Iron and aluminium oxides containing industrial wastes as adsorbents of heavy metals: Application possibilities and limitations.

584841

research-article2015

WMR0010.1177/0734242X15584841Waste Management & ResearchJacukowicz-Sobala et al.

Review

Iron and aluminium oxides containing industrial wastes as adsorbents of heavy metals: Application possibilities and limitations

Waste Management & Research 1–18 © The Author(s) 2015 Reprints and permissions: sagepub.co.uk/journalsPermissions.nav DOI: 10.1177/0734242X15584841 wmr.sagepub.com

Irena Jacukowicz-Sobala, Daniel Ociński and Elżbieta Kociołek-Balawejder

Abstract Industrial wastes with a high iron or aluminium oxide content are produced in huge quantities as by-products of water treatment (water treatment residuals), bauxite processing (red mud) and hard and brown coal burning in power plants (fly ash). Although they vary in their composition, the wastes have one thing in common – a high content of amorphous iron and/or aluminium oxides with a large specific surface area, whereby this group of wastes shows very good adsorbability towards heavy metals, arsenates, selenates, etc. But their physical form makes their utilisation quite difficult, since it is not easy to separate the spent sorbent from the solution and high bed hydraulic resistances occur in dynamic regime processes. Nevertheless, because of the potential benefits of utilising the wastes in industrial effluent treatment, this issue attracts much attention today. This study describes in detail the waste generation processes, the chemical structure of the wastes, their physicochemical properties, and the mechanisms of fixing heavy metals and semimetals on the surface of iron and aluminium oxides. Typical compositions of wastes generated in selected industrial plants are given. A detailed survey of the literature on the adsorption applications of the wastes, including methods of their thermal and chemical activation, as well as regeneration of the spent sorbents, is presented. The existing and potential ways of modifying the physical form of the discussed group of wastes, making it possible to overcome the basic limitation on their practical use, are discussed. Keywords Red mud, water treatment residuals, fly ash, adsorption, heavy metals, arsenic, iron and aluminium oxides waste materials

Introduction The presence of heavy metals in the aquatic environment tends to be of great concern in terms of human health and environmental impact, therefore they must be effectively removed from industrial effluents prior to their discharging into natural receiving waters. Several methods, e.g. precipitation, solvent extraction, membrane processes, ion exchange and adsorption, are used for this purpose. Among them, adsorption processes feature prominently because of their simplicity, low cost and high effectiveness in bringing down the concentration of the admixtures to a very low level. Much attention is devoted to adsorbents containing iron and aluminium oxides as active substances, which show very good adsorption properties towards the heavy metals and semimetals, in the form of both cations (e.g. cadmium, zinc, lead, chromium(III), copper, nickel) and anions (arsenates, chromates or selenates), present in natural waters and in processing solutions. To date, several adsorbents based on aluminium or iron oxides (e.g. activated aluminium oxide, granular iron hydroxide, iron oxide-coated sand or activated carbon, zero-valent iron, hybrid polymers containing iron oxides permanently deposited in the structure of their macromolecules, as well as iron oxide

nanoparticles with magnetic properties) have been developed (Allahdin et al., 2014; Chung et al., 2014; Cumbal et al., 2003; Fu et al., 2013; Giles et al., 2011; Han et al., 2009; Hua et al., 2012; Kumar et al., 2014; Mak et al., 2011; Üzüm et al., 2009; Xu et al., 2012). Some of them are already used in industry. Recently the possibility of utilising (as adsorbents) industrial wastes containing iron or aluminium oxides, particularly wastes from water treatment processes and from the processing of bauxites, has received much attention. Such wastes are produced in great quantities and tend to be deposited, which is at odds with the current approach to industrial waste management which emphasizes waste generation prevention or effective utilisation.

Department of Industrial Chemistry, Wroclaw University of Economics, Wrocław, Poland Corresponding author: Irena Jacukowicz-Sobala, Department of Industrial Chemistry, Wroclaw University of Economics, ul. Komandorska 118/120, 53-345 Wrocław, Poland. Email: [email protected]

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Waste Management & Research

Most of the relevant research deals with the basic characterisation of the produced wastes – their physical and chemical structure, crystalline forms, sorptivity towards selected elements and the possibility of their multiple use. Much attention is also devoted to the possibilities of the chemical or thermal modification of crude wastes aimed at improving their sorptivity towards selected groups of pollutants present in aqueous solutions.

a)

O Me

≡ FeOH + An − ↔ ≡ FeAn + OH − ≡ Fe ( OH )2 + An − ↔ ≡ Fe2 An + + 2 OH − In the case of cations, specific adsorption consists in the formation of coordinate bonds with a deprotonated hydroxyl

L

b) O

O Me

Mechanisms of anion and cation adsorption on oxide sorbents The adsorption of both anions and cations on the surface of iron or aluminium oxides can proceed in accordance with two basic mechanisms referred to as specific and nonspecific adsorption (Cornell and Schwertmann, 2003). Nonspecific adsorption is based mainly on electrostatic interactions between the sorbent surface with a given electric charge and the ions present in the solution. Since the adsorbed ions preserve their own hydration envelope, the characteristic feature of nonspecific adsorption is the presence of at least one layer of water molecules between the sorbent surface and the adsorbed ion. The adsorption of anions proceeds most effectively in low pH conditions because of the occurrence of a large number of positively charged active sites (e.g. ≡FeOH2+). As the pH increases, the total positive charge of the oxide surface decreases, but at the same time the number of negatively charged dissociated molecules of the adsorptive increases. Thus the pH optimum for the adsorption of anions is within a range between the pKa1 of a proper acid and pHpzc (the point of zero charge) of the sorbent. Above this range the sorbent surface acquires an overall negative charge, whereby the nonspecific adsorption of anions practically does not occur. These conditions are theoretically favourable for the adsorption of cations, but most heavy metals already precipitate below the pHpzc of such sorbents in the form of hydroxides, making it necessary to conduct the sorption process in an acidic environment in which the surface of iron and aluminium oxides has a positive charge. For the removal of anions and cations by means of metal oxide adsorbents specific adsorption (chemisorption, innersphere adsorption), in which coordinate bonds forming between the adsorbate and a metal atom on the adsorbent surface play the dominant role, is of greater importance. The proper active centres are hydroxyl groups coordinatively bonded with a single iron or aluminium atom (singularly coordinated groups). This mechanism is called ligand exchange since the bonding of anions on the adsorbent surface is accompanied by the release of hydroxyl groups into the solution, which contributes to the alkalisation of the environment (Cornell and Schwertmann, 2003):

L

L

c) O

O

Me

Me

Figure 1. Kinds of complexes forming on surface of iron or aluminium oxides: (a) monodentate mononuclear, (b) bidentate mononuclear, (c) bidentate binuclear. Me: Fe(III)

or Al(III), L: ligand.

group, which is accompanied by the acidification of the environment and, similarly as in the case of anions, by the formation of mono- or binuclear surface complexes (Cornell and Schwertmann, 2003): z −1 + ≡ FeOH + M z + ↔≡ FeOM ( ) + H +

≡ Fe ( OH )2 + M z + ↔ ≡ FeO) 2 M (

z − 2 )+

+ 2 H+

The kind of complexes (Figure 1) depends on the place on the surface of iron or aluminium oxides in which the adsorption process proceeds and on the process conditions. In the case of iron oxides, the most stable and thermally privileged are bidentate binuclear complexes forming in the corners of adjacent regular octahedrons (the basic structural components of most iron oxides) (Farquhar et al., 2002; Guo et al., 2007). Places situated along the edges of adjacent regular octahedrons favour the formation of bidentate mononuclear complexes in the initial phase of adsorption when the surface is only slightly coated with sorbate (Cornell and Schwertmann, 2003; Fendorf et al., 1997; Manceau, 1995). The proportion of the number of individual active sites favouring the formation of specific complex bonds is characteristic of the different structural forms of iron and aluminium oxides, and to a large extent determines their adsorptive properties. The anions and cations adsorbing, according to the above mechanism, bond permanently with the surface of iron or aluminium oxides, modifying it as a result, whereby the pHpzc of the sorbent changes. At the same time, owing to the covalent character of the forming bonds, the adsorption of both cations and anions at a neutral surface charge of the sorbent, and even at same-sign surface and adsorbate charges, becomes possible. Thus, the specific adsorption of anions is possible in an environment whose pH is higher than the pHpzc of the sorbent, while the adsorption of cations can proceed in a slightly acidic environment in which their solubility is the highest. Also, the


3

Jacukowicz-Sobala et al. dependence between the effectiveness of the adsorption of cations on iron or aluminium oxides and the pH of the solution (the so-called adsorption edge) has a peculiar character. In a strongly acidic environment the positive charge of the sorbent surface makes cation adsorption practically impossible, but as the pH increases and the surface charge gradually changes to negative, the efficiency of the process increases very quickly in a rather narrow (1–2 pH units) range characteristic of the particular cations. Moreover, at higher solution pH values cations are adsorbed in a hydrolysed form (e.g. PbOH+), and as the environment acidity further decreases, they precipitate in the form of hydroxides on the sorbent surface, whereby the heavy metal cation removal effectiveness reaches nearly 100% in solutions with a pH lower than the pHpzc of the sorbent (Zhou and Haynes, 2011). Not only specific and nonspecific adsorption, but also other mechanisms, such as the co-precipitation or permanent immobilisation (intercalation), in the structure of the minerals present in the adsorbent play an important role in the removal of toxic admixtures from waters by means of iron and aluminium oxides. Also the so-called reductive dissolution that accompanies the adsorption of arsenates(III) on the surface of iron oxides is of significance since the Fe2+ ions released in the process adsorb on the sorbent surface, forming new positively charged active centres, whereby the total sorption capacity increases and efficient adsorption of anions in the alkaline environment becomes possible (Greenleaf et al., 2003; Leupin and Hug, 2005; Nagar et al., 2010).

Sorptive properties of industrial wastes containing iron and/or aluminium oxides In the huge mass of iron and aluminium oxides containing wastes generated by the different branches of industry, one can distinguish the following three main groups: water treatment residuals (WTRs), bauxite processing wastes (red mud) and fly ashes from the combustion of coal and biomass. The wastes are generated in very large quantities (hundreds of billion kilograms per annum) and within the particular groups they show similar physicochemical properties. However, the groups of wastes, despite their common feature – a high iron and aluminium oxide content – significantly differ in their physicochemical properties. Therefore, depending on their origin, they require different processing and modification methods to endow them with desirable adsorption material characteristics. Still, there are literature reports, according to which, studies of heavy metal adsorption by sorbents obtained from raw materials belonging to the different groups mentioned above were carried out in the same conditions to compare their sorption capacity and draw general conclusions concerning their intended use and way of processing aimed at endowing them with better adsorption and utilitarian properties.

Bauxite ore processing wastes (red mud) Bauxite ore processing wastes (red mud) are generated in the first stage of aluminium production, when aluminium oxide is separated from bauxite. This stage is usually conducted using the Bayer method, consisting in the extraction of aluminium oxide from bauxite ore with caustic soda at elevated temperatures and under increased pressure. In the next stage (the Hall-Heroult process) the obtained aluminium oxide (with an addition of cryolite) is subjected to electrolytic reduction to aluminium (Saternus, 2006). The wide range of applications of aluminium for construction purposes in the transport industry, the automotive industry, the aviation industry, the building industry and the packaging industry, and consequently its rising production, contribute to the generation of great quantities of red mud, estimated at about 120 billion kilograms per annum, most of which is deposited on the bottom of seas or in land-based impoundments (Power et al., 2011). Depending on the composition of the raw material, bauxite processing wastes contain mainly iron oxides, aluminium oxides, silicon oxides, titanium oxides, sodium oxides, potassium oxides, vanadium oxides and gallium oxides. According to the relevant EC directive, after its neutralisation red mud does not constitute a hazardous waste. Under the trade name Bauxsol (red mud neutralised with sea water) it is used by industry (Power et al., 2011). Its adverse effect on the environment stems from the scale of the generation and deposition of this waste characterised by a very high pH (10.5–12.5) and high grain fineness, which may result in the alkalisation of underground waters and in dust emissions into the atmosphere, adversely affecting the flora. The management of the huge quantities of bauxite processing wastes generated per annum (the production of 1 t of aluminium generates 0.9–1.76 t of red mud) entails the high cost of their deposition, increasing with the area of the land designated for their disposal, with the cost of operating the disposal sites and with the environment use charges (Power et al., 2011; Rai et al., 2012; Sutar et al., 2014). Wastes generated in the Bayer process. The processing of bauxites by the Bayer method consists of several unit operations: bauxite digestion, desilication, aluminium compound leaching, the clarification of the sodium aluminate solution, from which by lowering its temperature aluminium hydroxide is precipitated as a result of NaAlO2 decomposition, and further operations involved in the treatment of the solid residue (red mud). The preliminary desilication consists in dissolving readily soluble silica (which in bauxites usually occurs in the form of kaolin) and precipitating the desilication product in the form of sodium aluminosilicates: sodalite or cancrinite. This process proceeds in a dilute solution of NaOH at a temperature of 95 °C–100 °C, and its aim is to reduce the silicate content in the sodium aluminate solution obtained in the next stage and to improve the quality of the final product. The main stage in the Bayer process is the leaching of aluminium oxide from the preconditioned bauxite, which is conducted in the strongly alkaline medium of concentrated NaOH solution. The amphoteric character of aluminium compounds and their greater solubility (in comparison with the other raw material


4


components) in these conditions are exploited in the process. The sodium aluminate extraction conditions depend on the mineralogical composition of the ore. In the case of gibbsite (γ-Al(OH)3), NaOH solutions with a concentration of 3.6–8.9 M, a temperature of 104 °C–145 °C and a pressure of 0.1–0.3 MPa are used while bauxites containing boehmite (γ-AlO(OH)) are digested with 3.6–5.0 M NaOH solutions at a temperature of 200 °C–232 °C and under a pressure of 0.6 MPa. The extraction of diaspore (α-AlO(OH)) is conducted at an even higher temperature, i.e. at 240 °C–260 °C. The extraction process and the separation of the sodium aluminate are followed by the stages involved in the preparation of the red mud for its deposition, including washing, thickening and neutralisation. The aqueous washing is conducted in the countercurrent system in a series of clarifiers/thickeners using polyacrylic or polyacryloamide flocculants in order to wash away the excess of sodium aluminate and the excess of the leaching agent, and at the same time to densify the sludge. The final stage most often is the thickening of the sludge in superthickeners to the required solid phase content (to 55%) depending on the way of its deposition (lagooning, dry stacking). Much less often, filtration is additionally used for this purpose. Filtration combined with thickening makes it possible to obtain a slurry paste with a solid phase content above 65% (dry cake disposal). Properties and management possibilities. The chemical composition of red mud depends on the composition of the raw material (bauxite) and on the course of its processing. Wastes from different aluminium oxide production plants in the world have the same basic components, but they considerably differ in their content: Fe2O3 (30%–60%), Al2O3 (10%–20%), SiO2 (3%–50%), Na2O (2%–10%), CaO (2%–8%), TiO2 (trace amounts–25%). Exemplary chemical compositions of the residues from the processing of bauxites coming from different regions in the world are shown in Table 1. The mineralogical composition varies even more: besides the minerals originally occurring in the raw materials (e.g. hematite, geothite, boehmite, diaspore, anatase, rutile, quartz and calcite), new crystalline phases (e.g. sodalite, cancrinite, portlandite and perovskite) form in the course of the Bayer process (Table 2) (Rai et al., 2012). The average crystalline and amorphous phase content is estimated at respectively 70 wt% and 30 wt% (Gräfe et al., 2011). Most of the generated bauxite processing wastes are managed by depositing them in the environment. About 30% of the major aluminium oxide producers discharge the wastes directly to seas, while the others deposit them (in the form of more or less thickened sludge) in sealed and drainage-protected basins or natural depressions. The simplest method consists in depositing slurry with a low dry mass content. But this solution poses the greatest uncontrolled leakage and groundwater alkalisation hazard and requires large areas for building waste deposition basins. Today, more and more producers deposit red mud in a concentrated form, containing 48%–55% of solid phase, or in the form of paste in which the dry mass content amounts to 65% and in some cases exceeds 75%. The preferred dry stacking possesses less leakage hazard and needs less deposition area, but requires more protection against the spread of alkaline dust from the waste disposal site.

Some aluminium oxide producers, before depositing bauxite processing residues, subject them to partial neutralisation, usually using sea water or CO2. Attempts are also made to use such agents as SO2 containing exhaust gases, mineral acids or acidic industrial effluents. As opposed to acidic agents that neutralise the OH− ions in the aqueous phase of red mud, the action of sea water consists in the precipitation of the anions, which mainly determine its strongly alkaline properties (OH-, CO32-, Al(OH)4-), in the form of insoluble calcium and magnesium compounds. These methods do not ensure the complete neutralisation of the alkalinity of the waste, but merely its reduction to a pH of 8.5– 10.5 depending on the initial pH and the neutralising agent used (Power et al., 2011; Rai et al., 2012; Sutar et al., 2014). Because of the scale of red mud generation, numerous attempts are undertaken to utilise this waste as a building industry raw material for the production of bricks, cement, concrete, paints and pigments, and as the main component of ceramic masses or glazes in the ceramic industry. Also, studies on the possibility of recovering the main bauxite processing waste components such as iron, aluminium and titanium, and admixtures such as zinc, cadmium, vanadium and lead, or rare earth metals such as scandium, thorium and uranium, are being conducted. Many studies are devoted to the utilisation of this waste as a support and an active substance in chemical catalysis and as a coagulating agent or (because of its iron and aluminium oxide content) as an adsorbent for the cleaning of flue gases, water and industrial wastes in environmental protection (Bertocchi et al., 2006; Rai et al., 2012; Sutar et al., 2014; Wang et al., 2008). Owing to its ability to permanently bind heavy metals, red mud can be used as reactive permeable barriers or as an additive to deposited wastes (mainly mine wastes, but also municipal wastes) in contaminated soil and industrial waste remediation (Bertocchi et al., 2006; Klauber et al., 2011). Much attention is also devoted to the possibilities of utilising red mud as a heavy metals adsorbent in industrial sewage treatment. In comparison with other sorbents, it is characterised by a relatively poorly developed specific surface area (typically 10–25 m2 g-1). Nevertheless, in the literature on the subject, one can find examples of the use of red mud of different origins (Jamaica; Australia, Kwinana; Weipa, Eurallumia; Vietnam, Tan Binh Alumina (Gräfe et al., 2011; Ha et al., 2011), whose specific surface area often exceeds 50 m2 g-1). Moreover, thanks to thermal or chemical activation one can obtain sorbents with an even larger specific surface area, reaching a few hundred m2 g-1. Several red mud-based sorbents for the removal of toxic admixtures from aqueous solutions, mainly cations of heavy metals such as Zn(II), Cu(II), Pb(II), Cd(II), Ni(II), toxic anions of mainly As(III), As(V) and Cr(VI), but also V(V), Mo(V), Se(IV) and Se(VI) compounds, as well as radionuclides of radium, uranium and thorium, have been obtained (Clark et al., 2011a, 2011b). The sorbents exhibit different sorptivity (from a few to tens milligrams of the admixture removed per gram of sorbent) depending on the origin of the red mud, the way of neutralising it (in the Bayer process) and the method of modification aimed at obtaining a material with the desired sorptive properties (Table



a,bThe

3.0

12.6a +

31.0b

+

−

3.0

2.0

9.58

16.9

+

9.64

33.44 13.4 7.0 7.0

SiO2

rest of the mass is cancrinite (Na6Ca1.5Al6Si6O24(CO3)1.6): (a) 51%, (b) 42%.

3.0

12.0

27.0a

+

8.61

5.68

17.91

17.6

37.3

−

18.8

6.88 12.2 3.5 3.5

TiO2

30.45

+

+

Eurallumina, Portovese, Sardinia, Italy Eurallumina, Portovese, Sardinia, Italy Queensland Alumina Ltd Gladstone, Australia

17.28

38.80

HONDALCO, Renukoot, India Etibank Seydisehir Aluminium, Konya, Turkey Etibank Seydisehir Aluminium, Konya, Turkey

18.44 12.8 13.0 13.0

Al2O3

16.02 24.0 54.0 54.0

Fe2O3

Chemical composition, %

Shandong Aluminium Corp. China Shandong Aluminium Corp. China NALCO, Damanjodi, Orissa, India NALCO, Damanjodi, Orissa, India

Plant

+

−

−

+

−

7.77

4.4

11.54 13.0 − −

CaO

+

−

−

12.06

−

6.86

5.24 6.2 8.0 8.0

Na2O

Table 1. Composition and main properties of bauxite refining residuals (red mud).

Neutralisation with sea water

Neutralisation with sea water and acid treatment

Neutralisation with sea water Neutralisation with sea water

Acid and thermal treatment

Rinsing with FeCl3 solution

Granulation Rinsing with FeCl3 solution Neutralisation with CO2 Neutralisation with CO2 and thermal treatment Thermal treatment

Residuals treatment

14.2

25.2

18.9

31.8

28.0

−

108.0

15.3 192.0 63.0 68.22

SBET, m2 g-1

Gupta et al., 2001 Kocabaş and Yürüm, 2011 Apak et al., 1998 Bertocchi et al., 2006 Santona et al., 2006 Santona et al., 2006 Zhou and Haynes, 2012

88.2 mg Pb(II)/g (pH 4.0) 75 mg Cr(VI)/g (pH 2.0) 5.25 mg As(III)/g (pH 7.0)

389 mg Pb(II)/g 151 mg Cd(II)/g 162 mg Zn(II)/g 389 mg Pb(II)/g 151 mg Cd(II)/g 162 mg Zn(II)/g 11.2 mg As(III)/g 30.7 mg As(V)/g 15.8 mg Se(IV)/g 30.8 mg Se(VI)/g

11.64 mg As(V)/g (pH 2.0) 112.0 mg Cd(II)/g; 87.8 mg Cu(II)/g; 137.2 mg Pb(II)/g (pH 5.7) 0.95 mg As/g (pH 6.5)

Zhu et al., 2007 Zhang et al., 2008 Sahu et al., 2010 Sahu et al., 2011

References

52.10 mg Cd(II)/g 0.95 mg As/g (pH 6.5) 55.55 mg As(V)/g (pH 5.0) 14.92 mg Zn(II)/g (pH 6.0)

Sorption capacity, mg g-1)

Jacukowicz-Sobala et al. 5

6


Table 2. Mineralogical composition of the bauxite refining residuals from various plants. Plant

Minerals

Content, %

References

Etibank Seydisehir Aluminium, Konya, Turkey

Sodalite, Na2OxAl2O31.68 SiO21.73H2O Hematite, α-Fe2O3 Cancrinite, 3NaAlSiO4NaOH Diaspore, AlO(OH) Rutile, TiO2 Calcite, CaCO3 Rutile TiO2 Hematite, α-Fe2O3 Magnetite, Fe3O4 Ilmenite, FeTiO3 Sodalite, Na2OxAl2O31.68 SiO21.73H2O Hematite, α-Fe2O3 Gibbsite, Al(OH)3 Sodalite, Na2OxAl2O31.68 SiO21.73H2O Quartz, SiO2 Cancrinite, 3NaAlSiO4NaOH Rutyle and anatase, TiO2 Hydroxides and bicarbonates (Ca, Al) Hydroxides and bicarbonates (Mg, Al) Calcite and aragonite, CaCO3 Halite, KCl

32.3 34.9 4.6 2.5 1.5 1.2 >50 30-50 3-10 3-10 trace 35.0 20.0 20.0 8.0 8.0 6.0 6.0 5.0 4.0 1.7

Altundogan et al., 2002

Alumina-Aluminio of San Ciprian, Lugo, Spain

Eurallumina, Portovese, Sardinia, Italy

1). The neutralisation and modification of red mud have a direct bearing on the sorbent suspension reaction, and in addition, they usually lead to phase transitions of its components. Since these minerals show different values of pHpzc (e.g. pHpzc amounts to 8.7–9.8 for hematite, 6.8 for magnetite, 5.0 for gibbsite and 4.0 for rutile), this results in a change in the magnitude of the surface charge and in the efficiency of the adsorption process in different pH conditions (Power et al., 2011). Modification of sorptive properties. In order to utilise bauxite processing residues as an adsorbent, they must be neutralised at least after the Bayer process. Attempts have been made to remove As(III) and As(V) compounds using red mud directly from a superthickener, but the obtained sorptivity values were very low, amounting to 0.380 mg As(V)/g at pH = 3.2 and 0.323 mg As(III)/g at pH = 9.0. At the same time the sorbent would show affinity to arsenates(III) and arsenates(V) in narrow pH ranges beyond which its sorptivity would sharply drop (Altundogan et al., 2000). The neutralisation of bauxite processing wastes alone considerably improves their sorptive properties as a result of a change in the chemical structure in their surface. The varied mineralogical structure of red mud is characterised by the presence of minerals exhibiting a constant negative or positive charge and minerals containing amphoteric groups whose charge depends on the pH of the environment. As a result of the neutralisation of red mud with sea water, the anions responsible for its alkalinity precipitate in the form of various hydrotalicites (Mg4Al2(CO3)(OH)12xnH2O, Mg8Al2Cl(CO3)0,5(OH)20xnH2O). These minerals have a constant positive charge whereby they increase the number of active sites showing affinity to anions (Palmer et al., 2009). Red mud neutralised (to pH = 10.7) with sea water showed substantially better sorptivity towards As(III) and As(V) compounds, as well as towards Se(IV) and Se(VI)

Lopez et al., 1998

Clark et al., 2011a

compounds, amounting to 11.85 mg g-1, 30.75 mg g-1, 15.80 mg g-1 and 30.81 mg g-1, respectively (Zhou and Haynes, 2012). As a result of the next stages in the modification of the chemical and physical properties of red mud, the obtained sorbents have much better sorptive properties. Among the activation methods the following stand out: washing with Fe(II) or Fe(III) salt solutions, washing with hydrochloric acid or sulphuric acid and heating at a temperature of 500 °C–600 °C or at 900 °C, as well as a multistage treatment being a combination of the above methods. The aim of chemical treatment with iron(II) or (III) salts is to lower the pH of the red mud slurry and to precipitate additional quantities of hydrated iron oxides and iron hydroxyoxides on the surface of the solid phase. By modifying the sea water-neutralised wastes with an FeCl3 solution, a product whose specific surface area (192 m2 g-1) was larger than that of the starting material (115 m2 g-1) was obtained. At the same time the iron content increased from 10.6% to 17.5%, while the calcium content decreased from 4.2% to 2.0%. An X-ray diffraction analysis showed that the mineralogical composition of the sorbent changed as a result of the activation process, i.e. the calcite content and the silicate minerals (quartz and sodalite) content decreased while the hematite content increased (Kocabaş and Yürüm, 2011; Zhang et al., 2008). Probably owing to the higher iron oxide content and the lower sodalite content (characterised by a constant negative charge), the obtained sorbent exhibited relatively high sorptivity towards As(V) compounds (68.5 mg g-1 and 23.2 mg g-1 at a pH of 6.0 and 9.0, respectively) (Gräfe et al., 2011; Zhang et al., 2008). Continuing their research on the adsorption of arsenate(V), the authors would contact red mud of the same origin with FeCl2 solutions. The obtained sorbent showed a 10% higher capacity to remove As(V) from aqueous


7

Jacukowicz-Sobala et al. solutions than the previous product. The highest sorptivity was obtained at a pH of 4.0–8.0. A more acidic environment is usually the environment most favourable to sorbents containing iron and aluminium oxides. This is owing to the properties of Fe(II) salts, which at a pH > 4.0 undergo oxidation and hydrolysis, resulting in the precipitation of hydrated iron oxides. Consequently, the removal of As(V) on the sorbent modified in this way, would proceed via specific adsorption on the surface of iron oxides, and in addition through the coprecipitation of iron and arsenic compounds. At the same time, iron(II) salts would aid the flocculation and sedimentation of red mud particles after the adsorption process (Li et al., 2010). The activation of bauxite processing residues with acids is usually conducted at elevated temperatures using sulphuric acid solutions or hydrochloric acid solutions with a concentration of 0.2%–20%. The process results in the partial dissolution of the components of red mud, whereby its mineralogical composition and surface charge change and its specific surface area increases. Similarly, as in the case of the previous methods, the properties of the obtained products depend on the composition of the raw material and on the activation process conditions, i.e. the concentration of the acid, the process temperature and the way of conducting the process (the batch method or the column method). Studies (Altundogan et al., 2002) showed that when acted on with low-concentration hydrochloric acid, red mud slurry would acquire colloidal properties, and when the sorbent surface (particularly its active oxygen centres) was covered with silicic acid, its sorptivity towards arsenates would decrease. At higher HCl concentrations (2%–3%), aluminosilicates, sodalite (Altundogan et al., 2002) and cancrinite (Santona et al., 2006) would dissolve, whereby the permanent negative surface charge would decrease while the hematite content would relatively increase, and simultaneously the additional hydroxyl groups on its surface would become exposed, which would result in an increase in the sorptivity of the red mud towards As(III) and (V) compounds from 0.664 and 0.514 mg g-1 to 0.885 to 0.942 mg g-1, respectively. The further increase in the concentration of the hydrochloric acid used in the treatment would cause the dissolution of small particles of the sorbent, a reduction in its specific surface area and consequently, a deterioration in its sorptive properties (Altundogan et al., 2002). As opposed to the removal of anions, the acid activation of red mud did not result in an increase in its sorptivity towards heavy metal cations, which is owing to a different mechanism of their binding on the sorbent surface. The presence of aluminosilicates in the form of sodalite or cancrinite is desirable in this case because of the low solubility of metal cations in neutral and alkaline environments and the necessity of conducting the adsorption process in acidic solutions (often below the pHpzc of such materials as hematite or aluminium and titanium oxides). In experiments on lead, zinc and cadmium adsorption with sea water-neutralised bauxite processing residues, the sorptivity values obtained were 389, 162 and 151 mg g-1, respectively, whereas the sorbent after activation with the HCl solution showed nearly half lower capacities to remove the

cations. Hematite and cancrinite were found to be the main phases participating in their adsorption. On the hematite surface, the sorption of the cations involved specific bonds in the form of inner-sphere complexes, while in the case of cancrinite the dominant mechanism was ion exchange on its outer surface and more permanent intercalation inside its crystalline structure (Santona et al., 2006). The results of many studies indicate that red mud exhibits much higher sorption capacity towards cations of heavy metals than towards anionic pollutants such as arsenates or chromates (Table 1). This is owing to the fact that heavy metals precipitate in the form of hydroxy-oxides, which show high affinity to the sorbent surface. The share of this phenomenon is dominant when the adsorption process is conducted at a pH slightly below the precipitation limit of the metal being removed. Hence there is a clear correlation between the sorptivity towards cations of such metals as copper, lead and cadmium and their solubility products. At the same time the layer of metal hydroxide favours the adsorption of subsequent portions of cations, which results in considerably higher capacities for purifying solutions in column processes than in batch processes (Apak et al., 1998). Thermal activation is the most popular way of modifying red mud. As a result, the latter undergoes stabilisation, its specific area considerably increases and its mineral components undergo phase transitions. The operations that precede the thermal treatment (washing with acids or with H2O2 solutions to remove organic compounds) and the selection of a proper process temperature make it possible to obtain materials with the desired properties (Ha et al., 2011). Thermogravimetry (TG) and X-Ray diffraction (XRD) analyses showed that the heating of red mud to a temperature of about 310 °C leads to the evaporation of free moisture and to the dehydration of gibbsite (Al(OH)3) to aluminium oxide (Tong et al., 2013). At this temperature the removal of water from between the layers of the hydrotalcyte minerals precipitated during the neutralisation of the wastes with sea water also takes place. After this modification, red mud exhibits not only a permanent positive surface charge, but also has free spaces in its crystalline structure enabling the intercalation of anions, which contributed to an increase in its sorptivity towards arsenates, vanadates and molybdates (Palmer et al., 2010). At higher temperatures (311 °C–600 °C) geothite would change into hematite, while in the temperature range of 600 °C–700 °C calcite would undergo decomposition with the release of carbon dioxide. The gaseous products released during the transitions cause the development of the porous structure of the sorbent and a severalfold increase in its specific surface area, e.g. from 14.0 to 48.5 m2 g-1 or from 68.0 to 275.0 m2 g-1 (Ha et al., 2011; Tong et al., 2013). The heating of red mud at temperatures exceeding 600 °C causes a reduction in its surface area owing to the formation of meso- and macropores, and at the same time, as a result of sintering, the properties of its surface change, which often leads to a reduction in sorptivity (Ha et al., 2011). Sorbents obtained as a result of thermal activation usually show lower sorptivity towards metal cations than red mud merely neutralised with sea water. Bauxite processing residues neutralised


8


with CO2, after calcination at 500 °C, would show maximum sorptivity towards Zn(II), amounting to 14.92 mg g-1 (Sahu et al. 2011). In other column experiments in which uncalcinated red mud was subjected only to heating at 500 °C, better sorptivity towards Zn(II), Cd(II), Pb(II) and anions of chromate(VI), amounting to 25.5, 28.7, 64.8 and 35.7 mg g-1, respectively, was obtained (Gupta and Sharma, 2002, Gupta et al., 2001). After the heat treatment, the tested sorbent exhibited a low pHpzc of 3.2 and the dominant crystalline phases in its structure were hematite, anatase and rutile (also cancrinite was identified). After sorbent calcination, metal cations are probably removed mainly via specific and nonspecific adsorption on the surface of oxides, and to a lesser degree with the participation of their precipitation in the form of hydroxides. When red mud of a different origin was subjected to a similar treatment (neutralisation with CO2 and heating at 500 °C), a sorbent with a high iron oxide content (exceeding 50%) exhibiting high sorptivity towards arsenates(V) (55.6 mg g-1) was obtained (Sahu et al., 2010). As part of another research, the thermal activation of red mud was preceded by washing with HCl and an improvement in its sorptive properties towards As(V) was obtained. The acid treatment in the first stage removed sodalite, which had blocked the active centres on the surface of iron oxides, while the heat treatment at a temperature of 500 °C contributed to the development of the sorbent’s porous structure and to an increase in its iron and aluminium oxide content. Considering that the sorption of arsenates and chromates proceeds mainly via ligand exchange on the surface of iron and aluminium oxides, it seems that this activation method is the most effective when sorbents for the removal of the above anions from water are to be obtained (Genc-Fuhrman et al., 2004). The factor limiting the possible wide use of red mud as a sorbent is its physical form. Red mud consists of micrometeric grains varying rather widely (from 2 to 100 µm) in size, which is disadvantageous from the point of view of its potential applications (high hydraulic bed resistances and difficulties in separating the spent sorbent from the solution being purified). Therefore, studies aimed at modifying the physical form of red mud are underway. Their results are presented further on in this article.

accompanied by the coprecipitation of the pollutants dissolved in the water, such as heavy metals, phosphates, etc. (Pizzi, 2010). The composition of the forming deposit depends mainly on the type of coagulant used, the process conditions and the composition of the treated water. As a result, wastes in which iron oxides (Fe-WTRs) or aluminium oxides (Al-WTRs) predominate are obtained. Water deironing usually consists in water aeration aimed at oxidising Fe2+ (occurring as Fe(HCO3)2, FeSO4 and FeCl2) to Fe3+, which precipitates to the form of insoluble hydroxide:

( )

4 Fe 2 + + O 2 + 8 OH − + 2 H 2O → 4 Fe ( OH )3 ↓

During the removal of undesirable admixtures present in natural waters, considerable quantities of sediments, consisting mainly of aluminium or iron oxides, are generated. These are mainly wastes from the coagulation of usually surface waters and the deironing and demanganisation of underground and infiltration waters.

Similarly, as in the coagulation process, the precipitating iron hydroxide simultaneously brings about the removal of other pollutants (heavy metal cations, phosphates and arsenates) through coprecipitation and adsorption. Moreover, owing to the presence of manganese in the water being treated, Mn(II) oxidises to Mn(IV), which precipitates in the form of MnO2. Although this process proceeds with much greater difficulty than the oxidation of Fe(II), it significantly contributes to a change in the properties of the forming deposit, which owing to the presence of manganese dioxide acquires oxidising properties. Exemplary compositions of deposits coming from selected water treatment plants are shown in Table 3. Iron and aluminium oxides, being the principal components of WTRs, exhibit mainly an amorphous structure. This is owing to, among other things, the presence of phosphates, silicates and organic matter in the water being treated. Even small amounts of the substances practically block the crystallisation of the precipitating deposits, as evidenced by XRD analyses. On the other hand, studies into the selective extraction of the residuals with oxalates (in which only amorphous substances dissolve) show that some of the oxides (mainly iron oxides) present in them have a crystalline structure. Generally, iron oxides in WTRs are present in the form of ferrihydrite with a very low degree of crystallinity, referred to as 2-line ferrihydrite, hydrous ferric oxides or amorphous iron oxides. This is the principal structural form arising during the hydrolysis of Fe3+ conducted in standard conditions and a neutral environment. The structure of iron oxides precipitating in water treatment processes is unstable, particularly in the presence of moisture, and slowly undergoes recrystallisation leading to much better structurally ordered forms, e.g. α-geothite and hematite. Aluminium oxide occurs in the WTRs in almost exclusively the amorphous form, with only a small share (10 mg As(V)/g; >14 mg As(III)/g >15 mg As(V)/g; >8 mg As(III)/g 34.4 - 40.24 mg As(III)/g 44.95 - 49.98 mg As(V)/g 6.52 - 11.21 mg As(III)/g 4.92 - 9.18 mg As(V)/g 79.3 mg Hg(II)/g

62.16 mg Pb(II)/g; 86.83 mg Cr(III)/g 58.75 mg Cr(VI)/g 53.87 mg Pb(II)/g; 274.02 mg Cr(III)/g 58.23 mg Cr(VI)/g 20.98 mg As(V)/g; 18.73 mg As(III)/g 11.05 mg Se(VI)/g; 22.11 mg Se(IV)/g 17.55 mg Pb(II)/g; 15.86 mg Pb(II)/g

Sorption capacity, mg g-1

Lee et al., 2006 Irawan et al., 2011

Putra and Tanaka, 2011 Ippolito et al., 2009 Makris et al., 2006 Makris et al., 2006 Caporale et al., 2013 Caporale et al., 2013 Hovsepyan and Bonzongo, 2009 Gibbons and Gagnon, 2010 Gibbons and Gagnon, 2011 Gibbons and Gagnon, 2010 Gibbons and Gagnon, 2011 Chiang et al., 2012 Wu et al., 2004a

Zhou and Haynes, 2012



References

Jacukowicz-Sobala et al. 9

10


processes (Petruzzelli et al., 2000). But in recent years special attention has been given to the possibilities of utilising WTRs as adsorbents for removing toxic admixtures, mainly arsenates and to a lesser degree heavy metals, from sewage and natural waters. Arsenic is effectively removed from solutions with the use of both Fe-WTRs and Al-WTRs, and the obtained sorptivity ranges from a few to tens of mg g-1 sorbent, depending on the composition and specific surface area of the waste, as well as on the process conditions. Arsenates(V) are best adsorbed in a slightly acidic environment, and as the pH increases, their affinity to oxide surfaces clearly decreases. Arsenates(III) are effectively adsorbed in an environment close to neutral, because of an increase in the concentration of the dissociated form (H2AsO3-) in these conditions, whereas in a slightly alkaline environment the adsorption efficiency decreases owing to the unfavourable (negative) surface charge of oxide sorbents. However, the reductive dissolution of the adsorbent surface contributes to the widening of the effective range of As(III) adsorption when iron oxides are used as the adsorbent. For this reason, in a neutral environment arsenates(III) are sorbed more effectively than arsenates(V) by residuals with a predominant iron oxide content, whereas aluminium oxides exhibit higher sorptivity towards arsenates(V) (Caporale et al., 2013; Makris et al., 2006, 2009). At the same time, the rate of As(III) adsorption on aluminium oxides has been found to be considerably lower (Makris et al. 2006), which limits the possibility of conducting the adsorption process in column conditions. Thanks to the use of dried coagulation residuals (FeWTR) as the column filler, the treatment of natural groundwater containing 43 µg As/dm3 reduced the arsenate concentration to a level below 10 µg dm-3 (the allowable As concentration in drinking water) in water amounting to over 26,000 bed volumes. In the same conditions, the waste containing aluminium oxides (AlWTR) was found to be less useful and the sorption capacity determined in batch experiments amounted to merely 3 µg As/g at pH = 8.1 (Gibbons and Gagnon, 2010). But Al-WTR is characterised by high affinity to selenates(IV) and selenates(VI): the maximum sorption capacities determined in an environment with pH = 5 amounted to 11.05 mg Se(VI)/g and 22.11 mg Se(IV)/g (Zhou and Haynes, 2012). Similarly, as in the case of arsenates(V), selenates (especially Se(VI)) adsorption effectiveness is the highest in an acidic environment (pKa1[Se(IV)]=2.62, pKa1[Se(VI)]=1.7) and it quickly decreases with increasing pH. This is owing to the different mechanism of Se(IV) and Se(VI) adsorption. Selenates(VI) form weaker bonds with the surface of oxide adsorbents. The bonds are based mainly on electrostatic interactions (nonspecific adsorption) significantly dependent on the pH and the ionic strength of the solution. Selenates(IV) bind with the surface of oxides by forming considerably more stable coordination bonds (specific adsorption) (Hayes et al., 1987; Ippolito et al., 2009). Less attention is devoted to the utilisation of coagulation residuals in the removal of heavy metal cations, even though they are characterised by high sorptivity towards copper, mercury, lead or chromium(III) (Table 3). The reported research indicates that specific adsorption is the principal mechanism of binding heavy metals on the surface of iron and aluminium oxides, which ensures that the toxic admixtures are permanently bound with the adsorbent. But the mechanism of binding mercury on the surface

of WTRs is different (probably based on electrostatic interactions). As a result, the effectiveness of removing mercury ions considerably decreases when other cations are present in the solution being purified (Wu et al., 2004a). So far only a few studies on the utilisation of the residuals that form in the process of water deironing and demanganisation have been published, even though the residuals have both adsorption and oxidising properties. This combination of properties can be particularly useful in the removal of arsenates(III), which, because of their different behaviour in the water medium (pKa1=9.2 while for arsenates(V) pKa1=2.3), are usually adsorbed less effectively than arsenates(V). For this reason, to remove arsenates(III), preliminary oxidation of As(III) to As(V) is often applied, or adsorbents showing oxidising properties are used al., 2013; Kociołek-Balawejder and (Jacukowicz-Sobala et  Ociński, 2006). According to the authors’ research on the adsorption of arsenates(III) and (V), in which deironing process residuals were used, the latter showed a higher sorption capacity for As(III) (145 mg As/g) than for As(V) (111 mg As/g) (unpublished data). At the same time, the pH was found to have only a slight effect on the effectiveness of As(III) adsorption, which is probably owing to the substantial manganese(IV) oxide content in the waste. The oxidation of As(III) to As(V) on the surface of manganese(IV) oxide is combined with its reductive dissolution, followed by the adsorption of the reduced Mn2+ ions on the adsorbent surface whereby its total positive charge is increased, which favours the adsorption of the anions (Lafferty et al., 2011; Ociński et al., 2014; Scott and Morgan, 1995). Water deironing residuals also show a high affinity to cations of heavy metals, such as lead, cadmium, zinc and cobalt (Table 3). Deironing residuals were also used as an additive to soils polluted with arsenic (225–1033 mg kg-1) and chromium (27–371 mg kg-1) in order to immobilise the toxic admixtures. The introduction of dried deironing sludge (in the amount of 5 %wt.) into soil resulted in a reduction of the amount of leached arsenic and chromium by 98% and 91%, respectively (Nielsen et al., 2011). An attempt also was undertaken to use a sand bed as an adsorbent on which the sediments precipitated in the water deironing and demanganisation process settle. As the processes proceed, a layer of manganese dioxide, and to a lesser extent of iron oxides, covers the sand beads giving them oxidising and adsorptive properties. Thanks to its physical form this adsorbent shows very good hydraulic and mechanical properties, whereby it can be easily separated from the solution being purified and used as a bed in column processes. However, because of the small specific surface area of sand coated with manganese dioxide and iron oxides, such a sorbent is characterised by a very small sorption capacity for arsenates(III) and (V), amounting to 0.55 and 0.77 mg As/g, respectively (Jovanovic et al., 2011). WTRs are usually in the form of sludge with a high moisture content (about 50%–60%), with a solid phase particle size ranging from a few to several hundred micrometers. Depending on their intended further use, the wastes are subjected to thickening to the desired aqueous phase content or to drying. Since they are


11

Jacukowicz-Sobala et al. characterised by a neutral reaction (pH 6–7), they do not need (as opposed to bauxite processing wastes) any additional neutralisation and can be used as adsorbents without any further treatment. Owing to a large number of easily accessible adsorption sites on the surface of the particles and the easy diffusion of the adsorptive into them, the micrometric size of WTRs makes for high adsorption effectiveness in the removal of cations and anions from aqueous solutions. In studies on the adsorption of arsenates with coagulation residuals (both Fe-WTR and Al-WTR) carried out for four fractions with a graining of 1000–590, 590–250, 250–125 and below 125 µm, sorption capacities of 5–9 mg As(V)/g and 6–11 mg As(III)/g for Fe-WTR and 45–50 mg As(V)/g and 34–40 mg As(III)/g for Al-WTR were obtained (at pH = 6.0). The fractions with the smallest particle size were characterised by the highest specific surface area. The higher sorption capacities of the aluminium oxide containing residuals were owing to their considerably larger specific surface area in comparison with the residuals containing iron oxides (Caporale et al., 2013).

Fly ashes – their properties and possibilities of use as adsorbents Fly ashes are generated in high-temperature processes of burning solid fuels (hard coal, lignite and biomass), therefore their properties significantly differ from the properties of the sludges (red mud, WTRs) discussed earlier. The main components of fly ashes are silicon, aluminium, iron and calcium oxides, but their exact chemical and mineralogical composition depends on both the kind of solid fuel and the combustion conditions, as well as on the way of separating dust from the flue gas. Besides the main components, they contain oxides of alkali metals, small amounts of adsorbed heavy metals and a certain amount of unburnt coal. Fly ashes are mostly amorphous substances and the small amounts of crystalline fractions present in them include mainly quartz and to a lesser degree calcite, anhydrite, mullite, hematite, kaolin and geothite (Ahmaruzzaman, 2010; Apak et al., 1998; Bertocchi et al., 2006; Hsu et al., 2008; Khan et al., 2009). The presence of heavy metals and cyclic aromatic hydrocarbons adsorbed on the surface of fly ashes, which may gradually undergo leaching to ground and surface waters and soils, and the fact that fly ashes can be easily carried for longer distances by wind, contribute to their negative impact on the environment. It is estimated that in the world as much as 500 billion kilograms of fly ash are generated per annum, of which only 10%–20% is industrially utilised. The rest is subject to deposition (Ahmaruzzaman, 2010; Hsu et al., 2008). The necessity of properly protecting flyash disposal sites increases the total cost of generating electricity from coal. Therefore practical uses for these wastes are sought. One of the principal parameters that determine the properties of fly ashes and directions for their utilisation is the calcium oxide content. Fly ashes in which calcium oxide amounts to no more than 10%–20%, and which contain over 70% SiO2, Al2O3 and Fe2O3, show pozzolanic properties and are classified as grade F. Fly ashes with a high calcium oxide content (30%–40%), in

which the total content of silicon, iron and aluminium oxides does not exceed 70%, show properties similar to cement (harden in contact with water) and are classified as grade C (Ahmaruzzaman, 2010). Owing to these properties, the primary uses of fly ashes today are the production of clinkers, Portland cement, noncementitious binders, concrete and lightweight aggregates and the building of roads, embankments and dumping grounds (Fu et al., 2002; Haiying et al., 2007; Kruger, 2005; Kula et al., 2002; Lingling et al., 2005). Fly ashes with a high amorphous silicon dioxide content and a low calcium content have particularly advantageous properties – they can replace as much as 40% of the cement in concrete and at the same time improve its utilitarian properties (Kruger, 2005). Fly ashes show several properties that make it possible to extend the range of their practical applications. Thanks to the regularity of the shape (usually spherical) of the particles, their porosity, chemical composition (a high silica and aluminium oxide content) and to a large extent amorphous structure, this waste can be used as a raw material for the production of synthetic zeolites (Apiratikul and Pavasant, 2008; Lee et al., 2000; Medina et al., 2010; Nascimento et al., 2009). The use of fly ashes (raw or subjected to additional treatment) as adsorbents of toxic admixtures (both cationic and anionic) from industrial wastes has attracted much attention in recent years (Table 4). According to the above data, the sorption capacity of fly ashes, as compared with heavy metals and arsenates, ranges very widely depending on the origin of the fly ash, its pre-treatment and the adsorption process conditions. In comparative studies in which Cu(II), Pb(II) and Cd(II) ions were removed from aqueous solutions in similar conditions, using both fly ashes and red mud, as much as four-fold higher adsorption effectiveness was achieved in the case of fly ashes, even though they were not subjected to any additional treatment, except for washing and drying (Apak et al., 1998). Other studies showed red mud to be a much more effective adsorbent of heavy metals (copper, cadmium, lead, zinc) and arsenates present in mine flotation tailings. The latter mixed with red mud or fly ash would be placed in columns and subjected to leaching with distilled water for 80 days. Cadmium and copper leaching from the tailings was negligible for both column packings. The concentration of arsenic, lead and zinc in the effluent from the column containing the tailings mixed with red mud was 216, 17 and 40 times lower, respectively, than the permissible values for industrial wastes discharged into natural waters. Thanks to the use of fly ash as an immobilising substance, the concentration of arsenic, lead and zinc in the effluent was reduced to a level 17.5, 4.5 and 2.5 times lower, respectively, than the permissible one. In the authors’ opinion, the higher effectiveness of red mud is owing to its larger specific surface area, its aluminium and iron oxide content and its higher alkalinity ensuring a higher environment pH favouring heavy metal adsorption on oxide sorbents (Bertocchi et al., 2006). The considerable differences in sorption properties between the analysed groups of wastes are, to a large extent, owing to their different chemical structure (the mutual ratios between silica and iron, aluminium and calcium oxides) and physical structure


Coal combustion Biomass combustion Volcanic ash

Source of residuals

8.4

6.4

9.4

6.3

3.7

8.7

4.9

4.9

7.7 7.7

60.1

48.2

52.5

57.7

38.1

60.5

60.5

51.3 51.3

Fe2O3

57.2

SiO2


38.0 38.0

15.4

15.4

23.1

29.5

27.9

18.2

18.6

22.0

Al2O3

0.78 0.78

2.9

2.9

13.9

2.49

6.27

12.9

6.30

2.97

CaO

Chemical composition, %

Table 4. Fly ashes for heavy metal removal from water.

149.7 147.1

450

60 °C, 24 h (in H2O2) 105 °C, 24 h Rinsing with FeCl3, calcination (300 °C), 1 h

450

7.5 3.1 13.9 − − −

10.2

40.2

−

SBET, m2 g-1

− 600 °C, 4 h 90 °C, 0.5 h (in 6 M NaOH) Rinsing with 0.6 M MgCl2 Rinsing with 0.3 M MnCl2 and KMnO4 Rinsing with 1 M NaOH, 24 h; 60 °C, 48 h 60 °C, 24 h (in H2O2)

Pelletisation

110 °C, 24 h

105 °C, 2 h

Residuals treatment

15.7 mg Cu(II)/g (pH 4) 17.2 mg Pb(II)/g (pH 4) 0.17–0.97 mg Cd(II)/g 0.343–1.234 mg Pb(II)/g 6.23–14.7 mg Cr(VI)/g 20.92 mg Cu(II)/g 18.98 mg Cd(II)/g 178.5–249.1 mg Cu(II)/g 126.4–214.1 mg Cu(II)/g 76.7–137.1 mg Cu(II)/g 192.0–197.5 µg As(V)/g 178.5–195.6 µg As(V)/g 14.33 mg Cd(II)/g 55.55 mg Ni(II)/g 1.24–2.0 mg Cd(II)/g 1.12–1.7 mg Ni(II)/g 2.26–2.36 mg Cu(II)/g 2.34–2.54 mg Zn(II)/g 5.3–6.51 mg As(V)/g 6.13 mg As(V)/g

Sorption capacity, mg g-1

Alinnor, 2007 Khan et al., 2009 Papandreou et al., 2007 Hsu et al., 2008 Hsu et al., 2008 Hsu et al., 2008 Li et al., 20012 Li et al., 20012 Chaiyasith et al., 2006 Gupta et al., 2003 Gupta and Ali, 2000 Chen et al., 2010a Chen et al., 2011

References

12 Waste Management & Research

13

Jacukowicz-Sobala et al. (specific surface area, the degree of crystallinity and the kind of crystalline phases). This makes it difficult to explicitly compare the usefulness of the particular groups of wastes as adsorbents or heavy metals immobilising substances and usually requires detailed studies.

Possible ways of managing spent adsorbents A major problem connected with the use of iron and/or aluminium oxide materials as adsorbents is their management after the adsorption process. Because of the high content of toxic admixtures, their management usually comes down to their safe deposition in proper conditions. This is mainly so because the removed toxic admixtures are permanently immobilised on the surface of iron and aluminium oxides, whereby desorption proceeds with difficulty if no proper regenerating solutions (hydroxides or acids with proper concentrations) are used. At the same time, owing to the above fact, spent adsorbents can be treated as non-toxic substances (not requiring any special deposition conditions), which can be used in such things as land reclamation. The basic criterion for classifying spent sorbents as non-hazardous wastes is their limited environmental impact – mainly the fact that only small amounts of toxic elements and chemical compounds are leached from them. Several tests for determining the stability of toxic admixture binding by an adsorbent have been developed. Such tests are conducted in strictly defined conditions and the test procedures depend on the regulations binding in a particular country. Usually tests based on the methods developed by the Environmental Protection Agency (US EPA), such as singlepoint pH tests (Toxicity Characteristic Leaching Procedure (TCLP)) (USEPA, 1992) and the more comprehensive Leaching Environmental Assessment Framework (LEAF) (Garrabrants et al., 2010) involving four different test methods, or European leaching tests, such as pH-dependence CEN/TS 14429, liquid/ solid ratio dependence EN 12457, ‘up-flow’ percolation CEN/ TS 14405 (Insa Lyon, 2005), are used. However, because of its simplicity, TCLP is most often used in screening tests. The procedure consists of washing the spent sorbent with deionised water with pH = 3.0 or 5.0, obtained through correction with acetic acid or nitric acid. If the leached element concentration in the supernatant liquid does not exceed the maximum contaminant level binding in a given country, the waste is classified as nonhazardous. TCLP tests are often used to assess the environmental impact of the considered groups of industrial wastes (red mud, WTRs and fly ashes), previously utilised as adsorbents of heavy metals and metalloids. Studies have shown that such metals as copper, cadmium, cobalt and arsenic compounds are highly stable and immobilised on the surface of the above sorbents, which is probably owing to the mechanism of their adsorption (Brunori et al., 2005; Chaiyasith et al., 2006; Escudero et al., 2009; Genc et al., 2003; Genc-Fuhrman et al., 2004; Wu et al., 2004b). Moreover, arsenic compounds can be incorporated into the structure of slightly soluble minerals, such as calcium–iron arsenates and calcium–aluminium arsenates, and they can also undergo

isomorphous replacement in the structure of slightly soluble hydroxy-oxide minerals, e.g. calcium–iron arsenates or calcium– aluminium arsenates (Genc et al., 2003). It has been shown that also Cu2+ adsorption on the surface of red mud is accompanied by the production of atacamite, whereby the immobilisation of copper is much more stable than that of Zn2+ and Cd2+ cations (Ma et al., 2009). The adsorption of metal cations (e.g. Zn2+ or Mn2+) proceeding with precipitation or co-precipitation results in lower resistance to leaching. At the same time, despite the much higher leachability (10%–30% of the total adsorbed amount) of these cations in comparison with that of As(III), As(V) and Cu(II) compounds (0.2%–1.4%), their content in the supernatant liquid did not exceed the allowable value (Brunori et al., 2005). Also the influence of pH on the leaching of adsorbed radium, uranium and thorium radionuclides from red mud has been studied (Clark et al., 2011a, 2011b). The highest concentration of the radionuclides in the solution after desorption was recorded at pH = 3.0, but after 8 h of the equilibration of the sorbent in the solution, the thorium and radium concentrations in the solution decreased again. The effect of spent sorbent stabilisation on its resistance to the leaching of adsorbed metals was observed also in the case of red mud after the adsorption of copper, cadmium, zinc and lead. Probably at a low pH, iron oxides undergo dissolution together with the adsorbed species, which after stabilisation again undergoes adsorption and translocation from the surface to the lattice defect sites (Clark et al., 2009, 2011a, 2011b; Li et al., 2010). The industrial waste-based adsorbents described in this article can be managed (besides through storage and land reclamation) after they have lost their sorption capacity also, in a similar way as the primary wastes. This applies particularly to red mud and fly ash whose physicochemical characteristics make them highly suitable for use in the construction industry. Both bauxite processing wastes and fly ashes are used as raw materials in the production of cement, special concretes, lightweight cellular concretes, bricks and other ceramic blocks (Ahmaruzzaman, 2010; Klauber et al., 2011; Rai et al., 2012; Sutar et al., 2014; Wang et al., 2008). It has been shown that the leachability of toxic elements from this kind of material is very low. Especially fly ashes are characterised by very good properties as regards the permanent immobilisation of adsorbed toxic admixtures (Ahmaruzzaman, 2010; Papandreou et al., 2007). The possibility of desorbing cations and anions in deposition conditions is similar as in the case of the other considered wastes, however, owing to the special properties of grade C and F fly ashes they can be permanently solidified by adding a small amount of cement and/ or fresh fly ash. By solidifying the spent sorbent with concrete (in which 50%–75% of the cement had been replaced with fresh fly ash), the blocks in which practically no leaching of cadmium and copper ions in the alkaline, neutral and acidic environment occurred were obtained (Papandreou et al., 2007). Fly ashes are also used as a raw material for the production of synthetic zeolites and as an auxiliary material for building roads (Ahmaruzzaman, 2010). Even though WTRs could also be used in the production of ceramic materials (Wu et al., 2004a), their processing is limited to dewatering (thickening, conditioning,


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mechanical pressing, drying) aimed at reducing the surface area of the waste disposal sites (Dharmappa et al., 1997). Also research on the possibility of regenerating spent absorbents, whereby they could be reused many times, is being conducted in order to extend effective sorbent operating time and minimise the amount of deposited wastes. For this purpose, strong base (NaOH) or strong acid (HNO3) solutions, depending on the type of the admixture being removed, are usually used. In the case of oxyanions (As(V), Se(VI) or Cr(VI)), thanks to the competitive action of hydroxyl anions on the sorbent surface, the adsorbed anions can be desorbed with an almost 90% effectiveness (Sahu et al., 2010; Zhang et al., 2008; Zhou and Haynes, 2012). In the case of the desorption of heavy metal cations, the best results (about 90% effectiveness) are achieved using nitric acid solutions (Gupta and Ali, 2000; Gupta and Sharma, 2002; Gupta et al., 2001; Zhou and Haynes, 2011). Research on the use of an ethylenediaminetetraacetic acid (EDTA) solution for this purpose has also been conducted (Chiang et al., 2012). Because of the usually incomplete desorption of the bound ions (both cations and anions) only a few sorption/regeneration cycles can be run without a significant loss in the sorption capacity of the waste (Genc-Fuhrman et al., 2005; Gupta and Sharma, 2002; Gupta et al., 2001; Sahu et al., 2010; Zhou and Haynes, 2011, 2012; Zhu et al., 2007).

Limitations in use of iron and aluminium oxides containing industrial wastes in water treatment Leachability of toxic substances from sorbents into treated water Since all the three discussed groups of industrial wastes contain toxic substances (mainly heavy metals), much attention is devoted to their leachability, which may contribute to the secondary pollution of the treated solutions. This problem applies to mainly bauxite processing wastes and fly ashes owing to their particularly high heavy metal content. Both the standardised leachabilty tests (TCLP, LEAF) mentioned earlier and ecotoxicologic tests (e.g. the MicrotoxTM test and the ASTM microalgae toxicity test), making it possible to determine the effect of the tested substances on living organisms (Brunori et al., 2005; Orescanin et al., 2003; Singh and Singh, 2002; Wang et al., 2008), are used to assess the suitability of the above groups of wastes as adsorbents. The results of such tests indicate above all the very stable bonding of heavy metals, arsenates and selenates with the aluminium and iron oxides present in WTRs, red mud and fly ashes, making the desorption of the former significantly difficult. The amount of heavy metals and metalloids leached from the wastes depends on the process conditions, but usually (also in a slightly acidic environment) their concentration in the effluent is very low and does not exceed the allowable standards, whereby no secondary contamination hazard is posed to the solutions being treated (Ahmaruzzaman, 2010; Sutar et al., 2014; Wang et al., 2008). In neutral and slightly alkaline environments, the amount of leached heavy metals is usually negligibly small.

However, because of the considerable differences in the physicochemical properties of the wastes (even within the same groups), each time it is necessary to determine their stability in the environment in which they are to be used as an adsorbent (Klauber et al., 2011; Wang et al., 2008).

Physical form of sorbents obtained from wastes containing iron and aluminium oxides For the processes of heavy metals immobilisation in soils or in deposited industrial wastes, the physical form of the bauxite ore processing or water treatment sludges is highly advantageous because of the large interphase surface area. However, in the processes of adsorptive cleaning of solutions, this physical form makes it difficult to separate the sorbent from the solution being cleaned, and also, because of the high hydraulic resistance leading to bed clogging, limits the possibility of conducting adsorption in column conditions. However, so far only a few publications on this subject (mostly dealing with the research on the modification of bauxite processing wastes) have appeared. One of the presented solutions consisted in depositing the wastes on grains of sand, whereby a material similar to the one used in water treatment sand filters was obtained. A slurry of red mud (neutralised with sea water or activated with sulphuric acid and then thermally) was deposited and dried and used as column bed sorbents in the process of purifying water from As(V) compounds. The best results in the removal of arsenic from water were obtained in a process in which the waste was neutralised with only sea water. The column breakthrough capacity (C/C0 = 0.9) of 2388 BV and 3192 BV at a flowrate of 0.175 and 0.275 dm3 h-1, respectively, was achieved (Genc-Fuhrman et al., 2005). Another solution consists of the use of red mud with a phosphogypsum addition (8% of the CaSO4 content). In this way, stable aggregates, which can be easily filtered out or used as a column bed, are obtained. This material was used to treat wastes containing cations of heavy metals: nickel, zinc, copper and organic compounds, nitrates, phosphates and ammonium cations. Phosphates and Ni(II) were completely removed from the wastes while copper and zinc cations were removed to a certain degree (the probable cause of the lower effectiveness of adsorption of these ions was the formation of their organic complexes showing weaker affinity to the sorbent) (Lopez et al., 1998). A similar sorbent was used to remove arsenates(V) from water. The sorbent was characterised by better physical properties and it was also found that as the phosphogypsum content increased (from 0% to 25%), the sorbent’s sorption capacity for As(V) increased from 0.91 mg g-1 to 3.33 mg g-1 (Lopes et al., 2013). Red mud was also used to produce granulates showing good sorptivity towards phosphates, fluorides and cadmium (Tor et al., 2009; Zhao et al., 2012; Zhu et al., 2007). Since red mud does not show good cohesion properties in order to obtain a granulate, it is admixed with sodium silicate or clay minerals (e.g. bentonite), which perform the function of a binder. Thanks to the use of this material in a column bed, effluents containing Cd(II) could be treated at a sorption capacity


15

Jacukowicz-Sobala et al. of 5.92–6.69 mg g-1 while the regeneration of the bed with dilute hydrochloric acid made it operate effectively for four sorption/ desorption cycles (Zhu et al., 2007). Only a small number of published studies deal with the modification of the physical form of WTRs. The methods of modifying WCRs reported to date are based mainly on their sintering at a temperature of 1000 °C and under high pressure, whereby, besides an improvement in the physical properties, also a reduction in the amount of heavy metals washed out of the deposit is obtained. At the same time this process leads to a significant reduction in the waste’s specific surface area and so to a deterioration in its adsorption properties (Table 3) (Wu et al., 2004a, 2004b). An interesting solution consists in conducting adsorption in a fluidised-bed reactor, whereby high adsorption effectiveness can be achieved without it being necessary to additionally separate the sorbent as is the case in processes run in the batch regime. Besides, no high bed hydraulic resistances typical of column systems occur in the fluidised-bed reactor. Thanks to the use of the fluidised-bed system in the removal of copper from a solution with an initial concentration of 20 mg dm-3 at pH = 4, a purification efficiency of 90 % was achieved in 60 min (Lee et al., 2006). Because of the peculiar properties of fly ashes, research on the modification of their physical form focuses on forming them into durable blocks suitable for use as column packing. Fly ashes from hard coal burning, showing cementitious properties (grade C), were subjected to forming (pelletisation) without the use of any additional binder and then kept in an environment whose humidity exceeded 90%, whereby a stable sorbent in the form of beads 3–8 mm in diameter, characterised by high mechanical strength, was obtained. Despite its small specific surface area of 10.2 m2 g-1, its sorption capacity in the experimental conditions amounted to 21 mg Cu(II)/g and 19 mg Cd(II)/g (Papandreou et al., 2007). Another interesting idea consisted in pressing fly ash with an addition of starch and Fe2O3 at elevated temperatures to obtain ceramic blocks (Chen et al., 2010b, 2012). In this way, a sorbent in the form of highly porous spherical beads characterised by very good hydraulic properties was obtained. In order to improve its adsorption properties, it was coated with iron oxides on both its surface and inside its macro- and mesopores. Even though the specific surface area of the sorbent decreased (from 149 to 45 m2 g-1) as a result, its sorption capacity for arsenates(V) increased by about 30% relative to the raw fly ash (Chen et al., 2012). To date, however, this method has been used only in order to modify natural volcanic ashes.

Conclusion All the groups of iron or aluminium oxide-containing industrial wastes discussed in this article show high sorption effectiveness towards most of the heavy metals and semimetals present in the effluents. Moreover, owing to the presented methods of physicochemical treatment, their basic properties can be modified to improve the effectiveness of removal of selected groups of pollutants from aqueous solutions or to immobilise the pollutants

more permanently. However, a major drawback of most of the considered wastes is their physical form. Both sludge wastes and fly ashes consist of small solid particles from 10–20 to several hundred micrometers in size, ensuring on the one hand, a large specific surface area and low diffusional resistances during adsorption, but on the other hand, making the separation of the spent sorbent from the solution difficult. Moreover, the hydraulic resistances of a bed consisting of particles of such dimensions make it difficult or even impossible to conduct adsorption in the column system usually used in industrial installations because of its high effectiveness and technological simplicity. Although a few attempts to solve this problem have been undertaken, they applied mainly to fly ashes, which owing to their peculiar properties can be formed into beads that are durable in the water environment. For this reason, fly ash is used as an important additive in the pelletisation of red mud. Considering on the one hand, the generation of huge quantities of oxide wastes and the necessity to manage them, and on the other hand, their very good adsorption properties, one can expect that the interest in converting them into technologically useful adsorbents will grow in the near future. In the literature on the subject, one can find reports dealing with the preparation of sorbents based on very fine (nanometric particle size) synthetic iron oxides, characterised by good hydraulic and mechanical properties, and so suitable for column systems (Chen et al., 2010b; de Santa Maria et al., 2003; Guo et al., 2008; Mahajan et al., 2006; Paulino et al., 2011; Swedlund and Webster, 1999). The reports can provide some tips on the treatment of waste iron and/or aluminium oxides aimed at improving their physical form.

Declaration of conflicting interests The authors declare that there is no conflict of interest.

Funding This research received no specific grant from any funding agency in the public, commercial, or not-for-profit sectors.

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Unusual sources of aluminium and heavy metals in potable waters.

Microbial leaching of metals from solid industrial wastes.

Sources and fates of heavy metals in a mining-impacted stream: temporal variability and the role of iron oxides.

General facile approach to transition-metal oxides with highly uniform mesoporosity and their application as adsorbents for heavy-metal-ion sequestration.

Aluminium and heavy metals in potable waters of the north Ceredigion area, mid-Wales.

Sorption behavior of heavy metals on poorly crystalline manganese oxides: roles of water conditions and light.

Ion exchangers as adsorbents for removing metals from aquatic media.

Possibilities and limitations of reaction product analysis.

Properties of bacterial laccases and their application in bioremediation of industrial wastes.

Assessment of heavy metals exposure, noise and thermal safety in the ambiance of a vacuum metallurgy separation system for recycling heavy metals from crushed e-wastes.

Possibilities and limitations of intravital imaging.

[MRI in musculoskeletal imaging: possibilities and limitations].

Use of solvent extraction for the removal of heavy metals from liquid wastes.

The role of biochar, natural iron oxides, and nanomaterials as soil amendments for immobilizing metals in shooting range soil.

In silico protein modeling: possibilities and limitations.

Occurrences and inventories of heavy metals and brominated flame retardants in wastes from printed circuit board production.

Microbial treatment of sulfur-contaminated industrial wastes.

Chitin Adsorbents for Toxic Metals: A Review.

"Microscopic evidences of heavy metals distribution and anatomic alterations in breaching-leaves of Cupressus lindleyi growing around mining wastes".

Removal of heavy metals from tannery effluents of Ambur industrial area, Tamilnadu by Arthrospira (Spirulina) platensis.

Chitosan-based biosorbents: modification and application for biosorption of heavy metals and radionuclides.

Agro-industrial wastes as effective amendments for ecotoxicity reduction and soil health improvement in aided phytostabilization.

The possibilities and limitations of forensic hand comparison.

Endoscopic therapy of esophageal cancer: possibilities and limitations.