In this paper, goethite waste from hydrometallurgy of zinc was used as a raw material for arsenic adsorbent preparation. The goethite waste adsorbent (GWA) was characterized with scanning electron microscope (SEM), X-ray powder diffraction (XRD), and particle size distribution analysis. The adsorption of As(III) on GWA was studied as a function of contact time, pH, and coexisting anions. The safety of GWA usage in the wastewater treatment process was assessed by toxicity characteristic leaching procedure (TCLP) tests. The equilibrium adsorption data fitted well with the Langmuir isotherm model, and the maximum adsorption capacity of As(III) on GWA was 51.47 mg.g−1. GWA showed higher adsorption capacity at weak alkaline pH values (7.0–9.5). The coexisting PO43− and SiO32− presented significant adsorption competition with As(III) in aquatic systems. No significant heavy metals leaching was observed for GWA and As(III) loaded GWA in TCLP tests, which implied the safety of GWA as an adsorbent for arsenic containing wastewater treatment.

INTRODUCTION

Arsenic, one of the most toxic contaminants, has attracted increasing attention in recent years due to widely occurring arsenic contamination events. In aquatic environments, arsenic exists in the state of arsenite As(III) and arsenate As(V), but As(III) has been reported to be more toxic than As(V) (Straif et al. 2009; Zhao et al. 2010). Furthermore, As(III) is more mobile than As(V) due to its weaker affinity to soils and sediments (Smedley & Kinniburgh 2002; Tufano & Fendorf 2008). Hence, it is necessary to develop economical and efficient methods for arsenic removal, especially for As(III).

Numerous treatment technologies, such as adsorption (Pratap et al. 2009), sulfide precipitation (Harper & Kingham 1992), coagulation (Divagar et al. 2010), filtration (Damodar & Thiruvenkatachari 2009), and reverse osmosis (Ning 2002) have been developed to remove arsenic from water. Among them, adsorption is considered to be one of the most promising technologies for its low cost and high efficiency. Iron oxides and oxyhydroxides are classes of common adsorbents for arsenic removal because of their adsorption capacity, strong binding of adsorbed arsenic, and low cost (Swedlund & Webster 1999; Altundoğan et al. 2002; Mamindy-Pajany et al. 2009). Recently, attempts to recycle industrial waste iron oxide as adsorbent have been made (Tien Vinh et al. 2009; Randhawa et al. 2014). The utilization of waste iron oxide may offer many advantages including a lesser amount of overall sludge production and low cost.

Goethite is the most common iron oxide in soils, and has good adsorption capacity for both As(III) and As(V). The adsorption of arsenic onto goethite has been extensively studied (Giménez et al. 2007; Asta et al. 2009; Amstaetter et al. 2010). The goethite waste is generated by the ‘Goethite’ process, which is used to remove iron from acid leach liquors for the production of electrolytic zinc (Pelino et al. 1996). The electrolytic process for zinc metal production accounts for about 80% of the zinc produced in the world. The impurities including iron, cadmium, copper, and so on are released into leach solution in the leaching procedure, and these impurities should be removed from the leach solution in order to produce high purity zinc metal. Iron is subsequently precipitated from the leach solution in the form of hydrated ferric oxide or basic salt. The iron precipitation technologies include the jarosite process, goethite process, and hematite process. The goethite waste from zinc production contains significant amounts of goethite, which suggests its possible application in arsenic adsorption in aqueous solution. Zinc hydrometallurgical companies produce about 1,000,000 tons of goethite waste per year in China. Due to the large amount of production and the presence of impurities in goethite waste, the disposal of goethite waste causes serious economic and environmental difficulties for zinc companies. However, the recycling of goethite waste as an adsorbent for arsenic removal has not been reported.

The aim of this work was to examine the As(III) adsorption behavior of goethite waste generated from electrolytic zinc production. The adsorption isotherms, kinetics and the effects of solution pH and coexisting ions in the adsorption process were studied in detail. Moreover, the heavy metals leaching characteristic of goethite waste was studied in the course of As(III) adsorption.

MATERIALS AND METHODS

Materials

The goethite waste sample was collected from the goethite process in a zinc hydrometallurgy company located in Guangdong province, China. Twenty-five grams of wet samples were dispersed in 50 mL of deionized water, filtered with a 0.45 μm filter membrane with a suction filtration device, then rinsed with 2 mM NaOH and deionized water. During the rinsing procedure with deionized water, most of the soluble impurity ions such as Zn2+, Pb2+, Cd2+, which adhere to the surface of goethite waste particles, were removed. After the deionized water rinse, trace amounts of residual soluble impurity ions were precipitated using NaOH solution rinsing procedure. Finally, the products were dried at 70 °C, ground to fine powder with a mortar, and stored as goethite waste adsorbent (GWA).

Adsorbent characterization

Imaging was performed by a scanning electron microscope (SEM, Cambridge Instruments S-360). Phase analysis was conducted using the quantitative X-ray diffraction (XRD) method using Cu Kα radiation and a graphite monochromator (X-ray diffractometer) (Rigaku Rint 2200). Particle size distribution was tested using a laser diffraction particle size analyzer (Mastersizer 2000).

Batch adsorption test

Batch experiments were carried out to examine the adsorption behavior of As(III) on goethite waste under different operation conditions. Stock solutions of As(III) (1,000 mg L−1) were prepared by dissolving sodium arsenite (NaAsO2) (Beijing Chemicals, China) in deionized water. Working solutions for experiments were freshly diluted from the stock solution. All experiments were conducted in triplicate. The relative standard deviation (RSD) of triplicate analyses was normally lower than 3%.

Adsorption isotherm studies were performed under conditions of initial As(III) concentration in the range of 0.5–150 mg L−1, initial pH 7, reaction time 24 h (equilibrium was achieved at 1.6 h), 120 rpm shaking at 25 ± 1 °C. All batch experiments were conducted in 250 mL conical flasks with a cover by taking 200 mg (dry weight) of goethite waste with 100 mL of As(III) working solution. All samples were collected, filtrated with a 0.45 μm filter membrane, and stored for analysis.

To determine the adsorption kinetic of As(III) on goethite waste, batch experiments were performed under conditions of initial As(III) concentrations 35 mg L−1 and 150 mg L−1, and initial pH 7. The samples were taken at reaction times of 5, 15, 30, 45, 60, 100, 150, and 200 min. The effect of pH on arsenite adsorption efficiency was determined at varying pH values (pH 3–12) using 0.1 M NaOH or 0.1 M HCl to adjust the pH of the solution. The heavy metals concentration of the solution, including arsenic, lead, cadmium, and so on, was measured with inductively coupled plasma–atomic emission spectrometry (ICP-AES) (725, Agilent). The amount of arsenite adsorbed on the goethite waste was measured by calculating the arsenic concentration difference in the initial and residual working solution. All experimental were run in triplicate.

To study the effects of competing anions, solutions were made by simultaneously adding As(III) and a single competing anion of interest. The competing anions were sulfate, silicate, and phosphate. The competing anion concentrations of As(III) solution were controlled at 5 mM under the conditions of As(III) concentrations 5.0 mg·L−1, initial pH 7.0 ± 0.1, and GWA 2.0 g·L−1.

Toxicity characteristic leaching procedure

The toxicity characteristic leaching procedure (TCLP) tests of goethite waste and As(III) absorbed goethite waste (0.572 mmol As·g−1) were carried out according to the procedure described by the US EPA method 1311. The extraction fluid consisted of a buffered acetic acid solution (pH 4.93 ± 0.05), prepared by adding NaOH and glacial acetic acid to water, and diluting to a final volume.

The dry test solid samples were mixed with extraction fluid (solid/liquid, ratio 1:20), then agitated (speed, 280 ± 5 rpm) for 18 h at 30 °C using a mechanical shaker. The supernatant was separated by centrifugation at 4,000 rpm for 15 min, filtered through a 0.45 μm filter membrane, acidified with concentrated nitric acid, and stored in amber vials at 4 °C. Then, the Cd, Cr, Pb, and As concentrations in the supernatant were measured by ICP-AES. Concentrations of Hg were determined by XGY-1011A atomic fluorescence spectrometer.

RESULTS AND DISCUSSION

Characterization of the GWA

In order to determine the surface characterization of GWA, the scanning electron microscopy (SEM) images are presented (Figure 1) under different magnifications. The GWA particles showed no regular characteristics, which may result from grinding during sample preparation. It can be seen that the adsorbent particles are aggregated with many small particles, resulting in a rough surface and porous structure.
Figure 1

SEM images of GWA.

Figure 1

SEM images of GWA.

The particle size distribution of GWA is depicted in Figure 2. The d (0.1), d (0.5), and d (0.9) of GWA particles are 2.69 μm, 7.34 μm, and 16.58 μm, respectively. The particle size distribution of GWA is suitable for application as a type of additive adsorbent in the wastewater treatment process. The small particle size indicates the relatively large specific surface area in the aqueous solution, and it is responsible for the high arsenic adsorption capacity of GWA. However, the smaller particle size also makes the removal of particles loaded with arsenic from wastewater more difficult. Figure 3 shows the XRD patterns for the GWA samples. The materials detected in GWA are primarily goethite (FeOOH), gypsum (CaSO4.2H2O), franklinite (ZnFe2O4), and sphalerite (ZnS).
Figure 2

Particle size distribution of GWA.

Figure 2

Particle size distribution of GWA.

Figure 3

XRD patterns of GWA.

Figure 3

XRD patterns of GWA.

Adsorption isotherm

In order to determine the adsorption capacity of GWA for As(III), the equilibrium adsorption isotherms are of fundamental importance. The Langmuir isotherm models were studied to determine the isotherm constants and predict the maximum As(III) adsorption capacity of GWA. These models assume monolayer adsorption onto a homogeneous surface with a finite number of identical sites (Jeon et al. 2009). It can be written in the form:
formula
1
where Qmax is the maximum adsorption capacity (mmol g−1), qe is the equilibrium adsorption amount (mmol·g−1), Ce is the equilibrium adsorbate concentration (mmol·L−1), and Ka is the isotherm constant.
As shown in Figure 4, the Langmuir isotherm model fits the experimental adsorption data well. The fitting Qmax value of GWA was 0.687 mmol·g−1 (51.47 mg·g−1) and the fitting determination coefficient (R2) was 0.956. The As(III) adsorption capacity of other studied adsorbents is presented in Table 1.
Table 1

Comparison of As(III) adsorption capacities of GWA and other adsorbents

AdsorbentAdsorption capacity (mg·g−1)ConditionReference
GWA 51.47 pH 7.0 ± 0.1; 25 ± 1 °C The present work 
Goethite 7.5 pH 5.5, 25 ± 0.5 °C Ladeira & Ciminelli (2004)  
Iron oxide coated cement 0.69 pH 7, 27 ± 2 °C Kundu & Gupta (2007)  
Nanocrystalline magnetite 3.65 pH 7.9 Bujňáková et al. (2013)  
Zero-valent iron 2.28 pH 7 Sasaki et al. (2009)  
Nanostructured Fe(III)–Cr(III) mixed oxide 8.24 pH 7 ± 0.2 Basu & Ghosh (2011)  
AdsorbentAdsorption capacity (mg·g−1)ConditionReference
GWA 51.47 pH 7.0 ± 0.1; 25 ± 1 °C The present work 
Goethite 7.5 pH 5.5, 25 ± 0.5 °C Ladeira & Ciminelli (2004)  
Iron oxide coated cement 0.69 pH 7, 27 ± 2 °C Kundu & Gupta (2007)  
Nanocrystalline magnetite 3.65 pH 7.9 Bujňáková et al. (2013)  
Zero-valent iron 2.28 pH 7 Sasaki et al. (2009)  
Nanostructured Fe(III)–Cr(III) mixed oxide 8.24 pH 7 ± 0.2 Basu & Ghosh (2011)  
Figure 4

Equilibrium isotherms for As(III) adsorption on goethite waste (pH 7.0 ± 0.1; 25 ± 1 °C; sorbent dose 20 g L−1. The error bars represent standard deviation).

Figure 4

Equilibrium isotherms for As(III) adsorption on goethite waste (pH 7.0 ± 0.1; 25 ± 1 °C; sorbent dose 20 g L−1. The error bars represent standard deviation).

The adsorption capacity of GWA was higher than other experimental prepared iron oxides, including goethite, iron oxide coated cement, nanocrystalline magnetite, and Fe(III)–Cr(III) mixed oxide. This may be attributed to the micro size hydrated iron oxide particles, and is supported by the SEM images. In addition, it was reported that the addition of other metals, such as Cr (Basu & Ghosh 2011), Mn and Co (Zhang et al. 2010) into iron oxides can enhance the arsenic adsorption capacity of iron oxides. For goethite waste, the soluble impurity ions such as Zn2+, Pb2+, Cd2+, Cu2+ in leach solution were coprecipitated with the Fe3+, and the goethite waste was the combination of polymetallic hydrated oxides. The higher As(III) adsorption capacity of GWA may be due to the existence of other metals in the goethite waste as well.

Effect of pH

The pH is an important parameter in the arsenic adsorption process because the variation of proton concentration can strongly modify the chemical speciation of arsenic as well as the surface charge of sorbents.

Figure 5 demonstrates the efficiency of arsenic removal, and shows the increase in arsenic removal efficiency for As(III) with increasing pH from 3.1 to 7.0, which remained almost the same in the pH range 7.0 to 9.5, while decreasing with the increasing pH range of 9.5–11.8.
Figure 5

Effect of pH on As(III) adsorption by GWA (T: 25 ± 2 °C; initial As concentration 5.1 mg L−1; reaction time 12 h; sorbent dose 20 g L−1. The error bars represent standard deviation).

Figure 5

Effect of pH on As(III) adsorption by GWA (T: 25 ± 2 °C; initial As concentration 5.1 mg L−1; reaction time 12 h; sorbent dose 20 g L−1. The error bars represent standard deviation).

In general, the adsorption of As(III) is less pH dependent. According to the previous results reported, the isoelectric point (iep) of goethite was observed at pH 6.7 (Lakshmipathiraj et al. 2006), which means that the goethite carries a positive charge when the pH is lower than 6.7 and a negative charge when the pH is higher than 6.7. However, the As(III) exists in the state of H3AsO30 when the pH is less than 6.2 (Wolthers et al. 2005; Guo et al. 2009), which implies that the adsorption of As(III) under acidic conditions is not dominated by electrostatic attraction. In acidic conditions, As(III) competes with hydrogen ions for adsorption onto GWA. In strong alkaline conditions (pH 9.5–11.8), the GWA becomes negatively charged and electrostatic repulsion between resulted in a decrease of adsorption.

Adsorption kinetics

Effect of reaction time on the As(III) adsorption amount is one of the important factors determining the applicability of the adsorbent for practical wastewater treatment. Batch adsorption experiments were carried out for the adsorption of As(III), varying reaction time at constant temperature, pH, and shaking speed conditions. Results (Figure 6) reveal that a rapid uptake within the first 60 min was achieved. Then, the uptake stabilized after the first 100 min of agitation time, implying that equilibrium had been reached.
Figure 6

Adsorption kinetics of As(III) onto goethite waste (initial concentration of As(III) 35 mg L−1, 150.0 mg/L, pH 7.0 ± 0.1, temperature 25 ± 1 °C, sorbent dose 20 g L−1. The error bars represent standard deviation).

Figure 6

Adsorption kinetics of As(III) onto goethite waste (initial concentration of As(III) 35 mg L−1, 150.0 mg/L, pH 7.0 ± 0.1, temperature 25 ± 1 °C, sorbent dose 20 g L−1. The error bars represent standard deviation).

The kinetics data were fitted to the pseudo-first order and pseudo-second order models, respectively presented as follows in Equations (2) and (3):
formula
2
formula
3
where qe (mmol·g−1) and qt (mmol·g−1) are the amount of As(III) adsorbed onto GWA at equilibrium and reaction time t, respectively; t is the contact time (min), k1 (min−1) and k2 (min−1) are, respectively, the rate constants.

All the kinetics parameters of As(III) adsorption onto GWA are shown in Table 2. The pseudo-second-order model was observed to exhibit a better fit for the kinetics data than the pseudo-first-order model. The k2 values showed that As(III) sorption on GWA took place more rapidly at a lower concentration (35 mg L−1) than at a higher concentration (150 mg L−1). This might be due to the availability of a greater number of sorption sites on the oxide surface per molecule of the sorbent in the case of its lower concentration.

Table 2

Kinetics parameters of As(III) adsorption onto GWA

As(III) concentration (mg L−1)Pseudo-first order
Pseudo-second order
K1 (min−1)qe (mg g−1)R2K2 (min−1)qe (mg g−1)R2
35 0.168 0.285 0.983 3.110 0.175 0.998 
150 0.162 0.494 0.984 0.497 0.525 
As(III) concentration (mg L−1)Pseudo-first order
Pseudo-second order
K1 (min−1)qe (mg g−1)R2K2 (min−1)qe (mg g−1)R2
35 0.168 0.285 0.983 3.110 0.175 0.998 
150 0.162 0.494 0.984 0.497 0.525 

The adsorption process is a multi-step process, and the rate limiting step may be of three types: 1) diffusion across the liquid film surrounding the solid particles; 2) intra-particle diffusion; and 3) physical or chemical adsorption at a site (Wu et al. 2012). The better fitting of the pseudo-second model suggested the presence of a chemisorption reaction in the process of As(III) sorption on GWA. In addition, the k2 value decreased with increasing As(III) initial concentration, revealing that As(III) adsorption was more endothermic.

Effect of competing ions

Various anions, especially oxyanions, may negatively influence the adsorption of arsenic in aquatic systems. Therefore, three kinds of oxyanions were studied to investigate the effects of coexisting anions on the As(III) adsorption process. Figure 7 demonstrated the effect of individual competing anions on As(III) adsorption. It is evident that the ion competition between As(III) and coexisting anions followed the order: . The presence of 5 mM reduced the As(III) removal efficiency from 93.2% to 10.1%, 76.9% and 91.9%, respectively. This somewhat agrees with other results obtained from copper ferrite (Tu et al. 2013), cupric oxide nanoparticles (Martinson & Reddy 2009), and MnFe2O4 nanomaterials (Zhang et al. 2010). It is reported that arsenate, phosphate, and silicate have similar tetrahedral structures, and they can form an inner-sphere complex with the hydroxyl groups at the surface of adsorbents (Hui et al. 2008). The decrease of As(III) removal efficiency might result from the competition between arsenic and oxyanions.
Figure 7

Effect of competing anions on As(III) adsorption by GWA (T: 25 ± 2 °C, pH 7, concentration of GWA 2 g/L, initial As(III) concentration 5.0 mg/L. The error bars represent standard deviation).

Figure 7

Effect of competing anions on As(III) adsorption by GWA (T: 25 ± 2 °C, pH 7, concentration of GWA 2 g/L, initial As(III) concentration 5.0 mg/L. The error bars represent standard deviation).

Leaching test

According to the US EPA, if the leached heavy metals (Hg, Cd, Cr, Pb, As) concentration from the solid waste exceeds the regulated level of 40CFR26 1.24, the solid waste should be marked as hazardous waste. In the present work, the TCLP was carried out to assess the safety of goethite waste as an adsorbent for arsenic removal from wastewater. Table 3 shows the leached pollutants (Hg, Cd, Cr, Pb, As) concentration for goethite waste before and after As(III) was loaded. Among all samples, the leached Hg, Cd, Cr, Pb, and As concentrations in the TCLP fluid were significantly lower than the US EPA specified limit. These results suggest that the soluble impurities were effectively removed from the raw goethite waste in the adsorbent preparation process. In addition to that, the binding force between the arsenic and goethite waste surface was relatively strong. Thus, goethite waste could be marked as a non-hazardous material in the arsenic-containing wastewater treatment process.

Table 3

TCLP results of GWA before and after As(III) was loaded

Leached pollutantsBefore As(III) loadedAfter As(III) loadedEPA regulated level
Hg (mg L−10.004 0.002 0.2 
Cd (mg L−10.21 0.16 1.0 
Cr (mg L−10.08 0.07 5.0 
Pb (mg L−10.01 0.01 5.0 
As (mg L−10.03 0.05 5.0 
Leached pollutantsBefore As(III) loadedAfter As(III) loadedEPA regulated level
Hg (mg L−10.004 0.002 0.2 
Cd (mg L−10.21 0.16 1.0 
Cr (mg L−10.08 0.07 5.0 
Pb (mg L−10.01 0.01 5.0 
As (mg L−10.03 0.05 5.0 

CONCLUSIONS

GWA generated from the hydrometallurgy of zinc was applied as the raw material for arsenic adsorbent, and the adsorption conditions for As(III) removal from aqueous solution by GWA, kinetic, isotherm, pH, competing ions, and heavy metals leaching was performed. Kinetic studies revealed that the pseudo-second order model suitably described the adsorption kinetics of As(III) on GWA. The equilibrium data were well fitted to the Langmuir isotherm model. The maximum As(III) adsorption capacity of GWA was 51.47 mg·g−1. pH studies showed that maximum As(III) adsorption could be achieved at weak alkaline pH values (7.0–9.5). The ion competition between As(III) and coexisting anions followed the order: . No significant heavy metals leaching was observed for GWA and As(III) loaded GWA in TCLP tests, which implied the safety of GWA as an adsorbent for treatment of arsenic-containing wastewater.

ACKNOWLEDGEMENTS

This research was supported by the National Natural Science Fund, People's Republic of China (No. 51404028 and No. 51504028) and Beijing General Research Institute of Mining & Metallurgy research project (Grant YJ-02-1417).

REFERENCES

Altundoğan
H. S.
Altundoğan
S.
Tümen
F.
Bildik
M.
2002
Arsenic adsorption from aqueous solutions by activated red mud
.
Waste Management
22
(
3
),
357
363
.
Amstaetter
K.
Borch
T.
Larese-Casanova
P.
Kappler
A.
2010
Redox transformation of arsenic by Fe(II)-activated goethite (alpha-FeOOH)
.
Environmental Science & Technology
44
(
1
),
102
108
.
Asta
M. P.
Cama
J.
Martínez
M.
Giménez
J.
2009
Arsenic removal by goethite and jarosite in acidic conditions and its environmental implications
.
Journal of Hazardous Materials
171
(
1–3
),
965
972
.
Bujňáková
Z.
Baláž
P.
Zorkovská
A.
Sayagués
M. J.
Kováč
J.
Timko
M.
2013
Arsenic sorption by nanocrystalline magnetite: an example of environmentally promising interface with geosphere
.
Journal of Hazardous Materials
262
(
8
),
1204
1212
.
Damodar
P.
Thiruvenkatachari
V.
2009
Biological filtration for removal of arsenic from drinking water
.
Journal of Environmental Management
90
(
5
),
1956
1961
.
Giménez
J. G.
Martinez
M.
de Pablo
J.
Rovira
M.
Duro
L.
2007
Arsenic sorption onto natural hematite, magnetite, and goethite
.
Journal of Hazardous Materials
141
(
3
),
575
580
.
Harper
T. R.
Kingham
N. W.
1992
Removal of arsenic from wastewater using chemical precipitation methods
.
Water Environment Research
64
(
3
),
200
203
.
Jeon
C. S.
Baek
K.
Park
J. K.
Oh
Y. K.
Lee
S. D.
2009
Adsorption characteristics of As(V) on iron-coated zeolite
.
Journal of Hazardous Materials
163
(
2–3
),
804
808
.
Ladeira
A. C. Q.
Ciminelli
V. S. T.
2004
Adsorption and desorption of arsenic on an oxisol and its constituents
.
Water Research
38
(
8
),
2087
2094
.
Lakshmipathiraj
P.
Narasimhan
B. R. V.
Prabhakar
S.
Raju
G. B.
2006
Adsorption of arsenate on synthetic goethite from aqueous solutions
.
Journal of Hazardous Materials
136
(
2
),
281
287
.
Mamindy-Pajany
Y.
Hurel
C.
Marmier
N.
Roméo
M.
2009
Arsenic adsorption onto hematite and goethite
.
Comptes Rendus Chimie
12
(
8
),
876
881
.
Martinson
C. A.
Reddy
K. J.
2009
Adsorption of arsenic(III) and arsenic(V) by cupric oxide nanoparticles
.
Journal of Colloid & Interface Science
336
(
2
),
406
411
.
Ning
R. Y.
2002
Arsenic removal by reverse osmosis
.
Desalination
143
(
2
),
237
241
.
Pelino
M.
Cantalini
C.
Abbruzzese
C.
Plescia
P.
1996
Treatment and recycling of goethite waste arising from the hydrometallurgy of zinc
.
Hydrometallurgy
40
(
s1–2
),
25
35
.
Pratap
C.
Shigeru
K.
Toshinori
K.
Shigeo
S.
2009
Arsenic adsorption from aqueous solution on synthetic zeolites
.
Journal of Hazardous Materials
162
(
1
),
440
447
.
Randhawa
N. S.
Murmu
N.
Tudu
S.
Sau
D. C.
2014
Iron oxide waste to clean arsenic-contaminated water
.
Environmental Chemistry Letters
12
(
4
),
1
6
.
Sasaki
K.
Nakano
H.
Wilopo
W.
Miura
Y.
Hirajima
T.
2009
Sorption and speciation of arsenic by zero-valent iron
.
Colloids & Surfaces A Physicochemical & Engineering Aspects
347
(
s1–3
),
8
17
.
Smedley
P. L.
Kinniburgh
D. G.
2002
A review of the source, behaviour and distribution of arsenic in natural waters
.
Applied Geochemistry
17
(
5
),
517
568
.
Straif
K.
Benbrahim-Tallaa
L.
Baan
R.
Grosse
Y.
Secretan
B.
El Ghissassi
F.
Bouvard
V.
Guha
N.
Freeman
C.
Galichet
L.
Cogliano
V.
;
WHO International Agency for Research on Cancer Monograph Working Group
.
2009
A review of human carcinogens – part C: metals, arsenic, dusts, and fibres
.
Lancet Oncology
10
(
5
),
453
454
.
Tien Vinh
N.
Nguyen
T. V. T.
Tuan Linh
P.
Saravanamuth
V.
Huu Hao
N.
Kandasamy
J.
Khanh
N. H.
Duc Tho
N.
2009
Adsorption and removal of arsenic from water by iron ore mining waste
.
Water Science & Technology
60
(
9
),
2301
2308
.
Tu
Y. J.
You
C. F.
Chang
C. K.
Wang
S. L.
Chan
T. S.
2013
Adsorption behavior of As(III) onto a copper ferrite generated from printed circuit board industry
.
Chemical Engineering Journal
225
(
3
),
433
439
.
Tufano
K. J.
Fendorf
S.
2008
Confounding impacts of iron reduction on arsenic retention
.
Environmental Science & Technology
42
(
13
),
4777
4783
.
Wolthers
M.
Charlet
L.
Weijden
C. H. V. D.
Linde
P. R. V. D.
Rickard
D.
2005
Arsenic mobility in the ambient sulfidic environment: sorption of arsenic(V) and arsenic(III) onto disordered mackinawite
.
Geochimica et Cosmochimica Acta
69
(
14
),
3483
3892
.
Zhao
F. J.
McGrath
S. P.
Meharg
A. A.
2010
Arsenic as a food chain contaminant: mechanisms of plant uptake and metabolism and mitigation strategies
.
Annual Review of Plant Biology
61
(
4
),
535
559
.