Uptake of Zn2+ ions by a fully iron-exchanged clinoptilolite. Case study of heavily contaminated drinking water samples

Uptake of Zn2+ ions by a fully iron-exchanged clinoptilolite. Case study of heavily contaminated drinking water samples

ARTICLE IN PRESS WAT E R R E S E A R C H 41 (2007) 2763 – 2773 Available at www.sciencedirect.com journal homepage: www.elsevier.com/locate/watres ...

551KB Sizes 1 Downloads 49 Views

ARTICLE IN PRESS WAT E R R E S E A R C H

41 (2007) 2763 – 2773

Available at www.sciencedirect.com

journal homepage: www.elsevier.com/locate/watres

Uptake of Zn2+ ions by a fully iron-exchanged clinoptilolite. Case study of heavily contaminated drinking water samples Anthoula Dimirkou Department of Agriculture, Crop Production and Agricultural Environment, University of Thessaly, Soil Science Laboratory, N. Ionia 38446 Magnesias, Greece

art i cle info

ab st rac t

Article history:

Clinoptilolite, a natural zeolite, was used for the synthesis of a high surface area

Received 17 October 2006

clinoptilolite (Clin)–iron (Fe) oxide system, in order to be used for the removal of Zn2+

Received in revised form

ions from drinking water samples. The new system was obtained by adding natural Clin in

26 February 2007

an iron nitrate solution under strongly basic conditions. The Clin-Fe system has specific

Accepted 27 February 2007

surface area equal to 151 m2/g and is fully iron exchanged (Fe/Al ¼ 1.23). Batch adsorption

Available online 19 April 2007

experiments were carried out to determine the effectiveness of the Clin and the Clin–Fe

Keywords:

system in removal of Zn from drinking water. Adsorption experiments were conducted by

Clinoptilolite

mixing 1.00 g of each of the substrates with certain volume of water samples contaminated

Clinoptilolite–Fe system

with ten different Zn concentrations (from 7.65  105 to 3.82  102 M or from 5.00 to

Zn adsorption

2500 ppm Zn). For our experimental conditions, the maximum adsorbed Zn amount by Clin

Water treatment

was 71.3 mg/g, whereas by the Clin–Fe system 94.8 mg/g. The main factors that contribute

Hardness

to different adsorbed Zn amounts by the two solids are due to new surface species and

Dissolution

negative charge of the Clin–Fe system. In addition, the release of counterbalanced ions (i.e. Ca2+, Mg2+, Na+ and K+) was examined, as well as the dissolution of framework Si and Al. It was found that for most of the samples the Clin–Fe system releases lower concentrations of Ca, Mg and Na and higher concentrations of K than Clin, while the dissolution of Si/Al was limited. Changes in the composition of water samples, as well as in their pH and conductivities values were reported and explained. & 2007 Elsevier Ltd. All rights reserved.

1.

Introduction

Zinc is one of the most common elements in the earth’s crust. It is found in air, soil and water and is present in all foods. It is an essential element needed by the body in small amounts. Either too little or too much zinc can be harmful to health. In natural surface waters, the concentration of zinc is usually below 10 mg/l, and in groundwaters 10–40 mg/l. In tap water, the zinc concentration can be much higher as a result

of its leaching from piping and fittings. According to EC Drinking Water Directive, to World Health Organization and its Guidelines for Drinking Water Quality and to Environmental Protection Agency (EPA) in USA, zinc is an undesirable material with a guideline value of 5.00 mg/l in drinking water. Since zinc compounds are widely used in industry in paints, ceramics, batteries, wood, fabrics, drugs, sun blocks, deodorants, etc., the zinc manufacturing (and other industries) release during production large quantities of metals, mainly

Thermopilon 14 street, 15344 Pallini, Greece. Tel./fax: +30 210 6668215.

E-mail address: [email protected]. 0043-1354/$ - see front matter & 2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2007.02.045

ARTICLE IN PRESS 2764

WAT E R R E S E A R C H

41 (2007) 2763– 2773

Nomenclature

Co C X

initial Zn concentration (mol/l or ppm) solution Zn concentration (mol/l)

QP QA CSi CAl

maximum processing capacity (l/kg) practical specific capacity (eq/kg) solution concentration of dissolved Si (mol/l) solution concentration of dissolved Al (mol/l)

adsorbed Zn concentration (mol/kg)

Cd and Zn (Zoumis et al., 2000). Several treatment technologies, such as chemical precipitation, ultra filtration, adsorption and ion-exchange, reverse osmosis, and electrodialysis have been developed for eliminating zinc, and heavy metals in general, from water and wastewaters (Erdem et al., 2004). Many of these methods suffer from some drawbacks, such as high capital and operational costs and problem of disposal of residual metal sludge. Ion exchange is feasible when an exchanger has a high selectivity for the metal to be removed and the concentrations of competing ions are low. The metal may then be recovered by incinerating the metal-saturated resin and the cost of such a process naturally limits its application to only the more valuable metals. In many cases, however, the heavy metals are not valuable enough to warrant the use of special selective exchangers/resins from an economic point of view. This has encouraged research into using low-cost adsorbent materials to purify water contaminated with metals. Zeolites are microporous crystalline hydrated aluminosilicates that can be considered as inorganic polymers built from an infinitely extending three-dimensional network (similar to honeycomb) of tetrahedral TO4 units, where T is Si or Al, which form interconnected tunnels and cages. Each aluminum ion that is present in the zeolite framework yields a net negative charge, which is balanced by an extra framework cation, usually from the group IA or IIA. Water moves freely in and out of zeolite pores but the framework remains rigid. Porous zeolite is host to water molecules and ions of positive charge and the ability to exchange cations is one important property of them (Ruthven, 2001; Doula and Ioannou, 2003). Zeolites have been widely used in heavy metals adsorption experiments (Doula et al., 2002) due to their unique physical and chemical properties (crystallinity, thermal stability, welldefined cage structure of molecular size, ion exchange, etc.). The advantage of zeolites over resins, apart from their much lower cost, is their ion selectivities. Owing to zeolite structural characteristics and their adsorbent properties, they have been applied as chemical sieves, water softeners and adsorbents (Ghobarkar et al., 1999). They also have large specific surface area (SSA) and contain high concentrations of exchangeable cations, which give them a high cation exchange capacity (CEC). Iron (Fe) oxides are also active sorbents and play an important role in the reaction behavior of many ions in environmental sediments. Actually they can act as sinks in these environments because they have high affinity for metal contaminants, large surface area, microporous structure and their internal surface constitutes as much as 40% of their total sites (Trivedi and Axe, 2001). In the past many researchers have tried to synthesize mixed systems of clay minerals or zeolites with Fe-oxides, and they have proved that these systems are capable of adsorbing high concentrations of

inorganic species (Dimirkou et al., 1996, 2002) but also, for the case of Fe–zeolites systems, to improve zeolite catalytic characteristics (Morturano et al., 2001). Recently, a very simple Fe–overexchanged clinoptilolite (Clin) system was synthesized and tested for its effectiveness in removal of Mn2+ ions from drinking water (Doula, 2006), as well as in removal of Cu2+ ions from aquatic solutions (Doula, 2007). This Clin–Fe system has SSA equal to 151 m2/g, which is almost five times higher than that of untreated Clin (30.98 m2/ g). The experimental results obtained for the removal of Mn2+ ions from drinking water revealed that the Clin–Fe system has a noticeable higher Mn adsorption capacity (27.12 mg/g) than Clin (7.69 mg/g). Moreover, after the treatment with the Clin–Fe system, the water samples had significantly lower hardness because the new system acted simultaneously as a water softening material. Similar satisfactory results were obtained for Cu adsorption. It was proved that the Clin–Fe system not only adsorbs almost four times higher Cu concentration than Clin, but also has the ability to hold Cu2+ ions adsorbed. In the following, we synthesize again this Clin–Fe system in order to test it in removal of Zn2+ ions from contaminated drinking water samples.

2.

Materials and methods

2.1.

Clin

The Clin used in the present investigation comes from a layer situated in Thrace (North Greece). This material has been used in the past for metal adsorption experiments (Doula et al., 2002; Doula and Ioannou, 2003; Inglezakis et al., 2004) and consequently its physicochemical properties are well known. Because in a traditional aluminosilicates zeolite the source of the ion exchange capacity is the extent of isomorphous substitution of Al for Si in the tetrahedral framework, the theoretical exchange capacity can be derived from the elemental composition. Thus, the estimated CEC of Clin, with respect to its formula (Na0.2K0.6Mg0.7Ca2.0Al6.2Si29.8O7219.6H2O), is 235 meq/100 g. The CEC of a zeolite can be experimentally estimated by measuring the uptake of the ammonium cation at room temperature when equilibrium conditions are known to have been attained in the presence of a 1 M ammonium salt solution (Dyer, 2001). This practical CEC was measured and found equal to 115 meq/100 g. The difference between theoretical and experimental CEC is owing to ion-sieving phenomenon (Dyer, 2001). The powder XRD study showed that zeolite, feldspars and total micas+clays are present through the tuff whereas, the Clin used in the present study comes from the layer with the

ARTICLE IN PRESS WAT E R R E S E A R C H

higher Clin content (up to 90%). Its SSA is equal to 30.98 m2/g ˚ (Doula et al., 2002). and the average pore diameter is o20.0 A Before the synthesis of the new system as well as, the adsorption experiments the zeolite was finely ground and sieved to o0.02 mm.

2.2.

water. Their pH values ranged from 5.51 to 4.34, whereas their conductivities from 333 mS/cm to 91.8 mS/cm.

2.5.

Water sample

The water sample used comes from the Athens drinking water network and its physicochemical analysis is shown in Table 1. The concentrations of Na and K were measured by using a Korning 410 flame photometer, while, Si and Al by using a Varian Liberty 220 ICP emission spectrometer. Ca and Mg were volumetrically determined and from these two concentrations the value of sample hardness (in ppm CaCO3) was estimated.

2.4.

Zn2+ adsorption experiment

The adsorption experiment was divided into two stages. Firstly, several samples of 1.00 g of each solid were equilibrated with 95.0 ml of water sample (referred to as ‘‘equilibrium period’’). This period lasted for 48 h and after this period the sample pH values were measured. Chemical analysis was carried out for some of these equilibrated water samples in order to verify the changes caused by the contact with the adsorbents, as well as to verify the exact solution composition before the addition of the Zn2+ solutions. During the second experimental stage (referred to as ‘‘Zn adsorption’’) 5.00 ml of each of zinc stock solutions were added to the above-mentioned samples, at concentrations such that the total mixture contained 0, 7.65  105, 1.53  104, 3.06  104, 7.65  104, 1.53  103, 3.06  103, 7.65  103, 1.53  102, 2.29  102, and 3.82  102 mol/l Zn, (or 0, 5.00, 10.0, 20.0, 50.0, 100, 200, 500, 1000, 1500, and 2500 ppm Zn). The samples at this second stage were equilibrated for 48 h, were centrifuged for 10 min at 15,000 rpm and then their pH values and their conductivities were measured. According to literature, a 48-h contact period is sufficient to reach the studied system equilibrium (Ghanem and Mikkelsen, 1988; Puls and Bohn, 1988; Hui et al., 2005; ¨ ren and Kaya, 2006). However, Trivedi and Axe (2001) proved O that the adsorption of Zn on goethite was completed within the first 2 h of contact, while the adsorption on amorphous Fe-oxide did not reach equilibrium within a 7-day experiment, although the samples were close to equilibrium after the first 2 days. The researchers explained that this difference is due to high internally located sites of amorphous Fe-oxide, which contribute to significant intraparticle (surface) diffusion. However, the 48-h contact period was used also for a practical reason, since a good adsorbent should be effective and act quickly. Zinc concentrations were measured in the liquid phase by using a SpectrAA 300 Varian flame atomic spectrometer. The concentrations of adsorbed Zn species were calculated from the difference in measured and initial concentrations of the species in solution. The solution concentrations of Ca, Mg, Na, K, Si and Al were also measured.

Clin–Fe-oxide system

The Clin–Fe system was synthesized by following the method of pure goethite preparation, as described by Schwertmann and Cornell (1991). The change made in this study concerns the presence of Clin, which was added in the experimental flasks. The system was prepared by mixing 20.0 g of Clin, 100 ml of freshly prepared 1 M Fe(NO3)3 solution, and 180 ml of 5 M KOH solution in a 2 l polyethylene flask. The addition of KOH solution was rapid and with stirring. The suspension was diluted to 2 l with twice distilled water and was held in a closed polyethylene flask at 70 1C for 60 h. After the appropriate period the reaction vessel was removed from the oven, and the precipitate was centrifuged, washed (until free of NO 3 ions) and finally dried. After the 60-h period a dark red precipitate was obtained. The SSA of the Clin–Fe system is 151 m2/g and the average ˚. pore diameter is p20.0 A Spectroscopic methods (XRD, EPR, FTIR TG/DSC) were used for the characterization of the Clin–Fe-oxide system and were fully detailed elsewhere (Doula, 2007).

2.3.

2765

4 1 (200 7) 276 3 – 277 3

Stock zinc solutions

Ten stock zinc solutions from 100 to 50,000 ppm, were prepared by dissolving Zn(NO3)26H2O in twice distilled

Table 1 – Chemical analysis results of water samples before and after their contact with Clin and and Clin–Fe system

Waterc Clin Clin–Fe

pH

Cond.a

Ca

Mg

K

Na

Si

Al

Hardnessb

7.80 7.53 8.14

275 293 395

7.73  104 8.40  104 1.48  104

2.70  104 2.78  104 9.09  105

E0 4.00  105 2.86  103

1.80  104 5.00  104 2.51  104

2.97  105 3.36  104 8.16  105

1.05  105 1.10  105 2.41  106

104 112 23.9

The concentrations of Ca, Mg, Na, K, Si, and Al are in mol/l. a Conductivity in mS/cm. b In ppm CaCO3. c Water samples before treatment.

ARTICLE IN PRESS 2766

WAT E R R E S E A R C H

41 (2007) 2763– 2773

Since the subject of the present experiment was to remove Zn2+ ions from drinking water samples which could have been contaminated with heavy metals in different ways, for example by industrial wastes or accidentally, the pH values of the samples were not adjusted. In this way authors tried to simulate a possible real event and to study the response and the behavior of the adsorbents used to purify the contaminated samples. All experimental stages took place in triplicate, in a water bath at constant temperature (25 1C), and under a N2 atmosphere by placing the water bath in a glove box. The maximum processing capacity QP expresses the solution volume that can be purified from an ion with initial concentration C, by using a certain mass of sorbent or exchanger (Leinonen, 1999): QP ¼

½Q A  ½Co 

ðl=kgÞ,

(1)

where QA is the practical specific capacity of the sorbent (equal to experimentally determined adsorbed Zn concentration) in eq/kg and Co the initial Zn concentration in eq/l.

3.

Results and discussion

3.1.

Characterization of the Clin–Fe system

Elemental analysis of the Clin–Fe system revealed that it contains 14.1% Fe as amorphous Fe species. Its Si/Al ratio is almost equal to the respective ratio of the parent material, and its Fe/Al ratio is equal to 1.23. The CEC of zeolites is analogous to the Al content (Dyer, 2001) since these sites carry permanent negative charge because of the isomorphous substitution of Al3+ for Si4+. From this point of view, the Fe/Al ratio is a characteristic property of the substrate and indicates that, besides being fully Fe exchanged, the Clin–Fe system contains also an additional fraction of Fe ions that does not act as charge-balancing species. Generally, samples which contain more Fe (or other metallic cation) than calculated from the theoretical ion exchange capacity, are characterized as ‘‘overexchanged’’ samples. Overexchanged zeolites have been synthesized by numerous researchers and they have been used mainly as catalysts (Urquieta-Gonza`lez et al., 2002). Indeed, the percentage (14.1%) of loaded Fe corresponds to 500 meqFe/100 g zeolite, a value which is higher than the practical and theoretical CEC of the Clin, and is explained by the fact that one Fe3+ ion must compensate three spatially separated negative charges of the zeolite matrix. Because only a small amount of Fe3+ ions could be deposited at cationic sites, the additional Fe3+ ions most likely form Fe oxo- or hydroxo cations that undergo complex chemical transformation during subsequent washing. The Fe-overexchanged Clin can hold outside its framework one part of its Fe atoms in exchanged sites and the rest as neutral species. Amorphous extra framework metallic cations can be deposited on both internal and external surface. Generally, the Fe–overexchanged zeolites are characterized by the simultaneous presence of various types of Fe species, whose population depends on the procedure of Fe–zeolite preparation and it is evident that, in most cases, single di-

and trivalent Fe ions, oxo and hydroxo complexes, polymeric oxidic species and Fe-oxide species are present simultaneously (Chen et al., 1998; Ko¨gel et al., 1999; Doula, 2006).

3.2.

Equilibrium period

During the equilibrium period the chemical composition of the water samples was changed due to their contact with Clin and Clin–Fe system, as reported in Table 1. The two substrates changed the composition of the water samples; however the characteristics of these changes were different. The solution concentrations of Ca2+, Mg2+, Na+, K+ and framework Si/Al were slightly increased when Clin was used as a substrate. Especially for Na+ ions, the increase in their solution concentration is mainly due to the dissolution of zeolites impurities, but also due to ion exchange process. The limited release of Ca, Mg and K was assumed to occur through ion-exchange reactions, mainly between these surface cations and water molecules as well as with Na+ ions from impurities dissolution. Generally, this process is a dynamic one and occurs when aluminosilicates materials, such as Clin, are embedded in aquatic solutions. The motion of Si and Al, which are framework cations, from the substrate toward solution is a solid dissolution process and can cause a defect in zeolite framework structure and contamination of water samples. (Stumm, 1991; Oelkers and Schott, 1995). The new system not only retained its Ca and Mg, but also adsorbed these ions from the solution, decreasing the total hardness of the samples. The release of Na+ and the dissolution of Si were slightly increased during the equilibrium period but these increases were limited with respect to increases reported for Clin. The dissolution of Al was also limited and lower than in the water sample before treatment as well as in samples treated with Clin. The solution concentration of K+ ions was increased due to the dissolution of K+ ions deposited on Clin–Fe sites during the synthesis procedure, as will be better explained in the following. However, despite this process, the new system used its available active sites to retain Ca and Mg from solution as well as, water molecules. After the equilibration all the samples, treated with the Clin–Fe system, were characterized by basic pH, whereas the final solution pH values for Clin samples were stabilized within the neutral/basic pH range. The increase in solution pH corresponds to H+ adsorption by the two solids, whereas the decrease in solution pH to H+ release. By considering that the new system adsorbs larger H+ amounts, as well as experimental data from previous adsorption experiments (Doula, 2006) we conclude that the negative surface charge of the Clin–Fe system is noticeably higher than that of Clin. Because the K+ released concentrations are higher than the retained concentrations of the other cations, a final increase in sample conductivity values were measured.

3.3.

Zn2+ adsorption

In order to study the adsorption of Zn from the two substrates, an extent range of Zn initial concentrations was examined. Zinc concentrations used were exceedingly higher

ARTICLE IN PRESS WAT E R R E S E A R C H

1.4 1.2

Clinoptilolite Clinoptilolite-Fe system

X, mol/kg

1.0 0.8 0.6 0.4 0.2 0.00

0.005

0.010

0.015 C, mol/l

0.020

0.025

0.030

Fig. 1 – Adsorbed Zn concentrations (X) in relation to solution Zn concentrations (C) for all samples treated with clinoptilolite and clinoptilolite–Fe system.

100 Clin Clin-Fe system

90

% Zn Adsorption

80

2767

4 1 (200 7) 276 3 – 277 3

70 60 50 40 30 20 10 0 5.00 10.0 20.0 50.0 100 200 500 1000 1500 2500 Co ppm

Fig. 2 – Percentage adsorption of Zn for Clin and Clin–Fe system.

than the drinking water quality standard of 5.00 ppm but the aim of the study was to test the adsorption abilities and capacities of the two substrates under extreme conditions. Fig. 1 presents the adsorbed Zn amounts in relation to the solution Zn concentrations and Fig. 2 the percentage of adsorbed Zn concentrations for both substrates. Clin has a satisfactory adsorption behavior and is capable of adsorbing Zn2+ species from the contaminated samples. The adsorption graph consists of two regions however; it seems that the adsorption does not reach a plateau although the slope of the graph is gradually decreased. For low Zn initial concentrations (o10.0 ppm), Clin removes almost the entire concentration of the metal, whereas the percentage of adsorption decreases gradually with an increase in Zn initial concentration. The maximum adsorbed Zn amount by Clin, under our experimental conditions, is 1.09 mol/kg (or 71.3 mg/g). The Clin–Fe system can adsorb larger Zn amounts than its parent material (1.45 mol/kg or 94.8 mg/g). A two-region adsorption graph characterizes also the Clin–Fe system. The adsorption does not reach a plateau and the slope is gradually decreased. The same result was obtained by Trivedi and Axe (2001), who compared the adsorption of Zn2+ ions on amorphous Fe-oxide with that on goethite. They found that the adsorption of Zn adsorption by amorphous Fe-oxide does not reach a plateau whereas by goethite it does. From Figs. 1 and 2 it is obvious that when Zn2+ ion concentration is lower than 3.06  103 M (oro200 ppm), the newly synthesized system has the ability to retain almost the entire Zn2+ solution concentration. Table 2 summarizes the calculated QP values of the two substrates and for each one of the initial Zn concentrations. For Clin the higher values of QP were obtained within 5.00 and 100 ppm, while for the Clin–Fe system within 5.00 and 500 ppm. Beyond doubt Clin is a very good adsorbent material but it can be used only for low Zn initial concentrations. But since QP expresses the solution volume that can be purified, one concludes that the Clin–Fe system is capable of purifying larger solution volume than Clin. For Zn initial concentrations

Table 2 – Conductivities, initial and equilibrium pH, hardness and QP values for all water samples treated with Clin and Clin–Fe system after Zn adsorption Co (ppm)

5.00 10.0 20.0 50.0 100 200 500 1000 1500 2500

Clinoptilolite pH (initial)

PH (equil.)

6.74 6.83 6.84 6.74 6.70 6.62 6.38 6.17 5.95 5.64

7.48 7.48 7.30 7.06 6.74 6.71 6.35 6.19 6.09 5.98

Conductivity

289 306 338 405 555 866 1799 3240 4690 7290

Clinoptilolite–Fe system Hardness

QP

pH (initial)

PH (equil.)

107 111 116 127 173 205 268 296 312 297

96.3 91.8 85.7 75.3 66.6 52.9 41.6 34.6 30.2 28.5

6.79 6.90 6.91 6.79 6.75 6.66 6.40 6.19 5.96 5.64

8.39 8.37 8.20 7.94 7.46 6.73 6.17 5.94 5.85 5.74

Conductivity in mS/cm; hardness in ppm CaCO3 and Qp in l/kg.

Conductivity

391 407 440 535 691 994 1881 3320 4730 7180

Hardness

QP

28.8 33.2 40.6 65.8 113 225 326 462 477 472

99.1 100 103 99.8 103 96.5 68.6 48.1 43.3 37.8

ARTICLE IN PRESS 2768

WAT E R R E S E A R C H

41 (2007) 2763– 2773

between 100 and 500 ppm the Clin–Fe system can purify almost double the solution volume than Clin. It should also be pointed out that these results were obtained from only oneround sample treatment and significantly better results are expected, for high Zn concentrations, from an integrated water treatment process. Moreover, because retention values are dependent on the adsorbent dose, higher retention values, for both substrates, could be achieved by using higher adsorbents dose. From this point of view the Clin–Fe system could be used for the purification of heavily contaminated water samples or wastewaters. As mentioned, the Clin–Fe system has been already tested in the removal of Mn2+ and Cu2+ ions from aquatic solutions, as well as in the immobilization of heavy metals in soils (unpublished data). A general result that has been obtained and can be mentioned is that the Clin-Fe system is stable under a wide pH range and is not dissociated during heavy metals adsorption and thus, does not enrich the solutions with undesirable species. Moreover, it can adsorb higher metal concentrations than many other adsorbents which have been used to remove zinc from aquatic solutions. Table 3 presents maximum Zn adsorbed amounts by different substrates found in literature. Note that these values have been obtained from specific experimental conditions, expressing them and thus the comparison between the different sorbents could be only qualitative. These properties along with the fact that the new system has lower production cost than that of synthetic organic resins, often used for the same purpose, could characterize the Clin–Fe system as a very promising heavy metal adsorbent. The differences in chemical behavior, in adsorption of Zn2+ ions, in counterbalanced ions release and in Si/Al dissolution of the two substrates are owing to the different surface species, which are located on the two substrates. The new material, due to the presence of the Fe-oxides located in the zeolite channels or on its external sites, is characterized by

the presence of additional active sites (–Fe–OH), which are influenced by the solution pH and are potential adsorption sites (Cornell and Schwertmann, 1996). The presence of non-crystalline Fe formations located in cationic positions in the zeolite channels, of Fe binuclear and in general Fe complexes in extra-framework positions, as well as of amorphous Fe-oxides FeOx located at the surface of the zeolite crystal, gives to the Clin–Fe system higher SSA and thus, higher adsorption capacity than untreated Clin. (Pe´rezRamirez et al., 2002). It is also well-known that a poorly crystalline substrate is desirable for adsorption because the lack of a 3D crystalline structure results in high SSA (Trivedi and Axe, 2001; Dong et al., 2003). The higher adsorption detected for the Clin–Fe system is also due to its high negative surface charge (Doula, 2006, 2007). Zn2+ species can occupy the exchangeable sites of counterbalanced ions through an ion-exchange process and can form outer-sphere complexes (Stumm, 1991): ð S2O Þ2y C32n nþ þ Zn2þ #ð S2O Þ2y Zn2þ þ ð3  nÞCnþ ;

(2)

where C is the counterbalanced ion with charge n+(n ¼ +1 or +2), and S corresponds to the surface central metal (i.e. Si, Al, Fe). But metals can also form inner-sphere complexes with surface sites. During inner-sphere complexation, hydrogen ions are released as products, and the process causes a total decrease in solution pH (Stumm, 1991):  S2OH þ Zn2þ #  S2O2Znþ þ Hþ ;

(3)

 2S2OH þ Zn2þ #ð S2OÞ2 2Zn þ 2Hþ :

(4)

The work of May et al. (1986) emphasizes that except the above processes, the processes of heterogeneous nucleation cannot be ignored whenever solid phases are brought in contact with aqueous solutions. Since most specifically sorbed metal ions can form oxides and hydrous oxides, the

Table 3 – Maximum adsorbed Zn amounts on different, natural or synthetic, adsorbents used for the removal of zinc from contaminated aquatic samples Substrate

Maximum adsorbed Zn concentration, mg/g

Clinoptilolite from Gordes (Turkey) Clinoptilolite from Bigadic (Turkey) Na–montmorillonite Commercial zeolite 4A Zeolite 4A from coal fly ash Phosphatic clay (apatite+smectite) Iron hydrous oxide gel P–iron oxide Vermiculite Kaolinite Goethite Kaolinite–goethite system Bentonite Recycled iron-bearing material Clinoptilolite from Thrace (Greece) Clinoptilolite–Fe oxide system

6.00 3.00 3.61 140.8 140.1 25.1 5.86 9.00 39.2 1.99 5.29 2.30 52.9 22.0 71.3 94.8

(CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn, (CoZn,

¼ 20.0 mg/l, pH 4.00) ¼ 20.0 mg/l, pH 4.00) max ¼ 65.4 mg/l, pH 5.50) max ¼ 300 mg/l, pH 3.00) max ¼ 300 mg/l, pH 3.00) max ¼ 200 mg/l, 0.05M KNO3) max ¼ 10 mg/kg, pH 6.00) max ¼ 10 mg/kg pH 6.00) max ¼ 65.4 mg/l) max ¼ 65.4 mg/l, pH 7.00) max ¼ 65.4 mg/l, pH 7.00) max ¼ 65.4 mg/l, pH 7.00) max ¼ 300 mg/l) max ¼ 5.00 mg/l, pH 5.50) max ¼ 2500 mg/l, pH 5.64) max ¼ 2500 mg/l, pH 5.64) max max

Comax is the maximum initial Zn concentration used in experiments.

References ¨ ren and Kaya (2006) O ¨ ren and Kaya (2006) O Abollino et al. (2003) Hui et al. (2005) Hui et al. (2005) Singh et al. (2001) Ghanem and Mikkelsen (1988) Ghanem and Mikkelsen (1988) Fonseca et al. (2006) Nachtegaal and Sparks (2004) Nachtegaal and Sparks (2004) Nachtegaal and Sparks (2004) Mellah and Chegrouche (1997) Smith (1996) In this study In this study

ARTICLE IN PRESS WAT E R R E S E A R C H

4 1 (200 7) 276 3 – 277 3

potential for heterogeneous precipitation will always be present in aqueous systems because of the presence of the constituent OH ion as coordinated water molecules on the surfaces. Thus, the pH of the aqueous solution is an important controlling parameter in the sorption process because except the effect on the ionization degree of the surface groups of the adsorbent, it also influences metal speciation. Heavy metal ions (Mz+) may form complexes with inorganic ligands such as OH, and the extent of the complex formation varies with pH. Depending on the pH and metal concentration, zinc may form complexes with OH (ZnOH+, 2 and the precipitated hydroxide Zn(OH)2, Zn(OH) 3 , Zn(OH)4 Zn(OH)2 ) and as a result Zn–hydroxyl species may participate in the adsorption and precipitate onto the zeolite structure especially at pH above 8.00 (Lindsay, 1979; Hui et al., 2005). The extent of surface precipitation depends also on the sorbate concentration. Thus, according to Stumm (1991) at low sorbate cation concentrations, surface complexation is the dominant mechanism. As the sorbate concentration increases, the surface complex concentration and the mole fraction of the surface precipitate both increase until the surface sites become saturated. Referring to our experiment, despite the fact that the pH values of the samples were not adjusted during adsorption experiment, the pH values were precisely known for every experimental stage. After the equilibrium period the solution pH was stabilized at 7.53 for Clin and at 8.14 for Clin–Fe. Because the stock Zn solutions added after that period had different pH values due to their different Zn(NO3)2 content, H+ ions were added to samples. However, their quantity can be estimated and thus, one can also estimate the solution pH at the time of adsorption beginning (Table 2). The initial pH values ranged from 5.64 to 6.74 and despite that the solution pH was not constant for all samples, the samples with the same initial Zn concentration had almost equal pH values, meaning that the experimental results for each Zn concentration obtained for the two solids are comparable. Although solution pH at the beginning of adsorption does not favor precipitation, such a process could happen for samples with high Zn initial concentrations. However, by using a chemical equilibrium model (CHEAQS) it was proved that almost the entire quantity of Zn (499%) is present in solution as free Zn2+ ions for both substrates (Verweij, 2005). Also, according to Hui et al. (2005), because zeolites are not only influenced by solution pH but in turn are capable of affecting solution pH (especially in batch systems), they tend to have a higher internal pH. In addition, the zeolite surface may be influenced by the ambient pH which is not equal to the external solution pH value and precipitation within the channels and at the surface of zeolites may occur.

3.4.

Conductivity and samples pH

Table 2 presents the conductivities and pH values for all water samples after Zn2+ adsorption experiment. According to this data, Clin resists pH changes and after the adsorption process all the samples have pH values between 5.98 and 7.48. The sample conductivity values maintained are also low. Extre-

2769

mely high increase in sample conductivity was reported only for the four higher Zn concentrations. The range of solution pH values is wider for the Clin–Fe system than for untreated Clin, the higher pH value is 8.39 while the lower is 5.74. For Zn concentrations o200 ppm all the samples treated with Clin had lower solution pH than the samples treated with the Clin–Fe system, while the opposite phenomenon was reported for Zn concentrations 4200 ppm. Generally, zeolites and zeolite systems tend to neutralize the solutions, acting either as proton acceptors or as proton donors, exhibiting thus an amphoteric character. During Zn adsorption process the H+ ions participate in many reactions in the solution and in the solid phase. An explanation for the movement of H+ could be given by accepting that H+ ions from all sources (released, adsorbed and added) are constituents of the equilibrium: Hþ solid #Hþ aq :

(5)

H+aq

Thus, ions are adsorbed by the substrates, causing surface protonation:  S2OH þ Hþ aq #  S2OH2 þ ;

(6)

 S2O þ Hþ aq #  S2OH;

(7)

H+solid

while ions are released from the substrates as a result of Zn inner-sphere complexation (Eqs. (3) and (4)) or ion+ exchange processes (Eq. (2) if Cnþ 3n is H ). Hydrogen ions, + Haq, are added to the water samples along with Zn2+ solutions and their concentrations are increased as higher zinc concentrations are used. Moreover, H+aq ions are produced by Al3+ hydrolysis and during surface precipitation processes. Although the hydrolysis of Al3+ contributes to the final pH value this process is not significant since it is favored at high solution pH values, but high pH values have the samples with low Zn initial concentrations (Table 2) for which the dissolution of Al is limited (Fig. 5). By considering the decrease in solution pH values for both substrates, one may conclude that the two solids release H+ during Zn adsorption experiments. However, in order to calculate the real concentration of released/adsorbed H+, we have to factor into our calculations the concentration of H+ in equilibrated solutions (before and after adsorption), as well as the concentration of H+ added along with Zn stock solutions. After these calculations, it was confirmed that the adsorption of H+ is the predominant process for both solids, although H+ release was detected for Clin but only for the first Zn concentration (7.65  105 M). Thus, not only the concentration of the contaminator but also its solution pH value controls the reactivity of the substrates. Therefore, the two substrates by trying to neutralize the surrounding environments adsorb H+ ions from solutions and despite the final decrease in solution pH values; the calculations prove that the predominant process is the adsorption of H+ by both solids. If such a process does not happen, the solution pH of the samples will be much lower and obviously the samples will be unsuitable for consumption. Especially for the Clin–Fe system, it was calculated that it adsorbs larger amounts of H+aq than Clin for the six lower Zn concentrations. For the rest of Zn concentrations the adsorption of H+ is almost equal for the two sorbents. This behavior

ARTICLE IN PRESS 2770

WAT E R R E S E A R C H

41 (2007) 2763– 2773

is expectable because for low Zn concentrations (and low H+aq concentrations as well as limited movement of the counterbalanced ions) there is more negative charge available on the Clin–Fe system capable of adsorbing positive ions from solution.

2.80

3.5.

Movement of counterbalanced ions

Figs. 3 and 4 present the amounts of Ca2+, Mg2+, Na+ and K+ (in eq l1) released during Zn adsorption, as a function of initial Zn concentrations. They also present the overall release of

K Clinoptilolite

Na 2.40

Ca Mg

eq/l x 10-2

2.00

Total Desorption Total Adsorption

1.60 1.20 0.80 0.40 0.00 0

5

10

20

50 100 200 500 Initial Zn concentration, ppm

1000

1500

2500

Fig. 3 – Concentrations of released Na+, Mg2+, Ca2+, K+ ions, total adsorbed cations (H+ and Zn2+) and total desorbed cations during Zn adsorption by Clin.

2.80 Clinoptilolite-Fe system K

2.40

Na Ca

eq/l x 10-2

2.00

Mg Total Desorption

1.60

Total Adsorption 1.20 0.80 0.40 0.00 0

5

10

20

50 100 200 500 Initial Zn concentration, ppm

1000

1500

2500

Fig. 4 – Concentrations of released Na+, Mg2+, Ca2+, K+ ions, total adsorbed cations (H+ and Zn2+) and total desorbed cations during Zn adsorption by the Clin–Fe system.

ARTICLE IN PRESS WAT E R R E S E A R C H

4 1 (200 7) 276 3 – 277 3

positively charged ions from the two adsorbents, as well as the overall adsorption of positively charged ions (i.e., H+ and Zn2+). During adsorption stage, release of the counterbalanced Ca2+, Mg2+, Na+ and K+ ions takes place, partly as a consequence of Zn2+ adsorption. The retention of Zn2+ is undoubtedly the process, which determines and forces directly and indirectly the release of counterbalanced ions. By comparing the two figures, it is obvious that the behaviors of Na+ and K+ ions are significantly different for the two solids. Clin releases higher concentrations of Na ions and lower concentrations of K+ ions than Clin–Fe system. However, both released concentrations are constant during adsorption and equal to the concentrations released during equilibrium stage. Such an observation has been also reported by the authors for Cu adsorption by Clin (Doula et al., 2002) but also for Mn2+ adsorption by Clin and Clin–Fe system (Doula, 2006). It was suggested then, that almost the entire concentrations of Na and K were released during equilibrium stage and these ions did not participate in ionexchange reactions. Another reason for this constant release is that Clin contains impurities, which are dissolved and enriched the solution with Na+ ions. Owing to this dissolution the relation between released and retained equivalents of positive charges is not stoichiometric. For the Clin–Fe system the release of K+ is significantly higher than that of the other counterbalanced ions for Zn initial concentrations o200 ppm. The release of K+ ions maintains almost constant, with a small increase for the highest Zn concentration. The high release of K+ is owing to the synthesis procedure of the Clin–Fe system: the new system was in contact for 60 h with a 5 M KOH solution and thus, high concentrations of this ion were deposited on system sites, which were dissolved during equilibrium and adsorption stage. The observation that the concentrations of K+ and Na+ ions do not vary as a function of Zn adsorption is evidence that they neither influence nor are affected by the adsorption of Zn. For Zn initial concentrations o200 ppm the release of both Ca2+ and Mg2+ is higher from the surface of Clin than from the surface of the Clin–Fe system. Thus, the significant result obtained concerns the capability of the Clin–Fe system to adsorb high Zn concentrations from solutions and simultaneously to maintain the Ca and Mg solution concentrations low. As a consequence, the hardness of the water samples, treated with the Clin–Fe system, is significantly lower than the respective values for Clin when zinc initial concentrations were o100 ppm and almost the same for zinc concentration of 200 ppm (Table 2). By increasing initial Zn concentrations, the two solids are forced to adsorb even higher metal concentrations. Clin does not respond satisfactorily to this demand whereas, the newly synthesized system succeeds in adsorbing even larger Zn concentrations by simultaneously increasing Ca and Mg release. The amounts of released Ca and Mg from the surface of the Clin–Fe system are smaller than the adsorbed Zn and H+ amounts for all samples (Fig. 4). By considering that K+ and Na+ ions do not significantly affect the adsorption, one

2771

concludes that except ion exchange between Zn2+ ions and Ca2+/Mg2+ ions, there is also another type of Zn retention on system’s surface. Thus, it is possible that Zn ions form innersphere complexes with surface sites as described by Eqs. (3) and (4). The release of Mg2+ ions is almost constant for Clin, although there is a slight increase in Mg solution concentration as Zn retention becomes higher. On the contrary, the release of Ca2+ is noticeably high for Clin and it seems that the increase in its presence in equilibrated solutions is affected by Zn adsorption. However, the total amount of released Ca and Mg is higher than the adsorbed amounts of Zn for CoZno100 ppm, meaning that ion-exchange process is more likely to predominate for Clin when initial Zn concentrations are lower than 100 ppm and that part of the Ca, Mg release is owing to ion exchange between them and Zn from solution. As Zn initial concentrations increase, the adsorbed zinc amounts become larger than the total sum of released cations, confirming that inner-sphere complexation participate also in the retention of Zn2+ by Clin. Nevertheless, the fact that the inner-sphere complexation is more obvious for Co4100 ppm does not exclude the formation of such complexes for lower Zn concentrations. The dissolution of zeolite framework can also affect and control the presence of Ca and Mg (but also of other cations) in solutions. Under specific experimental conditions Si and Al from the framework move toward solution, and this process is characterized as dissolution and depends mainly on the solution pH, on the extent of surface protonation and on the nature of solution ions (Stumm, 1991). Generally, the dissolution of framework Si and Al causes local distraction of the framework and the release of more counterbalanced cations. Fig. 5 presents the concentrations of Si and Al found in solutions after Zn adsorption. The dissolutions of both Si and Al are higher for Clin than for Clin–Fe system and this is a possible explanation for the increased presence of Ca and Mg ions in solutions. Although the retention of low Zn concentrations by the two solids seems to be controlled mainly by ion-exchange (Fig. 3), it is also possible that Zn ions form inner-sphere complexes with surface sites. A method to confirm the formation of inner-sphere complexation is to carry out desorption experiments, because during this process the most available ions (outer-sphere complexed) enter in the solution while, the more stable inner-sphere complexes require more drastic conditions in order to break down their covalent bonds with surface and to enter into aquatic phase.

3.6.

Comparison between Mn2+ and Zn2+ adsorption

The characteristics of Zn2+ adsorption onto the two examined substrates were different as compared to adsorption characteristics of Mn2+ (Doula, 2006). Both solids are characterized by significantly higher adsorption capacity for Zn2+ than for Mn2+, and consequently all the other processes (movement of H, Ca, Mg, Na, K and Si/Al dissolution) are influenced by the different adsorbed metals amounts. The samples treated for Zn2+ removal have higher hardness and conductivities values because the higher zinc adsorption forces the counterbalanced ions to move towards solution

ARTICLE IN PRESS 2772

WAT E R R E S E A R C H

41 (2007) 2763– 2773

3.00 5.00

CAl, mol/l x 10-5

CSi, mol/l x 10-4

4.00

3.00

2.00

1.00

2.00

1.00

Clinoptilolite Clinoptilolite-Fe system 0.00

0.00 0.0125 0.025 Co, mol/l

0.0375

0.008

0.016

0.024

0032

0.040

Fig. 5 – Dissolution of framework Si and Al for Clin and Clin–Fe system vs. Zn2+ initial concentrations.

phase. Generally, zeolites, zeolite-mixed systems as well as other oxides or aluminosilicates adsorb larger Zn amounts than Mn and it seems that the Fe–overexchanged Clin of the present experiment follows this general rule. This preference has been reported by many researchers in the past and it was explained by the higher hydrated volume of Mn which inhibits its movement as compared to other smaller hydrated cations (Erdem et al., 2004; Hui et al., 2005; Fonseca et al., 2006).

4.

Conclusions

 After treatment with the Clin–Fe system most of the water



samples had significantly low hardness and it seems that, except the adsorption of Zn2+ ions, the new system acts simultaneously as a water softening material. It is also important that the same conclusion was obtained from the study of Mn removal from contaminated drinking water. The good behavior in adsorption experiments along with the fact that the Clin–Fe system is inexpensive, easily synthesized and harmless for human beings, as well as for the environment could characterize it as a very promising metal adsorbent.

 The Clin–Fe system used for the adsorption of Zn from







drinking water samples was synthesized by mixing Clin with aquatic solution of Fe(NO3)3 under strongly basic conditions (5 M KOH). The new system contains 14.1% Fe as amorphous Fe species and its Fe/Al ratio is equal to 1.23. It has a specific surface area (SSA) equal to 151.0 m2/g, which is significantly higher than the SSA of untreated Clin (30.98 m2/g). For one-round purification experiment and for the specific solid/sample ratio used (1/100) the results indicate that Clin has a satisfactory adsorption behavior. The maximum percentage of Zn adsorption reaches almost 100% when CoZn ¼ 5.00 ppm. For Zn concentrations between 5.00 and 20.0 ppm the percentage was also high (485%) whereas for zinc concentrations between 20.0 and 100 ppm the percentage was between 70% and 85%. For the experimental conditions used, the maximum Zn adsorbed amount was 71.3 mg/g. The Clin–Fe system adsorbed noticeably larger Zn amounts (maximum adsorbed concentration of 94.8 mg/ g) than its parent material. For Zn concentrations lower than 200 ppm the Clin–Fe system is capable of removing almost the entire Zn2+ solution quantity. The high Zn adsorption capacity of the Clin–Fe system is owing to Fe clusters located on its surface, high surface negative charge as well as, its high specific surface area.

R E F E R E N C E S

Abollino, O., Aceto, M., Malandrino, M., Sarzanini, C., Mentasti, E., 2003. Adsorption of heavy metals on Na–montmorillonite. Effect of pH and organic substances. Water Res. 37, 1619–1627. Chen, H.Y., Voskoboinikov, T., Sachtler, W.M.H., 1998. Reduction of NOx over Fe/ZSM-5 catalysts: adsorption complexes and their reactivity toward hydrocarbons. J. Catal. 180, 171–183. Cornell, R.M., Schwertmann, U., 1996. The Iron Oxides, Structure, Properties, Reactions, Occurrence and Uses. VCH Verlagsgesellschaft mbH, Germany, pp. 61–68. Dimirkou, A., Ioannou, A., Kalliannou, Ch., 1996. Synthesisidentification of hematite and kaolinite–hematite (k–h) system. Commun. Soil Sci. Plant Anal. 27, 1091–1106. Dimirkou, A., Ioannou, A., Doula, M., 2002. Preparation, characterization and sorption properties for phosphates of hematite, bentonite and bentonite–hematite system. Adv. Colloid Interface Sci. 97, 37–61. Doula, M., 2006. Removal of Mn2+ ions from drinking water by using clinoptilolite and a clinoptilolite–Fe-oxide system. Water Res. 40, 3167–3176. Doula, M., 2007. Synthesis of a clinoptilolite–Fe system with high Cu sorption capacity. Chemosphere 67 (4), 731–740. Doula, M., Ioannou, A., Dimirkou, A., 2002. Copper adsorption and Si, Al, Ca, Mg and Na release from clinoptilolite. J. Colloid Interface Sci. 245, 237–250.

ARTICLE IN PRESS WAT E R R E S E A R C H

4 1 (200 7) 276 3 – 277 3

Doula, M., Ioannou, A., 2003. The effect of electrolyte anion on Cu adsorption–desorption by clinoptilolite. Microp. Mesop. Mater. 58, 115–130. Dyer, A., 2001. Verified syntheses of zeolitic materials. In: Robson, H. (Ed.). Elsevier, New York, pp. 67–68. Erdem, E., Karapinar, N., Donat, R., 2004. The removal of heavy metal cations by natural zeolites. J. Colloid Interface Sci. 280, 309–314. Fonseca, M.G., Oliveira, M.M., Arakaki, L.N.H., 2006. Removal of cadmium, zinc, manganese and chromium cations from aqueous solution by a clay mineral. J. Hazardous Mater. 137, 288–292. Ghanem, S.A., Mikkelsen, D.S., 1988. Sorption of zinc on iron hydrous oxide. Soil Sci. 146, 15–21. Ghobarkar, H., Schaf, O., Guth, U., 1999. Zeolites—from kitchen to space. Prog. Solid State Chem. 27, 29–73. Dong, D., Deny, L.A., Lion, L.W., 2003. Pb scavenging from a freshwater lake by Mn oxides in heterogeneous surface coating materials. Water Res. 37, 1662–1666. Hui, K.S., Chao, C.Y.H., Kot, S.C., 2005. Removal of mixed heavy metal ions in wastewater by zeolite 4A and residual products from recycled coal fly ash. J. Hazardous Mater. B 127, 89–101. Inglezakis, V.J., Loizidou, M.M., Grigoropoulou, H.P., 2004. Ion exchange studies on natural and modified zeolites and the concept of exchange site accessibility. J. Colloid Interface Sci. 75, 570–576. Ko¨gel, M., Mo¨nning, R., Schwieger, W., Turek, T., 1999. Simultaneous catalytic removal of NO and N2O using Fe–MFI. J. Catal. 182, 470–478. Leinonen, H., 1999. Removal of harmful metals from metal plating waste waters using selective ion exchangers. In: Report Series in Radiochemistry, vol. 13, University of Helsinki, ISBN 951-458759-6 (PDF Version). Lindsay, W.L., 1979. Chemical Equilibria in Soils. Wiley, New York, p. 411. May, H.M., Kinniburgh, D.G., Helmke, P.A., Jackson, M.L., 1986. Aqueous dissolution, solubilities and thermodynamic stabilities of common aluminosilicates clay minerals. Geochim. Cosmochim. Acta 50, 1667–1677. Mellah, A., Chegrouche, S., 1997. The removal of zinc from aqueous solutions by natural bentonite. Water Res. 31, 621–629. Morturano, P., Drozdova´, L., Pirngruber, D., Kogelbauer, A., Prins, R., 2001. The mechanism of formation of the Fe species in

2773

Fe/ZSM-5 prepared by CVD. Phys. Chem. Chem. Phys. 3, 5585–5595. Nachtegaal, M., Sparks, D.L., 2004. Effect of iron oxide coatings on zinc sorption mechanisms at the clay–mineral/water interface. J. Colloid Interface Sci. 276, 13–23. Oelkers, E.H., Schott, I., 1995. Experimental study of anorthite dissolution and the relative mechanism of feldspar hydrolysis. Geochim. Cosmochim. Acta 59, 5039–5053. ¨ ren, A.H., Kaya, A., 2006. Factors affecting adsorption characterO istics of Zn2+ on two natural zeolites. J. Hazardous Mater. 131, 59–65. Pe´rez-Ramirez, J., Mul, G., Kapteijn, F., Moulijn, J.A., Overweg, A.R., Dome´nech, A., Ribera, A., Arends, I.W.C.E., 2002. Physicochemical characterization of isomorphously substituted FeZSM-5 during activation. J. Catal. 207, 113–126. Puls, P.W., Bohn, H.L., 1988. Sorption of cadmium, nickel and zinc by kaolinite and montmorillonite suspensions. Soil Sci. Soc. Am. J. 52, 1289–1292. Ruthven, M.D., 2001. Verified syntheses of zeolitic materials. In: Robson, H. (Ed.). Elsevier, New York, pp. 61–65. Schwertmann, U., Cornell, R.M., 1991. Iron Oxides in the Laboratory, Preparation and Characterization. VCH Verlagsgesellschaft mbH, Germany. Singh, S.P., Ma, L.Q., Harris, W.G., 2001. Heavy metal interactions with phosphatic clay: sorption and desorption behavior. J. Environ. Qual. 30, 1961–1968. Smith, E.H., 1996. Uptake of heavy metals in batch systems by a recycled iron-bearing material. Water Res. 30, 2424–2434. Stumm, W., 1991. Chemistry of the Solid–Water Interface. Wiley, New York. Trivedi, P., Axe, L., 2001. Ni and Zn sorption to amorphous versus crystalline iron oxides: macroscopic studies. J. Colloid Interface Sci. 244, 221–229. Urquieta-Gonza`lez, E.A., Martins, L., Peguin, R.P.S., Batista, M.S., 2002. Identification of extra-framework species on Fe/ZSM-5 and Cu/ZSM-5 catalysts typical microporous molecular sieves with zeolitic structure. Mater. Res. 5 (3), 321–327. Verweij, W., 2005. Chemical Equilibria in Aquatic Systems—CHEAQS-PC Calculating Program. /http://home.tiscali.nl/ cheaqs/index.htmlS. Zoumis, T., Calmano, W., Fo¨rstner, U., 2000. Demobilization of heavy metals from mine wastewaters. Acta Hydrochim. Hydrobiol. 28, 212–218.