Artículo en PDF
How to cite
Complete issue
More information about this article
Journal's homepage in redalyc.org
Sistema de Información Científica
Red de Revistas Científicas de América Latina y el Caribe, España y Portugal
Rev. Int. Contam. Ambie. 28 (3) 237-249, 2012
EVALUATION OF THE REMOVAL OF ARSENIC AND CADMIUM FROM AQUEOUS
SOLUTION USING NATURAL RHYOLITIC SEDIMENTS AND METALLURGICAL WASTES
Luis Gerardo MARTÍNEZ JARDINES
1
, Francisco MARTÍN ROMERO
1*
,
Margarita Eugenia GUTIÉRREZ RUIZ
2
and Águeda Elena CENICEROS GÓMEZ
3
1
Departamento de Geoquímica, Instituto de Geología, Universidad Nacional Autónoma de México
2
Grupo de Biogeoquímica Facultad de Química, Universidad Nacional Autónoma de México
3
Departamento de Química Analítica, Facultad de Química, Universidad Nacional Autónoma de México
*Autor responsable; fmrch@geologia.unam.mx
(Recibido octubre 2011, aceptado marzo 2012)
Key words: arsenic, cadmium, rhyolitic sediments, metallurgical waste, sorption isotherm
ABSTRACT
The use of natural materials abundant, ef±cient and inexpensive for use in stabilization
of contaminants is in development, so some sorbent materials for removal of Cd (II) on
aqueous solutions in the range of 10-100 mg/L and for As (III) and As (V) in the range
of 1-500 mg/L have been investigated. The sorbent materials studied are indigenous
rhyolitic sediments and metallurgical wastes from San Luis Potosi, Mexico. Mineral-
ogical analysis showed that rhyolitic sediments are characterized by the occurrence of
clay minerals, while the metallurgical wastes are characterized by Fe-bearing minerals
as ammoniojarosite, K-jarosite, hematite and goethite. The experimental results showed
that the rhyolitic sediments had high removal ef±ciency (94-99 %) for Cd (II); while
As (III) was barely removed (5-18 %) and As (V) was not retained by these natural
geological materials. By contrast, the removal of As (III) and As (V) by metallurgical
wastes had an ef±ciency of 88 and 77 %, respectively. However, these wastes were not
able to remove Cd (II). The experimental results were ±tted to the Linear, Langmuir,
and Freundlich isotherm models to obtain the characteristic parameters of each model.
The Linear model for As (III) on rhyolitic sediments, as well as the Langmuir model
for Cd (II) on rhyolitic sediments and As (III) and (V) on metallurgical wastes, were
found to well represent the measured sorption data.
Palabras clave: arsénico, cadmio, sedimentos riolíticos, residuo metalúrgico, isoterma de sorción
RESUMEN
El uso de materiales naturales, abundantes, e±cientes y de bajo costo para utilizarlos
en métodos de estabilización de contaminantes se encuentra en desarrollo por lo que
se investigó la e±ciencia de algunos materiales para la retención en solución acuosa
de Cd (II) en concentraciones de 10 a 100 mg/L y para As (III) y As (V) en concentra-
ciones de 1 a 500 mg/L. Los materiales evaluados son sedimentos riolíticos y residuos
metalúrgicos que provienen de San Luis Potosí, México. La composición mineralógica
de los sedimentos riolíticos indica la presencia de minerales de arcilla, mientras que
los residuos metalúrgicos se caracterizan por la presencia de minerales de hierro como
L.G. Martínez Jardines
et al.
238
son: jarosita, amoniojarosita, goetita y hematita. Los resultados de este estudio indican
que los sedimentos riolíticos son capaces de retener Cd (II) con una efciencia del 94-
99 %. Sin embargo, tienen una efciencia baja (5-18 %) para la retención de As (III) y
no tienen capacidad de retención de As (V). Los residuos metalúrgicos tienen mucha
capacidad de retención de As (V) y As (III) con 88 y 77 % de efciencia, respectiva-
mente; pero no tienen capacidad de retener Cd (II). Los resultados experimentales se
ajustaron con los modelos de isotermas de sorción Lineal (K
d
), Langmuir y Freundlich,
con el fn de obtener los parámetros característicos de cada modelo. Los datos expe-
rimentales mostraron un ajuste adecuado a los modelos Lineal K
d
para As (III) y de
Langmuir para el caso de Cd (II) en los sedimentos riolíticos y para As (III) y As (V)
en los residuos metalúrgicos.
INTRODUCTION
Arsenic and cadmium are toxic to plants and
animals because o± their mobility and solubility un-
der environmental conditions, and their a±fnity ±or
proteins, lipids and others cell components.
Arsenic (As) contamination o± soils and water is
mainly related to mining and metal processing, as
well as the use o± agricultural pesticides containing
this element (Lin and Puls 2000, Wang and Mulligan
2006, Gerente
et al
. 2010). Arsenic is ±ound in soils
and natural waters, mainly in the ±orm o± arsenate
(As (V)) and arsenite (As (III)). The distribution
between dissolved As (III) and As (V) is dependent
on redox potential and pH. Under oxidizing condi-
tions, the predominant specie is As (V), which exists
as deprotonated oxyanions o± arsenic acid (H
2
AsO
4
,
HAsO
4
2–
and AsO
4
3–
). Under reducing conditions, As
(III) is thermodynamically stable and exists in solu-
tion as arsenious acid, a neutral, uncharged molecule
(H
3
AsO
3
0
) that only ±orms deprotonated oxyanions
at pH > 9.2 (H
2
AsO
3
and HAsO
3
2–
) (Manning and
Goldberg 1997, Sadiq 1997).
The As (III) species are more toxic than As (V).
At the pH o± most natural soils and water, As (III) is
electrically neutral and consequently electrostatically
is not strongly adsorbed on most mineral sur±aces as
the negatively charged As (V) oxyanions (Brewster
1992).
Cadmium (Cd) contamination o± soils and water
is mainly associated with industrial activity usually
involving metal coatings, inadequate handling o±
nickel-cadmium batteries, use o± phosphate ±ertili-
zers, mining and metal processing, etc. (Mulligan
et al
. 2001, Tiller 1989). Under acidic conditions,
Cd is predominantly present as soluble Cd (II), it is
soluble in aqueous solution at pH values <10. So is
readily treatable by alkali precipitating as hydroxide.
Arsenic and Cd contamination o± soils and water
is a serious and recurring problem in many countries,
including México, which requires intervention in
order to decrease human health and environmental
risks. Di±±erent methods aimed at chemical stabiliza-
tion ±or the treatment o± soils and water contaminated
with As oxyanions and heavy metal cations have
been developed.
Arsenic stabilization processes include immobi-
lization by sorption in solid phases such as oxides
and hydroxides o± iron, aluminum and manganese
as well as organic matter (Wang and Mulligan 2006,
Daus
et al
. 2004).
For heavy metals, including Cd, sorption has
been reported on activated carbon (Mohan and Singh
2002), zeolites and clays (Celis
et al.
2000). Clays
can act as sorbents ±or heavy metal cations because
o± their negative charge, large sur±ace area associa-
ted with small particle size, high cation exchange
capacity, low cost and occurrence in most soils and
sediments.
In order to apply the stabilization methods in real
situations, the selected sorbent materials must be
abundant, inexpensive and e±fcient. Activated carbon
and activated alumina are sorbent materials whose
e±±ectiveness has been amply demonstrated (Trivedi
and Axe 2001, Mohan and Singh 2002); however,
their use is limited due to their high costs. Due to this
limitation, at the present time a number o± techno-
logies employing other natural sorbent materials are
being developed. Examples include: montmorillonite
±or Cd (II) (Mal±errari
et al
. 2007); kaolinite ±or Cd
(II), Cu (II), Pb (II) and Zn (II) (Srivastava
et al
.
2005); bentonite ±or Cu (II) and Zn (II) (Veli 2007);
basaltic volcanic slag ±or Zn (II) (Kwon
et al
. 2005);
industrial wastes ±or Cd (II), Cu (II), Pb (II) and Zn
(II) (Ciccu
et al.
2003) and limestone and calcareous
shales ±or As (V) (Romero
et al.
2004, 2011). Re-
cently, biosorbents derived ±rom ±ood industry waste
(Gerente
et al.
2010), agriculture (Mohan and Singh
2002) and biomass (Beesley and Marmiroli 2011)
have also been used as sorbent materials.
REMOVAL OF ARSENIC AND CADMIUM BY TWO NATURAL SORBENT MATERIALS
239
This research was conducted with the aim of
assessing As and Cd removal by natural rhyolitic
sediments and metallurgical wastes from San Luis
Potosi, Mexico, as sorbent materials. These materials
were characterized and sorption experiments were
carried out using batch-leaching tests to determine
their capacity to remove Cd (II) in the range of 10-
100 mg/L, and As (III) and (V) in the range of 1-500
mg/L from aqueous solutions.
MATERIALS AND METHODS
Sampling sorbent materials
The sampling of the sorbent materials was done in
an area near to San Luis Potosi City, Mexico. Seven
samples of rhyolitic materials (S1-S7) from different
natural deposits and one sample of metallurgical
wastes (MW) named ”jarosite” (a mixture of diffe-
rent compounds) stored in an impoundment of the
electrolytic zinc re±nery, were collected.
At each sampling site, ±ve sub-samples of 10 kg
were taken within a circle area of 25 m radius. The
±ve sub-samples were mixed and quartered to prepare
8 composite samples. All solid samples were stored
in hermetically sealed plastic bags to minimize dust
contamination during transport to the laboratory for
chemical and mineralogical analyses.
Characterization of sorbent materials
Samples were air dried, disaggregated and sieved
through a 2-mm mesh and homogenized. The pH
values were determined in solid suspensions (1:5
solid:water) following the EPA 9045 method (US
EPA 1995) using a Beckman model Φ 720 pH meter
and a glass electrode.
Also the pH value at the point of zero charge
(PZC or pHpzc), at which the surface charge beco-
mes neutral was calculated using the zeta potential
values determined with a Zeta-Meter System 3.0+.
The zeta potential was determined using a suspension
containing 100 mg of homogenized solid samples
and 500 mL of NaCl solution 0.01 M, as electrolyte
equilibrated for 30 minutes. The zeta potential was
measured at each pH values from 1 to 11, which were
previously adjusted stirring the suspension with 0.1
M NaOH or 0.1 M HCl. The PZC of the samples were
estimated by plotting the zeta potential as a function
of pH and determining the pH through interpolation
when zeta potential =0.
For chemical and mineralogical analysis, the
homogenized samples were pulverized in an agate
mortar to 200 mesh. The ±nely-milled samples were
analyzed using a Portable X-ray Fluorescence analy-
zer NITON XL3t, according to the EPA 6200 method
(US-EPA 2006). The mineral composition of samples
was determined by X-ray Diffraction (XRD) using
a Shimadzu XRD-6000 diffractometer, equipped
with a Ni ±lter, copper tube and monochromator. All
samples were analyzed at angular interval 2θ from
4 º to 70 º and a speed of 2 º/min, operated at 40 kV
voltage and with 30 mA applied potential.
Experimental sorption experiments
Sorption experiments on rhyolitic sediments
and metallurgical wastes were carried out in batch
systems, using solutions of As (III), As (V) or Cd
(II). The solid suspensions were prepared using 100
mL of As (III), As (V) or Cd (II) solutions and 5.0
g of sorbent materials. Sample suspensions were
equilibrated for 18 ² 0.25 hours in batch reactors at
room temperature (23 °C(+/- 1 ºC)) and continuously
shaken at 200 rpm, at constant ratio (solid:solution,
1:20). In the case of As (III), in order to prevent its
oxidation, the experiments were conducted in a ni-
trogenous atmosphere.
Solutions of As (III), As (V) and Cd (II) were
prepared from stock standard solutions of As
2
O
3
,
Na
2
HAsO
4
.7H
2
O and CdCl
2
.2.5 H
2
O, respectively.
The initial soluble concentration of these ions in the
tested suspensions was 50 mg/L. Batch experiments
were carried out in order to acquire data for cons-
tructing sorption isotherms for sorbent materials.
For rhyolitic sediments, concentrations of As (III)
and As (V) were used between 1 and 75 mg/L, and
for the metallurgical wastes, from 2.5 to 500 mg/L.
To test the sorption of Cd (II) in both materials ±ve
concentrations of this cation were used from 10 to
100 mg/L. To prevent hydroxide precipitation of Cd
(II), the Cd (II) solution used has pH =4.0.
After the equilibration time, the ±nal pH was
measured and the suspension was centrifuged and
±ltered through a 0.45
m
m membrane, transferred to
a vial, and stored in the dark at 4 ºC until chemical
analyses were performed. The soluble concentrations
of As (III), As (V) and Cd (II) were determined by
atomic absorption spectrometry (AAS) using a Va-
rian Spectra 110A model. Soluble concentrations of
As (III) and As (V) were determined using hydride
generation-AAS in samples with concentrations be-
low 3 mg/L (detection limit: 5
m
g/L). The samples
with concentrations >3 mg/L of As were measured
using AAS-³ame (detection limit: 3 mg/L). The Cd
(II) in all samples was also measured by AAS-³ame
(detection limit of 0.02 mg/L).
A quality control was implemented including
L.G. Martínez Jardines
et al.
240
blanks, spiked blanks and duplicate samples Preci-
sion and accuracy were better than 10 % for all the
analyzed elements.
RESULTS AND DISCUSSION
Mineralogical and chemical analyses of sorbent
materials
The mineralogical and chemical compositions of
rhyolitic sediments and the metallurgical wastes are
presented in
Tables I
,
II
, respectively. The results
show that rhyolitic sediments and metallurgical
wastes are chemically and mineralogically different.
XRD analysis showed that the mineralogical
composition of rhyolitic sediment samples (S1- S7)
is dominated by quartz (SiO
2
), plagioclase [(CaAl)
(SiAl)
4
O
8
], cristobalite (SiO
2
), feldspar (KSi
3
AlO
8
)
and clays, from the smectite and mica groups. Addi-
tionally, the S3 sample contains zeolite (CaAl
2
Si
7
O
18
6H
2
O).
The pH values of rhyolitic sediments varied
between 8.0 and 8.8. Chemically, these materials
are characterized by the absence of potentially toxic
elements (As, Cd, Cu, Pb, S) which were not detected
and relatively low total concentrations of Zn (40-82
mg/kg), Fe (1.39-1.96 %), Ca (0.76-1.33 %), K (1.06-
1.69 %) and Mn (0.01-0.03 %).
Mineralogical analyses of the metallurgical wastes
showed that it is mainly composed by ammonio-
jarosite [(NH
4
)
2
Fe
6
(SO
4
)
4
(OH)
12
] and K-jarosite
[KFe
3
(SO
4
)
2
(OH)
6
]. These metallurgical wastes con-
tain minor amounts of gunningite [(Zn,Mn)SO
4
·H
2
O],
gibsite, Al(OH)
3
, anglesite (PbSO
4
), quartz (SiO
2
),
sphalerite (ZnS), hematite (Fe
2
O
3
) and goethite
(FeOOH). This complex mineralogical composition
is due to the fact that these industrial wastes were
generated by the hydrometallurgical processing of Zn-
sul±de concentrates from different polymetallic mines
from Mexico, under strongly hot and acidic conditions.
The metallurgical wastes had low pH value (pH
= 2.7) and high total concentration of potentially
TABLE I.
MINERALOGICAL COMPOSITIONS OF RHYOLITIC SEDIMENTS (S1-S7) AND METALLURGICAL WASTES (MW)
Sample
Mineralogy
S1
Quartz: SiO
2
, plagioclase of intermediate composition:(CaAl)(SiAl)
4
O
8
, potassium feldspar sanidine: (NaK)(Si
3
Al)O
8
,
clays from the smectite group
S2
Quartz: SiO
2
, potassium feldspar sanidine: (NaK)(Si
3
Al)O
8
, clays from the smectite and mica groups
S3
Quartz: SiO
2
, potassium feldspar sanidine: (NaK)(Si
3
Al)O
8
, Cristobalite: SiO
2
, Zeolite type Heulandite: CaAl
2
Si
7
O
18
6H
2
O, clays from the smectite group
S4
Quartz: SiO
2
, potassium feldspar sanidine: (NaK)(Si
3
Al)O
8
, Cristobalite: SiO
2
, clays from the smectite and mica groups
S5
Quartz: SiO
2
, potassium feldspar sanidine: (NaK)(Si
3
Al)O
8
, Cristobalite: SiO
2
, clays from the smectite and mica groups
S6
Quartz: SiO
2
, potassium feldspar sanidine: (NaK)(Si
3
Al)O
8
, Cristobalite: SiO
2
, clays from the smectite and mica groups
S7
Quartz: SiO
2
, potassium feldspar sanidine: (NaK)(Si
3
Al)O
8
, Cristobalite: SiO
2
, clays from the smectite and mica groups
MW
Ammoniojarosite: (NH
4
)
2
Fe
6
(SO
4
)
4
(OH)
12
, K-Jarosite: KFe
3
(SO
4
)
2
(OH)
6
, Gunningite: (Zn,Mn)SO
4
·H2O, Gibsite: Al(OH)
3
,
Anglesite: PbSO
4,
Quartz: SiO
2
, Sphalerite: ZnS, Hematite: Fe
2
O
3
and Goethite: FeOOH
TABLE II.
SOME PHYSICOCHEMICAL PARAMETERS OF RHYOLITIC SEDIMENTS (S1-S7) AND META-
LLURGICAL WASTES (MW)
Sample
pH
PZC
As
Cd
Cu
Pb
Zn
Fe
Ca
K
Mn
S
mg/kg
%
S1
8.7
2.0
ND
ND
ND
ND
58.1
1.39
1.13
1.55
0.01
ND
S2
8
3.0
ND
ND
ND
ND
65.4
1.77
0.89
1.5
0.03
ND
S3
8.1
2.6
ND
ND
ND
ND
39.7
1.59
0.94
1.06
0.03
ND
S4
8.3
2.6
ND
ND
ND
ND
76.5
1.78
0.76
1.4
0.02
ND
S5
8.8
2.0
ND
ND
ND
ND
74.3
1.66
0.79
1.57
0.03
ND
S6
8.7
2.0
ND
ND
ND
ND
82
1.96
0.76
1.24
0.02
ND
S7
8.3
2.8
ND
ND
ND
ND
73.3
1.77
1.33
1.69
0.03
ND
MW
2.7
1.7
1968
2194
8347
10 939
72 332
48.34
0.48
ND
2.41
6.33
Detection Limit
4
100
42.7
3.7
35
0.01
0.5
0.02
0.01
2.3
Note: ND = Not detected
REMOVAL OF ARSENIC AND CADMIUM BY TWO NATURAL SORBENT MATERIALS
241
toxic elements, mainly As (1 968 mg/kg), Cd (2 194
mg/kg), Cu (8 347 mg/kg), Pb (10 939 mg/kg) and
Zn (72 332 mg/kg). It is important to emphasize
that total concentrations of Fe (48.3 %), S (6.3 %)
and Mn (2.4 %) are consistent with its mineralogy
composition (
Table I
).
Point of zero charge (PZC) values
The PZC of rhyolitic sediments varied between 2
and 3 (
Table II
) which are comparable to the values
reported for quartz and clays (Appelo and Postma
2005, Manning and Goldberg 1997). As mentioned
previously, the rhyolitic sediments contain quartz and
clay minerals from the smectite and mica groups,
which were identi±ed by DRX (
Table I
).
The PZC of the metallurgical wastes (“jarosite”) is
around 2.7. This low value is probably related to the pre-
sence of ammoniojarosite ((NH
4
)
2
Fe
6
(SO
4
)
4
(OH)
12
)
and K-jarosite (KFe
3
(SO
4
)
2
(OH)
6
), which, as pre-
viously noted, are the predominant minerals identi±ed
by XRD in these wastes. This PZC value is similar to
other reported values for jarosite minerals, which vary
from 1.8 to 3.9 (Sadowski
et al
. 2001).
Removal of Cd (II), As (III) and As (V) by rhyolitic
sediments and metallurgical wastes
The removal ef±ciency of the sorbent materials was
determined in suspensions prepared with solutions of
50 mg/L of Cd (II), As (III) and As (V). The initial
pH of the suspensions of rhyolitic sediments varied
between 8.0 and 8.7; while the suspensions of the
metallurgical wastes presented a constant pH = 2.7,
as expected by the acid behavior of jarosite minerals.
The percentage removal was estimated using
Equation 1. The experimental results are shown in
Table III.
% Removal =
x
100
C
i
–C
f
C
f
(1)
where C
i
, is the initial concentration (mg/L) of Cd
(II), As (III) and As (V), before the interaction with
the studying sorbent material.
C
f
, is the ±nal concentration (mg/L) of Cd (II), As
(III) and As (V) after 18 h interaction with the sorbent
materials.
The results indicate that the equilibrium pH de-
pends on the sorbent material. In rhyolitic sediments
the values varied from 6.6 to 8.8, and in the meta-
llurgical wastes the pH was 3.9 for As (III) and 4.1
for As (V) and Cd (II) (
Table III
).
Rhyolitic sediments
The ef±ciency of the rhyolitic sediments to remo-
ve Cd (II) was very high (94-99 %), while As (III)
was barely removed (5-18 %) and As (V) was not
retained by this natural geological material.
The high ef±ciency of rhyolitic sediments for re-
moval Cd (II) appears to be related to the presence of
smectite, mica and zeolite, which, under equilibrium
pH (pH = 6.6-8.0) reached during tests (
Table III
),
must be negatively charged because the PZC values
determined in these materials were low (PZC = 2-3),
favoring, in this way, the retention of metal cations
such as Cd (II), which has been widely reported in
the literature (Celis
et al
. 2000, Srivastava
et al
. 2005,
Rao
et al
. 2006, Malferrari
et al
. 2007, Veli 2007,
Panuccio
et al
. 2009).
The inef±ciency of the rhyolitic sediments to
remove As (V) is likely related to the fact that at the
pH of equilibrium (6.6-8.0) this element exists as
deprotonated oxyanion of arsenic acid (HAsO
4
2–
),
because the pK
a
values for arsenic acid are pK
3
a
= 2.3,
TABLE III.
REMOVAL OF CD (II), AS (III) AND AS (V) FROM AQUEOUS SOLUTIONS BY NATURAL
RHYOLITIC SEDIMENTS (S1-S7) AND METALLURGICAL WASTES (MW)
Samples
pH
(1)
Cd (II)
pH
(2a)
As (III)
pH
(2b)
As (V)
pH
(2c)
mg/L
%
mg/L
%
mg/L
%
S1
8.7
46.75
93.5
7.3
6.65
13.3
8.0
ND
0
8.5
S2
8.0
47.8
95.6
6.6
8.9
17.8
6.9
ND
0
7.2
S3
8.1
49.4
98.8
7.6
5.5
11
7.8
ND
0
7.9
S4
8.3
48.25
96.5
7.7
4.7
9.4
8.6
ND
0
8.9
S5
8.8
47.65
95.3
7.8
5.05
10.1
8.8
ND
0
8.9
S6
8.7
49.05
98.1
7.4
2.65
5.3
8.7
ND
0
8.9
S7
8.3
49.25
98.5
8.0
3.05
6.1
8.8
ND
0
9.0
MR
2.7
NDL
0
4.1
38.35
76.7
3.9
43.75
87.5
4.1
pH (1) = initial solution pH;
pH(2a), pH(2b) , pH(2c) = equilibrium pH values after interaction with testing
sorbent materials;
ND = Not detected. Detection limit Cd(II) = 0.02 mg/L, As (III) , As (V) = 0.005 mg/L
L.G. Martínez Jardines
et al.
242
pK
2
a
= 6.8, pK
1
a
= 11.6. At equilibrium pH, the clay
surfaces are negatively charged, so the oxyanions of
As (V) are repelled.
The removal of As (III) by rhyolitic sediments
varied from 5 to 18 %. The As (III) retention is
probably due to the fact that it is present as non-
charged arsenious acid (H
3
AsO
3
0
). This chemical
specie is a pyramidal molecule consisting of three
hydroxyl groups bonded to As (Bickmore
et al
.
2006) with a radius of 4.16 Å (Kim
et al
. 2004), a
bond distance As-O of 1.79 Å and an angle of 100º
31´ (Iakovleva 2003). The neutral arsenious acid
exists in most of the pH range since its pK
a
is very
high (pK
3
a
= 9.2, pK
2
a
= 12.1, pK
1
a
= 13.1), and can
approach the negative surfaces of clay minerals
forming inner complexes with metals mainly with
Al, since the greater reactivity of Al(OH)
3
toward
As (III) (Manning and Goldberg 1997).
Arsenite adsorption on clay minerals as kaolinite,
illite and montmorillonite had been reported by Gold-
berg (2002). The adsorption increases with increasing
solution pH up to a maximum near 7 to 8.5 (Goldberg
2000), that is the pH range of the experiments carried
out in this work.
Metallurgical wastes
The major minerals in metallurgical wastes are
ammoniojarosite and K-jarosite which, under equili-
brium pH (pH = 3.1-4.9) reached during tests (
Table
III
), must be negatively charged because the PZC
determined in these materials was low (PZC = 2.7),
favoring, in this way, the retention of metal cations
and disfavoring the removal of negatively charged
arsenic oxyanions. However, the experimental results
showed that the removal As (III) and As (V) by me-
tallurgical wastes was very high with an efFciency
of 88 and 77 %, respectively.
A possible explanation about the high efFciency
for removal As (III) and As (V) by metallurgical
wastes is the presence of ±e-oxyhydroxides be-
cause of their high afFnity for As (V) and As (III)
(Manning and Goldberg 1997, Gimenez
et al
.
2007, Villalobos and Antelo 2011). Mineralogical
analysis indicates that the metallurgical wastes
contain minor amounts of hematite (±e
2
O
3
) and
goethite (±eOOH). Additionally, it is possible
that ±e-oxyhydroxides could be formed under the
equilibrium pH (pH = 3.1-4.9) reached during tests.
Jarosite is stable at pH <3.0 (Babcan 1971; Baron
and Palmer 1996). However, as pH increases jarosi-
te is transformed to ±e-oxyhydroxides, represented
as ±e(OH)
3
. The reported reaction between these
two phases is:
K±e
3
(SO
4
)
2
(OH)
6
+ 3H
2
O = 3±e(OH)
3
+ K
+
+
2SO
4
2-
+ 3H
+
Another phenomenon that may be related with As
(V) removal by metallurgical wastes is the arsenate/
sulfate substitution in jarosite minerals identiFed
in these wastes. Some authors have reported the
arsenate-for-sulfate substitution in jarosite-group
minerals as an important control on arsenic mobility
(±oster 1998, Asta
et al
. 2009).
Sorption Isotherms
The sorption isotherm is of great importance in the
design of sorption systems. The sorption equilibrium
is described by the sorption isotherms that are charac-
terized by certain constants whose values express the
surface properties and afFnity of the sorbent. In order
to investigate the sorption isotherm, three equilibrium
models were analyzed: Linear, Langmuir and ±reun-
dlich isotherms, described by Equations 2, 3 and 4:
Linear Model
C
s
= K
d
C
e
(2)
where “C
s
”, is the amount of sorbed ion per unit
weight of solid (mmol/kg); “C
e
”, is the equilibrium
concentration on solution (mmol/L) and “K
d
”, is
the linear distribution coefFcient (L/kg). The Linear
isotherm is constructed by plotting “C
s
” against “C
e
”;
if the data lie on a straight line then the linear model
can be considered appropriate.
Q
max
*
K
*
C
e
1+K
*
C
e
C
s
=
Langmuir Model
(3)
“K” is the Langmuir bonding energy coefFcient and
“Q
max
(mg/kg) is the adsorption maximum capacity.
The Langmuir isotherm is constructed by plotting
“C
e
/C
s
” against “Ce”; if the data lie on a straight
line then the Langmuir model can be considered
appropriate. Using least squares linear regression,
the parameters “Q
max
” (1/slope) and “K” (slope/
intercept), may be found.
C
s
= K
f
C
e
1/n
Freundlich Model
(4)
“K
f
” is the ±reundlich distribution coefFcient related
to the total sorption capacity of the solid (L/kg) and
“n” is a constant relating to adsorption intensity.
When n = 1, the ±reundlich isotherm simpliFed to
the Linear isotherm (Equation 2). If data conform to
the ±reundlich model (Equation 4), a plot of “ln (C
e
)”
versus “ln (C
s
)” should result in a straight line. Using
a least squares linear regression, the parameters “1/n”
(slope) and “K
f
” (intercept) may be found.
REMOVAL OF ARSENIC AND CADMIUM BY TWO NATURAL SORBENT MATERIALS
243
Sorption isotherms of Cd (II) on rhyolitic sedi-
ments
Sorption isotherms of Cd (II) on rhyolitic sedi-
ments exhibited similarities in shape and the amount
sorbed (
Fig. 1A
). According to the classi±cation
proposed by Giles (1960), these isotherms may
be associated with group “L” and subgroup “2”,
indicating that the sorption sites of the theoretical
monolayer have been saturated.
The experimental results were ±tted to the Li-
near, Langmuir, and Freundlich isotherm models to
obtain the characteristic parameters of each model
(
Table IV
). Based on the coef±cients of determina-
tion (r
2
), the three isotherm models appear to pro-
duce a reasonable adjustment for the sorption of Cd
(II) on rhyolitic sediments. However, the Langmuir
model yielded the highest coef±cients of determina-
tion (r
2
= 0.9945 ² 0.0075), while for the Freundlich
and Linear models the r
2
values were 0.968 ² 0.020
and 0.873 ²0.030, respectively. Figure 1B shows
plots comparing the adjusted Langmuir isotherm
with the experimental data. The equation shows an
excellent ±t with the experimental data suggesting
that the results could be explained by this model.
There has been reported that clays can adsorb heavy
metals via ion exchange reactions and by formation
of inner-sphere complexes through Si-O and Al-O
groups at the clay particle edge (Celis
et al
. 2000).
Results shown in
Table IV
indicate that rhyo-
litic sediments have a maximum sorption capacity
(Q
max
) ranging between 18.5 and 26.0 mmol/kg and
sorption energy values (K) ranging between 41.1
and 86.8. The S3 has the greatest sorption capacity
for Cd (II) (Q
max
= 26 mmol/kg), most likely due
to the presence of zeolite which was identi±ed by
XRD, in addition to the clay minerals identi±ed in
all the other rhyolitic sediment samples (
Table I
).
It is important to mention, that values reported for
maximum sorption capacity on natural pure clays
are 170 mmol/kg for sepiolite and 250 mmol/kg for
montmorillonite (Celis
et al
. 2000).
Sorption isotherms of As (III) on rhyolitic sedi-
ments
Sorption isotherms of As (III) on rhyolitic sedi-
ments exhibited similarities in shape and amount
sorbed (
Fig. 2A
). According to the classi±cation
proposed by Giles (1960), these isotherms can be as-
sociated with the group “C” subgroup “1”, indicating
that sorption sites have not been saturated.
The experimental results were ±tted to the Li-
neal, Langmuir, and Freundlich isotherm models to
obtain the characteristic parameters of each model
(
Table IV
). Based on the coef±cients of determi-
nation (r
2
), only the Linear isotherm and Freundlich
appear to produce a reasonable model for sorption of
As (III) by rhyolitic sediments. However, the highest
coef±cients of determination (r
2
= 0.959 ² 0.044)
were obtained with the Linear model, while with
the Freundlich and Langmuir models, the r
2
values
were 0.939 ² 0.028 and 0.276 ²0.175, respectively.
Figure 2B
shows plots comparing the adjusted Linear
isotherm with the experimental data, demonstrating
an excellent ±t and suggesting that experimental data
could be explained by this model.
Results shown in Table IV indicate that the rhyo-
litic sediments have a K
d
ranging between 1.1 and
4.5 L/kg. Sample S2 is the rhyolitic material with
the best properties for As (III) retention. The ability
of rhyolitic sediments for As (III) sorption is related
with the presence of clay minerals, identi±ed in
these natural geological materials. In the literature
is reported that As (III) adsorption by clay minerals
is minimal at low pH and increases with increasing
pH (Wang and Mulligan 2006).
Sorption isotherms for As (III) and As (V) on
metallurgical wastes
Sorption isotherms of As (III) and As (V) on me-
tallurgical wastes exhibited similarities in shape and
amount retained (
Figures 3A1
and
3B1
). According
to the classi±cation proposed by Giles (1960), these
isotherms may be associated with group “L” and
subgroup “2”, indicating that sorption sites of the
theoretical monolayer have been saturated.
The experimental results were ±tted to the same
models used in the previous cases to obtain the
characteristic parameters of each one (
Table IV
).
Based on the coef±cients of determination (r
2
),
only the Langmuir model reasonably simulates
the As (III) sorption by metallurgical wastes. The
highest coef±cients of determination (r
2
= 0.972)
were obtained with this model, while with the Linear
and Freundlich models, the r
2
values were 0.414 and
0.05, respectively. In
Figure 3A2
the adjustment for
the Langmuir isotherm with the experimental data is
compared. The equation shows an excellent ±t with
the experimental data for the Langmuir isotherm.
The r
2
values for the sorption of As (V) by me-
tallurgical wastes were also calculated applying
the Langmuir, Freundlich and Linear models. The
values obtained were 0.995, 0.898 and 0.55, res-
pectively. These results suggest that both Langmuir
and Freundlich isotherm models could explain the
As (V) sorption by metallurgical wastes. However,
Figure 3B2
shows an excellent ±t with the experi-
mental data for the Langmuir isotherm.
Results shown in
Table IV
indicate that the me-
tallurgical wastes have a maximum sorption capacity
L.G. Martínez Jardines
et al.
244
(a)
(S1)
(S2)
(S3)
(S4)
(S7)
(S5)
(S6)
(b)
8
4
0
0.04
C
e
, mmol/L
24
20
16
12
8
4
0
0
0.04
0.08
C
e
, mmol/L
C
s
, mmol/kg
0.12
0.16
S1
S2
S3
S4
20
16
12
8
4
0
0
0.04
0.08
C
e
, mmol/L
C
s
, mmol/kg
0.12
0.16
S5
S6
S7
20
16
12
8
4
0
0
0.04
0.08
C
e
, mmol/L
C
s
, mmol/kg
0.12
0.2
0.16
20
16
12
8
4
0
0
0.04
0.08
C
e
, mmol/L
C
s
, mmol/kg
0.12
0.2
0.16
20
16
12
8
4
0
0
0.04
0.08
C
e
, mmol/L
C
s
, mmol/kg
0.12
0.16
20
16
12
8
4
0
0
0.04
0.08
C
e
, mmol/L
C
s
, mmol/kg
0.12
0.16
20
16
12
8
4
0
0
0.04
0.08
C
e
, mmol/L
C
s
, mmol/kg
0.12
0.16
25
20
15
10
5
0
0
0.02
0.04
C
e
, mmol/L
C
s
, mmol/kg
0.06
0.08
20
16
12
0
C
s
, mmol/kg
0.12
0.08
Fig. 1.
Cd (II) removal from aqueous solutions by rhyolitic sediments: (A) Plot of measured Cd (II) sorption per unit weight of solid (C
s
,
mmol/kg) against Cd (II) concentration in the equilibrium solution (C
e
, mmol/L). (B) Theoretical Langmuir isotherms (
) for the
Cd (II) sorption and experimental data (
)
REMOVAL OF ARSENIC AND CADMIUM BY TWO NATURAL SORBENT MATERIALS
245
(Q
max
) of 13.6 mmol/kg for As (III) and 18 mmol/
kg for As (V) and energy values for sorption (K) of
21 y 31.8, respectively. The ability of metallurgical
wastes for sorption of As (III) and As (V) is due to
the presence of the jarosite and Fe-oxyhydroxides
minerals because of their high af±nity for these ar-
senic oxyanions. Pierce and Moore (1982) reported
that As (V) is preferentially sorbed to Fe hydroxides
between pH 4 and 7 with an optimal adsorption pH
of about 4, whereas As (III) is preferentially sorbed
onto Fe hydroxides between pH 7 and 10 with an op-
timal adsorption pH of about 7. Spectroscopic studies
generally agreed that both As (III) and As (V) are
speci±cally sorbed, forming inner-sphere complexes
(Wang and Mulligan 2006). Goldberg and Johnston
(2001) reported that amorphous Fe oxides has a
maximum sorption capacity of 200 mmol/kg for As.
CONCLUSIONS
The results of this study indicate that the rhyolitic
sediments had high removal ef±ciency (94-99 %)
for Cd (II) in concentrations between 10-100 mg/L.
However, these materials presented low removal
ef±ciency (5-18 %) for As (III) in concentrations
between 1-500 mg/L and they were not able to
remove As (V) in the same concentrations range.
Experimental data for the sorption of Cd (II) and As
(III) were adjusted appropriately to the Langmuir
model (r
2
= 0.9945 ² 0.0075) and the K
d
Lineal model
(r
2
= 0.959 ² 0.044), respectively. The ef±ciency of
rhyolitic sediments for the removal of Cd (II) from
aqueous solutions is most likely due to the presence
of smectite, mica and zeolite. At the equilibrium
pH (pH = 6.6-8.0) they were negatively charged,
favoring the retention of metal cations such as Cd
(II). The As (III) removal ef±ciency by these natural
geological materials may be explained because at
the equilibrium pH, the predominant species is an
uncharged molecule (H
3
AsO
3
0
) which can approach
to the negative surfaces of clay minerals and diffuse
into the interlayer space of clays.
Metallurgical wastes had high removal ability for
As (III) and As (V) in concentrations between 1-500
mg/L from aqueous solutions with an ef±ciency of 88
TABLE IV.
COEFFICIENT OF DETERMINATION AND ISOTHERM PARAMETERS OF RHYOLITIC SEDI-
MENTS (S1-S7) AND METALLURGICAL WASTES (MW) DEDUCED FROM THE APPLICATION
OF LINEAR, LANGMUIR AND FREUNDLICH EQUATIONS
Sample
Ion
Sorption Isotherm Model
Linear
Langmuir
Freundlich
r
2
K
d
r
2
K
Q
max
,
r
2
K
f
n
L/kg
mmol/kg
L//kg
S1
Cd (II)
0.849
85.2
0.978
58.5
18.5
0.972
32.9
3
S2
0.921
107.1
0.994
48.2
21
0.978
42.1
2.7
S3
0.858
361.4
0.998
78
26
0.967
154.2
1.6
S4
0.845
101.3
0.996
42.7
20.5
0.923
49.1
2.2
S5
0.856
84.4
0.997
55.1
18.9
0.983
38.6
2.6
S6
0.906
174.6
0.999
41.1
22.1
0.979
61.4
2
S7
0.876
171.9
0.9998
86.8
20
0.972
49.7
2.5
Average
0.873
0.9945
0.968
SD
0.030
0.0075
0.020
S1
As (III)
0.97
3.2
0.241
NA
0.948
3.1
0.9
S2
0.985
4.5
0.138
0.919
3.6
1.0
S3
0.992
2.4
0.208
0.943
2
1.2
S4
0.989
2
0.366
0.983
1.8
1.2
S5
0.975
2.5
0.46
0.918
4.2
0.6
S6
0.871
1.1
0.017
0.902
1.7
0.9
S7
0.93
1.3
0.501
0.961
1.1
1.4
Average
0.959
0.276
0.939
SD
0.044
0.175
0.028
MR
As (III)
0.414
NA
0.972
21
13.6
0.05
NA
MR
As (V)
0.55
NA
0.995
30.8
18.0
0.898
20.1
3.9
NA: Not applied
L.G. Martínez Jardines
et al.
246
(a)
(b)
2
1.6
1.2
0.8
0.4
0
0
0.2
0.4
C
e
, mmol/L
C
s
, mmol/kg
0.6
0.8
S1
S3
S4
4
3
2
1
0
0
0.2
0.4
C
e
, mmol/L
C
s
, mmol/kg
0.6
0.8
1
S2
2.5
2
1.5
1
0.5
0
0
0.2
0.4
C
e
, mmol/L
C
s
, mmol/kg
0.8
0.6
2
1.6
1.2
0.8
0.4
0
0
0.2
C
e
, mmol/L
C
s
, mmol/kg
0.4
0.8
0.6
4
3
2
1
0
0
0.2
0.4
C
e
, mmol/L
C
s
, mmol/kg
0.8
0.6
1
1.6
1.2
0.8
0.4
0
0
0.2
C
e
, mmol/L
C
s
, mmol/kg
0.4
0.6
1
0.8
0.6
0.4
0.2
0
0
0.2
0.08
C
e
, mmol/L
C
s
, mmol/kg
0.6
1
0.8
2
1.6
1.2
0.8
0.4
0
0
0.02
0.04
C
e
, mmol/L
C
s
, mmol/kg
0.06
0.08
1.6
1.2
0.8
0.4
0
0
0.2
0.4
C
e
, mmol/L
C
s
, mmol/kg
0.6
0.8
S5
S6
S7
0.4
0.2
0
0.2
C
e
, mmol/L
1
0.8
0.6
0
C
s
, mmol/kg
0.8
0.6
0.4
(S1)
(S2)
(S3)
(S4)
(S7)
(S5)
(S6)
Fig. 2.
As (III) removal from aqueous solutions by rhyolitic sediments: (A) Plot of measured As (III) sorption per unit weight of solid
(C
s
, mmol/kg) against As (III) concentration in the equilibrium solution (C
e
, mmol/L).
(B) Theoretical Linear isotherms (
)
for the As (III) sorption and experimental data (
)
REMOVAL OF ARSENIC AND CADMIUM BY TWO NATURAL SORBENT MATERIALS
247
and 77 %, respectively. However, these wastes were
not able to remove Cd (II). Experimental data for the
sorption of As (III) and As (V) was adjusted appropria-
tely to the Langmuir model with r
2
values of 0.972 and
0.995, respectively. The high removal ef±ciency for As
(III) and As (V) by metallurgical wastes is due to the
presence of jarosite and Fe-oxyhydroxides minerals
because of their high af±nity for As (III) and As (V).
ACKNOWLEDGEMENTS
We express our gratitude to Industrial Minera
Mexico (IMMSA) for their ±nancial and logistic
support and CONACyT for LG-MJ PhD scholarship.
The authors thank Guillermo Pérez, Heriberto Rosas,
Inés Ramos Bautista, Raquel Domínguez Martínez
and José Iván Morales Arredondo for their assistan-
ce in sampling, sample preparation and laboratory
analysis. We also thank Teresa Pi Puig (Institute of
Geology, UNAM) for XRD analyses and Jardine Wall
and Alejandra Romo for the grammatical revision of
this document.
REFERENCES
Appelo C.A.J. and Postma D. (2005). Geochemistry,
groundwater and pollution, 2nd ed. A. A. Balkema,
Leiden, The Netherlands.
Asta M.P., Cama J., Martínez M. and Giménez J. (2009).
Arsenic removal by goethite and jarosite in acidic con-
ditions and its environmental implications. J. Hazard.
Mat. 171, 965-972.
Babcan J. (1971). Synthesis of Jarosite, KFe
3
(SO
4
)
2
(OH)
6
.
Geol. Zb. 22 (2), 299-304.
Baron D. and Palmer C.D. (1996). Solubility of jarosite
at 4–35°C. Geochim. Cosmochim. Acta. 60, 185-195.
Fig. 3.
As (III) and As (V) removal from aqueous solutions by metallurgical wastes: (a1)
Plot of measured As (III) sorption per unit weight of solid (C
s
, mmol/kg) against
As (III) concentration in the equilibrium solution (C
e
, mmol/L). (a1) Theoretical
Langmuir isotherms (
) for the As (III) sorption and experimental data (
) (b1)
Plot of measured As (V) sorption per unit weight of solid (C
s
, mmol/kg) against
As (V) concentration in the equilibrium solution (C
e
, mmol/L). (b2) Theoretical
Langmuir isotherms (
) for the As (V) sorption and experimental data (
)
(a)
(b)
16
14
12
10
8
0
0
0.4
0.8
C
e
, mmol/L
C
s
, mmol/kg
1.2
1.6
2
16
12
8
4
0
0
0.4
0.8
C
e
, mmol/L
C
s
, mmol/kg
1.2
1.6
2
20
16
12
8
4
0
0.2
0.4
C
e
, mmol/L
C
s
, mmol/kg
1
0.8
0.6
20
16
12
8
4
0
0
0.2
0.4
C
e
, mmol/L
C
s
, mmol/kg
0.6
0.8
(b1)
(b2)
(a1)
(a2)
L.G. Martínez Jardines
et al.
248
Bickmore B.R., Rosso K.M., Tadanier C.J., Bylaskaand
E.J. and Doud D. (2006). Bond-valence methods
for pKa prediction. II. Bond-valence, electrostatic,
molecular geometry, and solvation effects. Geochim.
Cosmochim. Acta.70, 4057-4071.
Beesley L. and Marmiroli M. (2011). The immobilisation
and retention of soluble arsenic, cadmium and zinc by
biochar. Environ. Pollution. 159, 474-480.
Brewster M. (1992). Removing arsenic from contaminated
waste water. Water Environ. Technol. 4, 54-57.
Celis R., Hermosín M.C. and Cornejo J.E. (2000). Heavy
metal adsorption by functionalized clays. Environ. Sci.
Technol. 34, 4593-4599.
Ciccu R., Ghiani M., Serci A., Fadda S., Peretti R. and
Zucca A. (2003). Heavy metal immobilization in the
mining-contaminated soils using various industrial
wastes. Minerals Engineering. 16, 187-192.
Daus B., Wennrich R. and Weiss H. (2004). Sorption ma-
terials for arsenic removal from water: a comparative
study. Water Res. 38, 2948-2954.
Foster A., Brown G., Tingle N.and Parks G. (1998). Quan-
tative arsenic speciation in mine tailings using X-ray
absorption spectroscopy. Am. Miner. 83, 553-568.
Gerente C., Andrès Y., McKay G. and LeCloirec P. (2010).
Removal of arsenic(V) onto chitosan: From sorption
mechanism explanation to dynamic water treatment
process. Chem. Eng. J. 158, 593-598.
Giles C.H., McEwan T.H., Nakhawa S.N. and Smith D.
(1960). Studies in adsorption. Part XI. A system of
classi±cation of solution adsorption isotherms, and
its use in diagnosis of adsorption mechanisms and
in measurement of speci±c surface areas of solids. J.
Chem. Soc. 3973-3993.
Goldberg S. (2000). Competitive Adsorption of Inorganic
Arsenic Species on Oxides and Clay Minerals. Preprint
of extended Abstracts, vol. 40. No. 2 Symposia Paper
presented before The Division of Environmental
Chemistry. American Chemical Society, Washington,
D.C.
Goldberg S. and Johnston C.T. (2001). Mechanisms of
arsenic adsorption on amorphous oxides evaluated
using macroscopic measurements, vibrational spectros-
copy, and surface complexation modeling. J. Colloid
Interface Sci. 234, 204-216
Goldberg S. (2002). Competitive adsorption of arsenate
and arsenite on oxides and clay minerals. Soil Sci. Soc.
Am. J. 66, 413-421.
Gimenez J., Martínez M., De Pablo J., Rovira M. and Duro
L. (2007). Arsenic sorption onto natural hematite, mag-
netite, and goethite. J. Hazard. Materials. 141, 575-580.
Iakovleva V.P. (2003). UV Spectrophotometric Studies of
Arsenic (III) and Antimony (III) Aqueous Chemistry
from 25 to 300 ºC. A dissertation submitted to the
Swiss Federal Institute of Technology Zurich. Degree
of Doctor of Science. http://e-collection.library.ethz.
ch/eserv/eth:27055/eth-27055-02.pdf
Kim Y., Kim Ch., Choy I., Rengarajaj S. and Yi J. (2004).
Arsenic Removal Using Mesoporous Alumina Pre-
pared via a Templating Method. Environ. Sci. Technol.
38, 924-931.
Kwon J.S., Yun S-T., Kim S-O., Mayer B. and Hutcheon
I. (2005). Sorption of Zn(II) in aqueous solutions by
scoria. Chemos. 60, 1416-1426.
Lin Y.Z. and Puls R.W. (2000). Adsorption, desorption
and oxidation of arsenic affected by clay minerals and
aging process. Environ. Geol. 39(7), 753-759.
Malferrari D.B., Brigatti M.F., Laurora A., Pini S. and
Medici L. (2007). Sorption kinetics and chemical
forms of Cd(II) sorbed by thiol-functionalised 2:1 clay
minerals. J. Hazard. Mater. 143, 73-81.
Manning B. and Goldberg S. (1997).
Adsorption
and stability of arsenic(III) at the clay mineral-water
interface. Environ. Sci. Technol. 31, 2005-2011.
Mohan D. and Singh K.P. (2002). Single and multicompo-
nent adsorption of cadmium and zinc using activated
carbon derived from bagasse-an agricultural waste.
Water Res. 36, 2304-2314.
Mulligan C.N., Yong R.N. and Gibbs B.F. (2001). Re-
mediation technologies for metal-contaminated soils
and groundwater: an evaluation. Eng. Geology. 60,
193-207.
Panuccio M.R., Sorgona A., Rizzo M. and Cacco G.
(2009). Cadmium adsorption on vermiculite, zeolite
and pumice: Batch experimental studies. J. Environ.
Management 90, 364-374.
Pierce M.L. and Moore C.M. (1982). Adsorption of arse-
nite and arsenate on amorphous iron hydroxide, Water
Res. 16, 1247-1253.
Rao G.P.C., Satyaveni S., Ramesh A., Seshaiah K., Mur-
thy K.S.N. and Choudary N.V. (2006). Sorption of
cadmium and zinc from aqueous solutions by zeolite
4A, zeolite 13X and bentonite. J. Environ. Manage-
ment. 81, 265-272.
Romero F.M., Armienta M.A. and Carrillo A., (2004).
Arsenic sorption by carbonate-rich aquifer material, a
control on arsenic mobility at Zimapán, México. Arch.
Environ. Contam.Toxicol. 47, 1-13.
Romero F.M., Núñez-Alvarez L., Gutiérrez M.E., Armien-
ta M.A. and Ceniceros-Gómez A.E. (2010). Evaluation
of the potential of indigenous calcareous shale for
neutralization and removal of arsenic and heavy metals
of acid mine drainage in Taxco mining area, Mexico.
Arch. Environ. Contam. Toxicol. 60, 191-203.
Sadiq M. (1997). Arsenic chemistry in soils: an overview
of thermodynamic predictions and ±eld observations.
Water Air Soil Pollut. 93, 117-136.
REMOVAL OF ARSENIC AND CADMIUM BY TWO NATURAL SORBENT MATERIALS
249
Sadowski Z., Polowczyk I.,
Farbiszewska T. and
Farbisze-
wska-Kiczma J. (2001). Adhesion of jarosite particles
to the mineral surface. Prace Naukowe Instytutu
Górnictwa Politechniki Wroc┼éawskiej. Konferencje
rok: 95, 93-102
Srivastava P.S., Singh B. and Angove M. (2005). Competi-
tive adsorption behaviour of heavy metals on kaolinite.
J. Colloid Interf. Sci. 290, 28-38.
Tiller K.G. (1989). Heavy metals in soils and their en-
vironmental signi±cance. Adv. Soil Sci. 9; 113-142.
Trivedi P. and Axe L. (2001). Predicting divalent metal
sorption to hydrous Al, Fe and Mn oxides. Environ.
Sci. Technol. 35, 1779–1784.
US EPA (2006). Method 6200 and ±eld portable X-ray
²uorescence analysis for metals in soil.
Veli S. and Alyüz B. (2007). Adsorption of copper and
zinc from aqueous solutions by using natural clays. J.
Hazard. Mat. 149, 226-233.
Villalobos M. and Antelo J. (2011). A uni±ed surface struc-
tural model for ferrihydrite proton charge, electrolyte
binding, and arsenate adsorption. Rev. Int. Contam.
Ambie. 27, 139-151.
Wang S. and Mulligan C.N. (2006). Occurence of arsenic
contamination in Canada: sources, behaviour and dis-
tribution. Sci. Total Environ. 366, 701-721.
logo_pie_uaemex.mx