www.geologicacarpathica.sk
GEOLOGICA CARPATHICA, APRIL 2010, 61, 2, 163—171 doi: 10.2478/v10096-010-0008-1
Sorption of heavy metal cations on rhyolitic and andesitic
bentonites from Central Slovakia
SLÁVKA ANDREJKOVIČOVÁ, MARTIN PENTRÁK, UBOŠ JANKOVIČ and PETER KOMADEL
Institute of Inorganic Chemistry, Slovak Academy of Sciences, Dúbravská cesta 9, SK-845 36 Bratislava, Slovak Republic;
slavka.andrejkovicova@savba.sk; martin.pentrak@savba.sk; lubos.jankovic@savba.sk; peter.komadel@savba.sk
(Manuscript received July 20, 2009; accepted in revised form October 2, 2009)
Abstract: The main purpose of this work was to determine adsorption characteristics of heavy metal cations on two Slovak
bentonites. Adsorption of Pb
2+
, Zn
2+
, Cu
2+
and Cd
2+
on Jelšový Potok (JP) and Lieskovec (L) bentonites was studied by the
batch equilibration technique using solutions of different concentrations. Higher smectite content (81 mass %) and higher
cation exchange capacity (CEC) (105 mmol M
+
/100 g) of JP bentonite cause higher adsorption of all heavy metals in
comparison with L bentonite. JP adsorbed heavy metals in the order Pb
2+
> > C d
2+
> Z n
2+
> C u
2+
while sorption on L was
slightly different, Pb
2+
> > Cd
2+
> Cu
2+
≥Zn
2+
. The Freundlich model of adsorption is more appropriate for adsorption of
Pb
2+
and Cd
2+
while lower uptake of Cu
2+
and Zn
2+
is better described by the Langmuir model. Negative
∆G° values
indicate that the adsorption process of all cations on both bentonites is feasible, spontaneous and exothermic. The decrease
in the d
001
spacings from 14.8—14.9
A
in natural dominantly Ca
2+
-saturated samples to 13.2—12.6
A
for both bentonites
saturated with four heavy metal cations shows the effect of less hydrated exchangeable cations on interlayer spacing.
Jelšový Potok bentonite of higher montmorillonite content and greater CEC is the more effective candidate for removal of
Pb
2+
, Zn
2+
, Cu
2+
and Cd
2+
from waste water than Lieskovec bentonite.
Key words: adsorption isotherm, bentonite, lead, zinc, copper, cadmium.
Introduction
With the current awareness of heavy metals as very toxic con-
taminants, extensive effort has been devoted to investigation
of their adsorption by solid surfaces as the most important
mechanism for controlling metal content in soil solutions and
natural waters (Helios Rybicka et al. 1995). Heavy metals are
dangerous environmental pollutants due to their toxicity and
strong tendency to concentrate in the environment and in food
chains (e.g. Farrah & Pickering 1977; Doner 1978; Jain et al.
2004; Bulut et al. 2006). One of the recent environmental ap-
plications of bentonites is connected with high-level nuclear
waste disposal in underground repositories using a multibarrier
system of two basic components, a host rock and an engineered
barrier made of metallic containers filled with radioactive waste
surrounded by bentonite blocks, as is documented by the results
of large scale experiments (e.g. Delay et al. 2007; Pacovský et
al. 2007; Stríček et al. 2009). Sorption of radioactive metal cat-
ions on the Land JP bentonites has been described recently in
detail by Galamboš et al. (2009a,b, 2010).
Cadmium, zinc, copper, nickel, lead, mercury and chromi-
um are often detected in industrial wastewaters originating
from metal plating, mining activities, smelting, battery manu-
facture, tanneries, petroleum refining, paint manufacture, pes-
ticides, pigment manufacture, printing and photographic
industries, etc. (Kadirvelu et al. 2001; Balistrieri & Blank
2008; Bhattacharyya & Gupta 2008).
The methods used for removal of these and other metals in-
clude chemical precipitation, ion exchange, solvent extraction,
reverse osmosis, adsorption, and others. Activated carbon is
highly effective in adsorbing heavy metals from wastewater
but high cost limits its use (Kumar 2006). Various soils were
also used for heavy metal adsorption (Fontes & Gomes 2003;
Veeresh et al. 2003). Viraraghavan & Kapoor (1994) noted
that the abundance and low cost of bentonite make it a strong
candidate as an adsorbent for the removal of heavy metals
from wastewaters or a retardant of the flow of leachates. Con-
sidering the favourable characteristics, adsorption of metal
ions and other substances on clays has received considerable
attention (e.g. Egozy 1980; Van Bladel et al. 1993; Yong et al.
2001; Abollino et al. 2003; Sezer et al. 2003; Egirani et al.
2005; Kaya & Ören 2005; Stathi et al. 2007; Abu-Eishah
2008; Zhang & Hou 2008; Al-Jlil & Alsewailem 2009).
Illite was shown to adsorb Cd
2+
(Tanabe 1981) and natural
bentonite to eliminate zinc from aqueous solution (Mellah &
Chegrouche 1997). Removal of Cr
3+
, Ni
2+
, Zn
2+
, Cu
2+
and
Cd
2+
by natural and Na-exchanged bentonites was also report-
ed (Álvarez-Ayuso & García-Sánchez 2003). Strawn et al.
(2004) discussed the adsorption of Cu
2+
by montmorillonite
and beidellite, whereas Lin & Juang (2002) used surfactant
modified montmorillonite for the removal of Cu
2+
and Zn
2+
.
Chantawong et al. (2001) studied adsorption of lead on a clay
consisting mainly of kaolinite and illite and confirmed that the
adsorption efficiency grows with increase in pH. Neverthe-
less, presence of other ions such as Cd
2+
, Cu
2+
, Ni
2+
, Zn
2+
and
Cr(VI) reduced the lead uptake from aqueous solutions due to
the fact that these ions bind strongly with organic matter
present in clay to form a complex.
Bentonite consisting of clay, silt and sand was used in zinc re-
moval (Mellah 1997). The sorption processes usually follow the
Langmuir isotherm (e.g. Veli & Alyüz 2007; Sari et al. 2007).
Adsorption capacities of 20 mg of Pb
2+
/g were achieved by
bentonite at pH 3.4 (Naseem 2001). Veli & Alyüz (2007) con-
firmed that pH is a significant factor in adsorption processes of
Å
Å
164
ANDREJKOVIČOVÁ, PENTRÁK, JANKOVIČ and KOMADEL
copper and zinc on bentonite causing electrostatic changes in
the solutions. Hydrated hydrogen ions are strongly competing
with other adsorbates. The highest removal efficiency in the
copper and zinc adsorption with natural clay was obtained at
pH > 6 (Veli & Alyüz 2007). Sari et al. (2007) calculated
changes in thermodynamic parameters, such as Gibbs free ener-
gy, enthalpy and entropy. The results showed that adsorptions
of Pb
2+
and Cr
3+
on clay were feasible, spontaneous and exo-
thermic processes in nature (Fan et al. 2009). Clays could be
modified to improve their sorption capacity (e.g. Bailey et al.
1999; Adebowale et al. 2005; Oyanedel-Craver & Smith 2006;
Bhattacharyya & Gupta 2006, 2008; Karamanis & Assimako-
poulos 2007; Eren & Afsin 2008; Guimar
a
es et al. 2009).
Two bentonite deposits in Central Slovakia underwent dif-
ferent alteration processes. The bentonite from Stará Krem-
nička-Jelšový potok (JP) developed from rhyolitic tuffs in a
lacustrine environment; the main component is an Al-rich
montmorillonite (Kraus et al. 1994). The deposit is located
in the SW part of the Kremnické Vrchy Mountains in the
Western Carpathians and belongs to the Jastrabá Formation.
Bentonite from Lieskovec has andesitic pyroclastics as par-
ent rocks and the main mineral is an iron-rich smectite (An-
drejkovičová et al. 2006). The Lieskovec (L) deposit,
belonging to Abčina Formation, is located in Zvolenská kot-
lina Basin, only about 25 km east of the JP deposit. In spite
of close occurrence of the two bentonites, they differ in min-
eralogical compositions. JP and L bentonites are well charac-
terized and commonly used. Andrejkovičová et al. (2008)
reported recently that the blend containing 65 mass % of
Na
+
-L and 35 mass % of Na
+
-JP bentonites meets all the re-
quirements on bentonites used in geosynthetic clay liners.
Smectite content in the blends was the dominant factor af-
fecting their properties. Osacký et al. (2009) investigated in
batch experiments stability of four bentonites and one K-ben-
tonite from Slovak deposits in the presence of iron to simu-
late possible reactions of a bentonite barrier in the contact
with a Fe container in a nuclear waste repository. The struc-
ture of illite-smectite deteriorated more than the structure of
smectites. These results support more extensive application
of JP and L bentonites in environmental protection. Howev-
er, systematic investigation of heavy metals sorption on
these materials has not been performed yet. The purpose of
this study was to determine adsorption characteristics of
Pb
2+
, Cd
2+
, Zn
2+
and Cu
2+
on these two Slovak bentonites.
The experiments were focused on simple adsorption of diva-
lent cations on dominantly Ca
2+
-bentonites to compare the
effect of mineralogical composition including smectite con-
tent on adsorption properties. Sorption mechanisms and the
effects of ionic strength are not discussed. It is hoped that the
results obtained will lead to further environmental applica-
tions of these clays.
Materials and methods
Materials
The commercially available JP and L bentonites were sup-
plied by Envigeo, Inc., Slovakia. Raw clays from the deposits
were dried and crushed to < 3 mm. A pendulum mill (Neuman
& Esser PM 05) connected to a cyclone classifier was used to
reduce the materials to particle size < 250
µm.
Methods
RockJock. Quantitative analysis of bentonites was per-
formed applying the RockJock program (Eberl 2003). The
program fits the sum of stored XRD patterns of pure standard
minerals (the calculated pattern) to the measured pattern by
varying fraction of each mineral using the Solver function Mi-
crosoft Excel to minimize the degree of fit parameter between
the calculated and measured pattern. Samples for analysis
were prepared by adding 0.111 g ZnO (internal standard) to
1.000 g sample. The mixture was ground in a McCrone mill
for 5 minutes with 4 ml of methanol then dried and sieved.
The diffraction patterns for RockJock analysis were collect-
ed in the 2-theta range from 4° to 65°, using steps of 0.02° 2
θ,
counting time 2 s per step, on a Philips PW 1710 diffractome-
ter with CuK
α (λ=1.54056
Ĺ
) radiation and a secondary
beam graphite monochromator PW 1752.
Powder X-ray diffraction (XRD) profiles of pressed powder
samples were collected with primary beam monochromatized
C
οKα (λ=1.78897
Ĺ
) radiation using a STOE Stadi P trans-
mission diffractometer (Stoe, Darmstadt, Germany) config-
ured with a linear position sensitive detector.
Films obtained on slides by evaporation of suspensions
were kept for 24 hours at 25 °C in a glass dessicator over satu-
rated magnesium nitrate solution at relative humidity of 53 %.
Oriented diffraction patterns were obtained with a Bruker D8
DISCOVER apparatus (Cu-K
α radiation, 40 kV/300 mA) us-
ing step of 0.05 2
θ and counting time 1 s per step.
Fourier transform infrared (FTIR) spectra were measured
in the 4000—400 cm
—1
region using KBr pressed-disk tech-
nique (1 mg of sample and 200 mg of KBr) on a Nicolet Ma-
gna 750 spectrometer with a DTGS detector and a KBr beam
splitter. Discs were heated in a furnace overnight at 150 °C to
minimize the amount of water adsorbed on KBr and the clay
samples.
Cation exchange capacity (CEC) was determined using
0.01 M solution of Cu
2+
triethylenetetramine [Cu Trien]
2+
pre-
pared according to Meier & Kahr (1999). 200 mg ( ± 0.5 mg)
of clay samples were added to 50 ml of distilled water and
10 ml solution of [Cu Trien]
2+
, then subjected to an ultrasonic
treatment for five minutes, filtered and concentration of Cu
2+
complex was determined in the filtrates by UV-VIS spectropho-
tometry (Cary 100, Varian) at 578 nm (Meier & Kahr 1999).
The amount of adsorbed [Cu Trien]
2+
was determined using
molar absorption coefficient
ε=0.245 mol
—1
· dm
3
· cm
—1
(Kauf-
hold & Dohrmann 2003) and the CEC values in milliequiva-
lents of cations per 100 grams of specimen were calculated.
ICP atomic emission spectroscopy. Amounts of adsorbed
Pb
2+
, Zn
2+
, Cu
2+
, Cd
2+
were calculated from the difference be-
tween their initial and final contents in solutions used for sorp-
tion experiments. They were obtained with a sequential,
radially viewed ICP atomic emission spectrometer Vista MPX
(VARIAN).
Morphology of the samples was studied by scanning elec-
tron microscopy using a CARL ZEISS-EVO 40 HV micro-
ã
Å
Å
165
SORPTION OF HEAVY METAL CATIONS ON SLOVAK BENTONITES
scope. Before the scanning process, all the samples were
coated with gold to enhance the electron conductivity. The
samples were examined by energy dispersive X-ray analysis
(EDX) with spectrometer QUANTAX 400 to analyse the
chemical composition in the samples sputtered with carbon.
Adsorption procedure: Adsorption experiments were car-
ried out using a batch method. 100 mg of each sample was
added to 10 ml of nitrate solutions of Pb
2+
, Zn
2+
, Cu
2+
and
Cd
2+
. Five different concentrations for every heavy metal cat-
ion solution were used: 0.01 M, 0.005 M, 0.0025 M,
0.00125 M and 0.0005 M. A 24-h contacting period was
found to be sufficient to achieve equilibrium sorption. The rel-
atively high pH of solutions with montmorillonites may in-
duce the precipitation of hydroxides of Pb
2+
, Cd
2+
, Cu
2+
and
Zn
2+
as was observed for example, by Barrer & Townsend
(1976) and Strawn & Sparks (1999). To avoid possible misin-
terpretation of exchange selectivity at higher cation concentra-
tions, HNO
3
was added to the stock solutions of these
elements to adjust pH to 6 ± 0.1. The theoretical equivalent ra-
tio [M
2+
]/[H
+
] in solution provided values close to 1
×10
—3
, at
which possible influence of H
+
ions on the exchange reaction
could be neglected. Stock solutions of Cu
2+
, Zn
2+
, Pb
2+
and
Cd
2+
nitrates were prepared with deionized water. All the
chemicals used were of analytical reagent grade and were ob-
tained from Aldrich.
The separation of the liquid from the solid phase was
achieved by centrifugation at 20,000 rpm for 30 min.
Amounts of adsorbed Pb
2+
, Zn
2+
, Cd
2+
and Cu
2+
were calcu-
lated from the difference between the heavy metal cation ini-
tially added into the system and that remaining in the solution
at the adsorption equilibration, as obtained by a ICP atomic
emission spectroscopy.
The distribution coefficient K
d
(dm
3
· kg
—1
) was calculated
using Eq. (1):
, (1)
where c
ad
(mol ·kg
—1
)
is the amount of the metal cation ad-
sorbed and c
eq
(mol ·dm
—3
) is the equilibrium concentration of
the metal cation in solution.
Change in Gibbs free energy
∆G° (kJ/mol) was calculated
according to Eq. (2) and K
d
obtained from Eq. (1):
∆G° = —RT lnK
d
, (2)
where R is the universal gas constant (8.314 J ·mol
—1
·K
—1
)
and T is temperature.
The sorption equilibrium data were applied to the equation
based on the Freundlich model:
c
ad
= K
F
·c
eq
1/n
, where K
F
(mol·kg
—1
) and n (kg ·dm
—3
) are
Freundlich constants related to adsorption capacity and ad-
sorption intensity, respectively.
K
F
and 1/n were determined from the intercept and slope of
linear plot of log c
ad
versus log c
eq,
respectively:
. (3)
Sorption capacity K
L
according to Langmuir model was cal-
culated from:
, (4)
graph
was plotted and K
L
(mol· kg
—1
) was cal-
culated by regression of linearized Langmuir isotherm accord-
ing to Eq. (5):
, (5)
where K
2
is an equilibrium constant dependent on sorption
energy.
Distribution coefficients were determined using a conven-
tional batch-equilibration technique in which 100 mg of clay
was contacted with 10 ml of solution and shaken at room tem-
perature for 24 hours.
Results and discussion
X-ray diffraction
XRD diffraction patterns of JP and L samples are shown in
Fig. 1. Smectite with prevailing Ca
2+
cation in the interlayer
space is the dominant mineral in both samples with basal
(001) reflection at 6.68 °2 theta (15.09
A
). The (060) reflec-
tion at 73.24 °2 theta (1.49
A
) shows that the main mineral in
both samples is a dioctahedral smectite (Brindley & Brown
1980). Identified admixtures include muscovite and kaolinite
in JP and quartz, muscovite, kaolinite, cristobalite and ortho-
clase in L (Fig. 1).
More information on the composition of the materials stud-
ied was obtained by RockJock analysis (Table 1). The main
differences are related to the smectite and non-clay mineral
contents, such as quartz and feldspars. Other non-clay miner-
als include microcrystalline forms of SiO
2
and possibly also
iron oxides/oxyhydroxides.
Fig. 1. XRD patterns of JP and L samples (S – smectite, K – kao-
linite, Q – quartz, Cr – cristobalite, O – orthoclase, M – mus-
covite/illite).
eq
F
ad
c
n
K
c
log
1
log
log
+
=
eq
ad
d
c
c
K =
eq
eq
L
ad
c
K
c
K
K
c
.
1
.
.
2
2
+
=
)
(
eq
ad
eq
c
f
c
c
=
2
.
1
K
K
K
c
c
c
L
L
eq
ad
eq
+
=
Å
Å
166
ANDREJKOVIČOVÁ, PENTRÁK, JANKOVIČ and KOMADEL
Fig. 2. Infrared spectra of JP and L samples.
Infrared spectroscopy
IR spectra provide further information on mineralogical
composition of the samples and chemistry of the dominating
smectite. The adsorption band near 3624 cm
—1
, assigned to
stretching vibrations of structural OH groups of dioctahedral
smectite (montmorillonite), appears in the spectra of both
samples (Fig. 2). The broad complex band near 1030 cm
—1
is
related to the stretching vibrations of Si-O groups while the
bands at 522 cm
—1
and 468 cm
—1
are attributed to Al-O-Si and
Si-O-Si bending vibrations, respectively (Farmer 1974). Cou-
pled Al-O and Si-O out of plane bending vibration at 626 cm
—1
also confirms dioctahedral smectite in the samples (Madejová
& Komadel 2001). The characteristic bands of kaolinite at
3698 cm
—1
and 693 cm
—1
(Farmer 1974) are clearly visible in
the spectrum of L.
The doublet of quartz at 797 and 779 cm
—1
is overlapped
with the band of microcrystalline SiO
2
in the spectrum of L
sample (Fig. 2). Central atoms in the octahedral sheets of
smectite affect the OH-bending bands in the 950—800 cm
—1
re-
gion (Fig. 2). The discrete and relatively intense peak at
915 cm
—1
corresponds to the AlAlOH bending vibrations of
smectite; the OH bending vibrations of kaolinite and illite/
muscovite can contribute to this absorption. A higher content
of Fe(III) in the octahedral sheets of L is confirmed by the
AlFeOH band near 871 cm
—1
. The AlMgOH vibration near
841 cm
—1
is not observed in the spectrum of L, thus proving a
Table 1: Mineralogical composition of samples as obtained from
the RockJock analysis.
Sample
JP L
mass %
smectite
81
45
kaolinite
1.5
5
muscovite/illite
5
8
quartz
– 13
feldspars
– 11
other non-clay minerals
12.5 18
relatively low Mg content in the octahedral sheets of the main
mineral of Lieskovec bentonite (Andrejkovičová et al. 2006).
The spectrum of JP (Fig. 2A) contains a clearly visible
AlMgOH bending band near 841 cm
—1
proving higher isomor-
phous substitution of Mg(II) for Al(III) in the octahedral
sheets of JP than in L and resulting in its higher layer charge.
Both XRD and IR results prove a higher smectite content and
lower amounts of accessory minerals in JP.
Adsorption of heavy metal cations
The process of adsorption on the samples is depicted in
Fig. 3, where the relationship between the amount of adsorbed
metal cation and their equilibrium concentrations in aqueous
solutions after adsorption is shown. All experiments were per-
formed in triplicates with standard deviations < 2 %. pH was
kept at 6 ± 0.1 to avoid precipitation of heavy metal cations at
higher pH and undesirable attack on clay layers by protons at
lower pH (Strawn & Sparks 1999). JP bentonite with higher
smectite content and cation exchange capacity (81 mass %
and 105 mmol/100 g, respectively) is a better adsorbent than
L bentonite (45 mass % and 51 mmol/100 g). The adsorption
curves of Pb
2+
for both bentonites confirmed the best sorption
results. The amount of adsorbed Pb
2+
for each equilibrium
concentration was higher than for the other metal cations, in
accord with its lowest ionic potential of 1.60 eV/nm. Both ben-
tonites adsorbed heavy metal cations in a similar order: JP –
Pb
2+
>>Cd
2+
>Zn
2+
>Cu
2+
and L – Pb
2+
>>Cd
2+
>Cu
2+
≥Zn
2+
(Fig. 3), correlating reasonably well in an opposite manner
with the ionic potentials of 1.60, 2.06, 2.70 and 2.74 eV/nm
for Pb
2+
, Cd
2+
, Zn
2+
and Cu
2+
, respectively.
Adsorption isotherms
Langmuir and Freundlich isotherm models were used to de-
termine correlation between the amounts of adsorbed heavy
metal cations on JP and L samples and their equilibrium con-
centrations in aqueous solution (Fig. 4). The graphs on the left
side of Fig. 4 (A, C, E, G) show transformed experimental
data to Freundlich model in logarithmic form expressed by
Eq. (3), where K
F
and 1/n constants were determined from the
intercepts and slopes of linear plots of logc
ad
versus logc
eq
.
The obtained equilibrium data also conformed to the linear
form of the Langmuir model (Eq. 5) and are displayed on the
right side of Fig. 4 (B, D, F, H). Langmuir adsorption con-
stants (K
L
) were calculated from the intercepts and slopes of
the linear plots of c
eq
/c
ad
vs. c
eq
.
The adsorption patterns of the metals on JP and L clays
were well fitted with both the Langmuir (R
2
= 0.91—0.99) and
Freundlich (R
2
= 0.86—0.99) models. The Freundlich isotherm
represents multi-layer unlimited adsorption. Sorption capacity
is achieved at a certain concentration and this phenomenon
describes the Langmuir isotherm. Based on the R
2
values, the
Freundlich model is better applicable for adsorption of Pb
2+
and Cd
2+
while lower uptake of Cu
2+
and Zn
2+
is better de-
scribed by the Langmuir model.
Table 2 shows calculated values for adsorption of Pb
2+
,
Cd
2+
, Zn
2+
, and Cu
2+
on both samples, including the distribu-
tion coefficient K
d
, Gibbs free energy change
∆G°, sorption
Absorbance
167
SORPTION OF HEAVY METAL CATIONS ON SLOVAK BENTONITES
capacities K
F
, K
L
and Freundlich coefficient (1/n). The distri-
bution coefficients K
d
increase while
∆G° values decrease
with decreasing concentration of heavy metal in solution for
all investigated cations (Table 2). The negative values of
∆G°
indicate that the adsorption process of Pb
2+
, Cd
2+
, Zn
2+
and
Cu
2+
and on JP and L is feasible, spontaneous and exothermic
in nature (Fan et al. 2009).
The most spontaneous processes are connected with adsorp-
tion of all heavy metals from the less concentrated solutions.
The
∆G° values are between —16.1 and —18.3 for JP and —14.6
and —16.8 kJ/mol for L. The most feasible is adsorption of
Pb
2+
on JP from 0.0005 M solution with the lowest
∆G° of
—18.3 kJ/mol. The values of Gibbs free energy change de-
crease in the order Pb
2+
>Cu
2+
>Cd
2+
≥Zn
2+
for both JP and L at
the lowest concentration of heavy metal cations of 0.0005 M.
The adsorption process from the most concentrated solutions
of heavy metals (0.01 M, Table 3) differs in some way com-
pared to adsorption from 0.0005 M solutions. L adsorbs the
cations with
∆G° in the order Pb
2+
> Cd
2+
> Cu
2+
> Zn
2+
. This is
in good accordance with the results of Helios Rybicka et al.
Fig. 3. Adsorption of Pb
2+
, Cd
2+
, Zn
2+
and Cu
2+
on JP and L bentonites as a function of equilibrium concentration of the metal cation.
Table 2: Pb
2+
, Cd
2+
, Zn
2+
and Cu
2+
adsorption parameters for JP and L samples, c is concentration of metal cations in initial solutions.
JP L
c/10
–3
(mol/dm
3
)
10
5
2.5
1.25 0.5
10
5
2.5
1.25 0.5
K
d
(Henry) (dm
3
/kg)
68.6 148.3 278.2 635.3
2178
47.6 70.2 135.0
338.2 1147
∆G° (kJ/mol)
–10.1 –11.9 –13.4 –15.4
–18.3
–9.2 –10.1 –11.7
–13.9 –16.8
K
L
(Langmuir) (mol/kg)
0.44 0.35
K
F
(Freundlich) (mol/kg)
3.15
1.87
Pb
1/n (kg/dm
3
) 0.39
0.37
K
d
(Henry) (dm
3
/kg)
44.1 88.5 187.7 341.7 911.2
24.5 47.4 87.4
182.1 507.4
∆G° (kJ/mol)
–9.0 –10.7 –12.5 –13.9 –16.2
–7.6 –9.2 –10.6
–12.4 –14.8
K
L
(Langmuir) (mol/kg)
0.33
0.21
K
F
(Freundlich) (mol/kg)
2.35
1.09
Cd
1/n (kg/dm
3
) 0.39
0.34
K
d
(Henry) (dm
3
/kg)
39.6 72.6 138.3 303.8 846.1
10.9 24.4 62.3
148.5 459.0
∆G° (kJ/mol)
–8.8 –10.2 –11.7 –13.6 –16.1
–5.7 –7.6 –9.8
–11.9 –14.6
K
L
(Langmuir) (mol/kg)
0.31
0.10
K
F
(Freundlich) (mol/kg)
1.91
0.28
Zn
1/n (kg/dm
3
) 0.39
0.20
K
d
(Henry) (dm
3
/kg)
33.5 59.0 85.0 310.5 1188
12.3 26.6 64.3
162.2 645.8
∆G° (kJ/mol)
–8.4 –9.7 –10.6 –13.7 –16.9
–6.0 –7.8 –9.9
–12.1 –15.4
K
L
(Langmuir) (mol/kg)
0.27
0.11
K
F
(Freundlich) (mol/kg)
1.07
0.30
Cu
1/n (kg/dm
3
) 0.31
0.19
168
ANDREJKOVIČOVÁ, PENTRÁK, JANKOVIČ and KOMADEL
Fig. 4. Linear plots of Freundlich (left) and Langmuir (right) isotherms of Pb
2+
, Cd
2+
, Zn
2+
and Cu
2+
adsorption on JP and L bentonites.
169
SORPTION OF HEAVY METAL CATIONS ON SLOVAK BENTONITES
(1995) obtained for Cheto montmorillonite and for JP in a
similar order except for the exchanged last two members:
Pb
2+
> Cd
2+
> Zn
2+
> Cu
2+
.
K
d
values were calculated according to the Henry isotherm
model (Eq. 1) confirming monolayer adsorption of metal
cations of low concentration in solution. As expected, K
d
values decrease in the same order as
∆G° values
Pb
2+
> Cu
2+
> Cd
2+
> Zn
2+
for 0.0005 M solutions and are high-
er for JP than for L. K
d
values for 0.01 M solutions also de-
cline in the same way as
∆G° values do in the order of
Pb
2+
> Cd
2+
> Zn
2+
> Cu
2+
and Pb
2+
> Cd
2+
> Cu
2+
> Zn
2+
for JP
and L, respectively (Table 2).
The Freundlich adsorption capacity K
F
is the highest for
Pb
2+
, with values of 3.15 and 1.87 mol/kg for JP and L, re-
spectively. L has comparable values of K
F
for Zn
2+
and Cu
2+
of 0.28 and 0.30 mol/kg, respectively. The K
F
values decrease
for JP in the order Pb
2+
> Cd
2+
> Zn
2+
> Cu
2+
while for L in or-
der Pb
2+
> Cd
2+
> Cu
2+
≥Zn
2+
. Moreover, the Freundlich coeffi-
cients 1/n are lower than 1 and indicate that the adsorption of
all heavy metal cations on JP and L is favourable under the
studied conditions.
Sorption capacity values (K
L
) provide information on
maximal sorption capacity of the materials in given condi-
tions. Both bentonites have maximum sorption capacities for
Pb
2+
, 0.44 mol/kg and 0.35 mol/kg for JP and L, respective-
ly. The lowest K
L
has L for Zn
2+
(0.10 mol/kg) and JP for
Cu
2+
(0.27 mol/kg). K
L
values decrease for JP and L in the
same order as K
F
values: Pb
2+
>Cd
2+
>Zn
2+
> Cu
2+
and
Pb
2+
>Cd
2+
>Cu
2+
≥ Zn
2+
, respectively.
Fig. 5. Oriented XRD patterns of JP and L samples: (a) native sam-
ples, (b) Cd
2+
-saturated, (c) Pb
2+
-saturated, (d) Zn
2+
-saturated and
(e) Cu
2+
-saturated, d
001
values in
A
.
JP L
c/10
–3
(mol/dm
3
)
10 5 2.5 1.25 0.5 10 5 2.5 1.25 0.5
Ca 1.85 0.57 0.95 1.16 1.32 1.80 0.86 –
–
0.43 0.58 0.63
M
as
s %
Pb
– 6.72 5.53 3.64 2.45 1.04 – 3.64 2.77 3.00 1.99 0.58
Table 3: EDX analysis for Ca and Pb in JP and L, c is concentration
of metal cations in initial solutions.
The type of cation in the interlayer space of a smectite influ-
ences its d
001
value. Figure 5 shows changes in interlayer dis-
tances after adsorption of individual heavy metals from the
0.01 M solutions for both clays. The d
001
values decrease pro-
gressively from ~ 14.9
A
obtained for native samples with
Ca
2+
as the prevailing cation to values in the 12.6—13.2
A
range for smectites with heavy metal cations which are less
hydrated than Ca
2+
. These numbers are in accord with the data
of Auboiroux et al. (1996) and Brigatti et al. (1995) for Pb
2+
and Zn
2+
-saturated Wyoming and Cheto montomorillonites,
respectively. The change in d
001
is a function ion exchange
also depending on solution concentration or heavy metal cat-
ion content available for ion exchange. No changes in d
001
val-
ues were observed after treatments with 0.0005 M solutions
because of an insufficient amount of heavy metal cation to get
adsorbed on montmorillonite and substitute substantial
amount of Ca
2+
in the interlayers, thus the final d
001
of 14.9
A
was the same as that obtained for the parent Ca
2+
-saturated
montmorillonites.
Bentonites after adsorption of heavy metal cations were
studied by EDX analysis to determine the contents of ad-
sorbed cations. Table 3 provides data on Pb and Ca in the
starting materials and samples treated in particular solutions.
Ca
2+
content in raw JP (1.85 %) was more than twice as high
as in L (0.86 %). With increasing Pb
2+
amount in solution, Pb
contents in treated bentonites increased and Ca
2+
decreased.
As expected, more Pb
2+
got adsorbed on JP than on L from so-
lutions of all concentrations. No Ca
2+
remained in L after Pb
2+
adsorption from 0.01 M and 0.005 M solutions; it was quanti-
tatively exchanged with Pb
2+
.
Conclusions
The adsorption process of Pb
2+
, Cd
2+
, Zn
2+
and Cu
2+
on both
bentonites is feasible, spontaneous and exothermic in nature.
Modified Langmuir and Freundlich equations are suitable to
characterize the adsorption of heavy metal cations on the stud-
ied bentonites. The adsorption patterns of metal cations on JP
and L bentonites are well fitted with the Langmuir and Freun-
dlich models, providing slightly better fits for the Langmuir
(R
2
= 0.91—0.99) than for the Freundlich (R
2
= 0.86—0.99) mod-
el. The greatest adsorption amounts are obtained on both ben-
tonites for Pb
2+
, followed by Cd
2+
and Zn
2+
or Cu
2+
. Higher
smectite content and higher cation exchange capacity are the
crucial factors affecting the bentonite adsorption properties
generally and causing the superior properties of the Jelšový
Potok bentonite in comparison with the Lieskovec bentonite.
Acknowledgments: The financial support of the Slovak Grant
Agency VEGA (Grant 2/0183/09) is much appreciated.
References
Abollino O., Aceto M., Malandrino M., Sarzanini C. & Mentasti E.
2003: Adsorption of heavy metals on Na-montmorillonite. Ef-
fect of pH and organic substances. Water Res. 37, 1619—1627.
Abu-Eishah S.I. 2008: Removal of Zn, Cd, and Pb ions from water by
Sarooj clay. Appl. Clay Sci. 42, 201—205.
Å
Å
Å
Å
170
ANDREJKOVIČOVÁ, PENTRÁK, JANKOVIČ and KOMADEL
Adebowale K.O., Unuabonah I.E. & Olu-Owolabi B.I. 2005: Adsorp-
tion of some heavy metal ions on sulfate- and phosphate-modi-
fied kaolin. Appl. Clay Sci. 29, 145—148.
Al-Jlil S.A. & Alsewailem F.D. 2009: Saudi Arabian clays for lead
removal in wastewater. Appl. Clay Sci. 42, 671—674.
Álvarez-Ayuso E. & García-Sánchez A. 2003: Removal of heavy
metals from wastewaters by natural and Na-exchanged bento-
nites. Clays Clay Miner. 51, 475—480.
Andrejkovičová S., Madejová J., Czímerová A., Galko I., Dohrmann
R. & Komadel P. 2006: Mineralogy and chemistry of Fe-rich
bentonite from Lieskovec deposit (Central Slovakia). Geol. Car-
pathica 57, 371—378.
Andrejkovičová S., Rocha F., Janotka I. & Komadel P. 2008: An in-
vestigation into the use of blends of two bentonites for geosyn-
thetic clay liners. Geotextiles Geomembranes 26, 436—445.
Auboiroux M., Baillif P., Touray J.C. & Bergaya F. 1996: Fixation of
Zn
2+
and Pb
2+
by a Ca-montmorillonite in brines and dilute solu-
tions: Preliminary results. Appl. Clay Sci. 11, 117—126.
Bailey S.E., Olin T.J., Bricka R.M. & Adrian D.D. 1999: A review of
potentially low-cost sorbents for heavy metals. Water Res. 33,
2469—2479.
Balistrieri L.S. & Blank R.G. 2008: Dissolved and labile concentra-
tions of Cd, Cu, Pb, and Zn in the South Fork Coeur d’Alene
River, Idaho: Comparisons among chemical equilibrium models
and implications for biotic ligand models. Appl. Geochem. 23,
3355—3371.
Barrer R.M. & Townsend R.P. 1976: Transition metal ion exchange
in zeolites. Part 1. Thermodynamics of exchange of hydrated
Mn, Co, Ni, Cu and Zn ions in ammonium mordenite. J. Chem.
Soc., Faraday Trans. 72, 661—673.
Bhattacharyya K.G. & Gupta S.S. 2006: Kaolinite, montmorillonite,
and their modified derivatives as adsorbents for removal of
Cu(II) from aqueous solution. Sep. Purif. Technol. 50, 388—397.
Bhattacharyya K.G. & Gupta S.S. 2008: Adsorption of a few heavy
metals on natural and modified kaolinite and montmorillonite: A
review. Adv. Colloid Interface Sci. 140, 114—131.
Brigatti M.F., Corradini F., Franchini G.C., Mazzoni S., Medici L. &
Poppi L. 1995: Interaction between montmorillonite and pollut-
ants from industrial waste-waters: exchange of Zn
2+
and Pb
2+
from aqueous solutions. Appl. Clay Sci. 9, 383—395.
Brindley G.W. & Brown G. 1980: Mineralogical society monograph
No. 5. Crystal structures of clay minerals and their X-ray identi-
fication. Miner. Soc. London, 171—173.
Bulut Y. & Baysal Z. 2006: Removal of Pb(II) from wastewater us-
ing wheat bran. J. Environ. Manage. 78, 107—113.
Chantawong V., Harvey N. & Bashkin V.N. 2001: Comparison of
heavy metal adsorptions by Thai kaolin and ball clay. Asian J.
Energy Environ. 1, 33—48.
Delay J., Vinsot A., Krieguer J.M., Rebours H. & Armand G. 2007:
Making of the underground scientific experimental programme
at the Meuse/Haute-Marne underground research laboratory,
North Eastern France. Physics and Chemistry of the Earth 32,
2—18.
Doner H.E. 1978: Chloride as a factor in mobilities of Ni(II), Cu(II)
and Cd(II). Soil Sci. Soc. Amer. J. 42, 882—885.
Eberl D.D. 2003: User’s guide to RockJock – a program for deter-
mining quantitative mineralogy from powder X-ray diffraction
data. U.S. Geol. Surv., Open-File Report, 03—78.
Egirani D.E., Baker A.R. & Andrews J.E. 2005: Copper and zinc re-
moval from aqueous solution by mixed mineral systems II. The
role of solution composition and aging. J. Colloid Interface Sci.
291, 326—333.
Egozy Y. 2008: Adsorption of cadmium and cobalt on montmorillo-
nite as a function of solution composition. Clays Clay Miner. 28,
311—318.
Eren E. & Afsin B. 2008: An investigation of Cu(II) adsorption by
raw and acid-activated bentonite: A combined potentiometric,
thermodynamic, XRD, IR, DTA study. J. Hazard. Mater. 151,
682—691.
Fan Q., Li Z., Zhao H., Jia Z., Xu J. & Wu W. 2009: Adsorption of
Pb(II) on palygorskite from aqueous solution: Effects of pH,
ionic strength and temperature. Appl. Clay Sci. 45, 111—116.
Farmer V.C. 1974: Layer silicates. In: Farmer V.C. (Ed.): Infrared
spectra of minerals. Miner. Soc. London, 331—363.
Farrah H. & Pickering W.F. 1977: Influence of clay-solute interaction
on aqueous heavy metal ion levels. Water, Air, Soil Pollut. 8,
189—197.
Fontes M.P.F. & Gomes P.C. 2003: Simultaneous competitive ad-
sorption of heavy metals by the mineral matrix of tropical soils.
Appl. Geochem. 18, 795—804.
Galamboš M., Kufčáková J. & Rajec P. 2009a: Sorption of stron-
tium on Slovak bentonites. J. Radioanal. Nucl. Chem. 281,
347—357.
Galamboš M., Kufčáková J. & Rajec P. 2009b: Adsorption of cesi-
um on domestic bentonites. J. Radioanal. Nucl. Chem. 281,
485—492.
Galamboš M., Kufčáková J., Rosskopfová O. & Rajec P. 2010: Ad-
sorption of cesium and strontium on natrified bentonites. J.
Radioanal. Nucl. Chem. 283, 803—813.
Guimar
a
es A. de M.F., Ciminelli V.S.T. & Vasconcelos W.L. 2009:
Smectite organofunctionalized with thiol groups for adsorption
of heavy metal ions. Appl. Clay Sci. 42, 410—414.
Helios Rybicka E.H., Calmano W. & Breeger A. 1995: Heavy metals
sorption/desorption on competing clay minerals; an experimen-
tal study. Appl. Clay Sci. 9, 369—381.
Jain C.K., Singhal D.C. & Sharma M.K. 2004: Adsorption of zinc on
bed sediment of river Hindon: adsorption models and kinetics. J.
Hazard. Mater. B 114, 231—239.
Kadirvelu K., Thamaraiselvi K. & Namasivayam C. 2001: Removal
of heavy metal from industrial wastewaters by adsorption onto
activated carbon prepared from an agricultural solid waste.
Bioresour. Technol. 76, 63—65.
Karamanis D. & Assimakopoulos P.A. 2007: Efficiency of alumi-
num-pillared montmorillonite on the removal of cesium and
copper from aqueous solutions. Water Res. 41, 1897—1906.
Kaufold S. & Dohrmann R. 2003: Beyond the methylene blue meth-
od: determination of the smectite content using the CuTriene
Method. Z. Angew. Geol. 2, 13—17.
Kaya A. & Ören A.H. 2005: Adsorption of zinc from aqueous solu-
tions to bentonite. J. Hazard. Mater. B 125, 183—189.
Kraus I., Šamajová E., Šucha V., Lexa J. & Hroncová Z. 1994: Di-
agenetic and hydrothermal alteration of volcanic rocks into clay
minerals and zeolites (Kremnické vrchy Mts., The Western Car-
pathians). Geol. Carpathica 45, 151—158.
Kumar U. 2006: Agricultural products and by-products as a low cost
adsorbent for heavy metal removal from water and wastewater:
A review. Sci. Res. Essay 1, 33—37.
Lin S.-H. & Juang R.-S. 2002: Heavy metal removal from water by
sorption using surfactant-modified montmorillonite. J. Hazard.
Mater. B 92, 315—326.
Madejová J. & Komadel P. 2001: Baseline studies of the clay miner-
als society source clays: infrared methods. Clays Clay Miner.
49, 410—432.
Meier L.P. & Kahr G. 1999: Determination of the Cation Exchange
Capacity (CEC) of Clay Minerals using the Complexes of
Copper(II) ion with Triethylenetetramine and Tetraethylenepen-
tamine. Clays Clay Miner. 47, 386—388.
Mellah A. & Chegrouche S. 1997: The removal of zinc from aqueous
solutions by natural bentonite. Water Res. 31, 621—629.
Naseem R. & Tahir S.S. 2001: Removal of Pb(II) from aqueous/acid-
ic solutions by using bentonite as an adsorbent. Water Res. 35,
3982—3986.
ã
171
SORPTION OF HEAVY METAL CATIONS ON SLOVAK BENTONITES
Osacký M., Honty M., Madejová J., Bakas T. & Šucha V. 2009: Ex-
perimental interactions of Slovak bentonites with metallic iron.
Geol. Carpathica 60, 6, 535—543.
Oyanedel-Craver V.A. & Smith J.A. 2006: Effect of quaternary am-
monium cation loading and pH on heavy metal sorption to Ca
bentonite and two organobentonites. J. Hazard. Mater. 137,
1102—1114.
Pacovský J., Svoboda J. & Zapletal L. 2007: Saturation development
in the bentonite barrier of the Mock-Up-CZ geotechnical experi-
ment. Physics and Chemistry of the Earth 32, 767—779.
Sari A., Tuzen M. & Soylak M. 2007: Adsorption of Pb(II) and
Cr(III) from aqueous solution on Celtek clay. J. Hazard. Mater.
144, 41—46.
Sezer G.A., Türkmeno
g
lu A.G. & Göktürk E.H. 2003: Mineralogical
and sorption characteristics of Ankara Clay as a landfill liner.
Appl. Geochem. 18, 711—717.
Stathi P., Litina K., Gournis D., Giannopoulos T.S. & Deligiannakis
Y. 2007: Physicochemical study of novel organoclays as heavy
metal ion adsorbents for environmental remediation. J. Colloid
Interface Sci. 316, 298—309.
Strawn D.G.1 & Sparks D.L. 1999: The use of XAFS to distinguish
between inner- and outer-sphere lead adsorption complexes on
montmorillonite. J. Colloid Interface Sci. 216, 257—269.
Strawn D.G., Palmer N.E., Furnare L.J., Goodell C., Amonette J.E. &
Kukkadapu R.K. 2004: Copper sorption mechanisms on smec-
tites. Clays Clay Miner. 52, 321—333.
Stríček I., Šucha V., Uhlík P., Madejová J. & Galko I. 2009: Mineral
stability of iron-rich bentonite in the Mock-Up-CZ experiment.
Geol. Carpathica 60, 5, 431—436.
Tanabe K. 1981: Solid acid and base catalysis. In: Anderson J.R. &
Boudart M. (Eds.): Catalysis science and technology. Springer,
New York, p. 231.
Van Bladel R., Halen H. & Cloos P. 1993: Calcium-zinc and calcium-
cadmium exchange in suspension of various types of clays. Clay
Miner. 28, 33—38.
Veeresh H., Tripathy S., Chaudhuri D., Hart B.R. & Powell M.A.
2003: Sorption and distribution of adsorbed metals in three soils
of India. Appl. Geochem. 18, 1723—1731.
Veli S. & Alyüz B. 2007: Adsorption of copper and zinc from aqueous
solutions by using natural clay. J. Hazard. Mater. 149, 226—233.
Viraraghavan T. & Kapoor A. 1994: Adsorption of mercury from
wastewater by bentonite. Appl. Clay Sci. 9, 31—49.
Zhang S.Q. & Hou W.G. 2008: Adsorption behavior of Pb(II) on
montmorillonite. Colloids and Surfaces A: Physicochem. Eng.
Aspects. 320, 92—97.
Yong R.N., Yaacob W.Z.W., Bentley S.P., Harris C. & Tan B.K.
2001: Partitioning of heavy metals on soil samples from column
tests. Engng. Geol. 60, 307—322.
ğ