HEAVY METAL CATION RETENTION BY UNCONVENTIONAL SORBENTS (RED MUDS AND FLY ASHES)

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Toxic heavy metals, i.e. copper (II), lead (II) and cadmium (II), can be removed from water by metallurgical solid wastes, i.e. bauxite waste red muds and coal fly ashes acting as sorbents. These heavy-metal-loaded solid wastes may then be solidified by adding cement to a durable concrete mass assuring their safe disposal. Thus, toxic metals in water have been removed by sorption on to inexpensive solid waste materials as a preliminary operation of ultimate fixation. Metal uptake (sorption) and release (desorption) have been investigated by thermostatic batch experiments. The distribution ratios of metals between the solid sorbent and aqueous solution have been found as a...

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Nội dung Text: HEAVY METAL CATION RETENTION BY UNCONVENTIONAL SORBENTS (RED MUDS AND FLY ASHES)

IVat. Res. Vol. 32, No. 2, pp. 430-440, 1998 © 1998 ElsevierScienceLtd. All rights reserved Pergamon Printed in Great Britain P II: S0043-1354(97)00204-2 0043-1354/98 $19.00 + 0.00 H EAVY METAL CATION R ETENTION BY U NCONVENTIONAL SORBENTS ( RED MUDS AND FLY A SHES) R E , A T APAK*, ESMA TOTEM, M E H M E T HLIGI]L and JI]LIDE H I Z A L Department of Chemistry, Faculty of Engineering, Istanbul University, Avcdar, 34850, Istanbul, Turkey (First received April 1996; accepted in revised form June 1997) A l~raet--Toxic heavy metals, i.e. copper (II), lead (II) and cadmium (II), can be removed from water by metallurgical solid wastes, i.e. bauxite waste red muds and coal fly ashes acting as sorbents. These heavy-metal-loaded solid wastes may then be solidified by adding cement to a durable concrete mass assuring their safe disposal. Thus, toxic metals in water have been removed by sorption on to inexpen- sive solid waste materials as a preliminary operation of ultimate fixation. Metal uptake (sorption) and release (desorption) have been investigated by thermostatic batch experiments. The distribution ratios o f metals between the solid sorbent and aqueous solution have been found as a function of sorbent type, equilibrium aqueous concentration of metal and temperature. The breakthrough volumes of the heavy metal solutions have been measured by dynamic column experiments so as to determine the sat- uration capacities of the sorbents. The sorption data have been analysed and fitted to linearized adsorp- tion isotherms. These observations are believed to constitute a database for the treatment of one industrial plant's effluent with the solid waste of another, and also to utilize unconventional sorbents, i.e. metallurgical solid wastes, as cost-effective substitutes in place of the classical hydrous-oxide-type sorbents such as alumina, silica and ferric oxides. © 1998 Elsevier Science Ltd. All rights reserved K ey words---cadmium (II), lead (II), copper (II), sorption, red muds, fly ashes INTRODUCTION V arious treatment technologies have been devel- o ped for the removal of these metals from water. C admium (II), lead (II) and copper (II) are well- T he hydrometallurgical technology extracts and k nown toxic heavy metals which pose a serious c oncentrates metals from liquid waste using any of t hreat to the fauna and flora of receiving water a v ariety of processes, such as ion exchange, electro- b odies when discharged into industrial wastewater. d ialysis, reverse osmosis, membrane filtration, I n spite of strict regulations restricting their careless s ludge leaching, electrowinning, solvent stripping, d isposal, these metal cations may still emerge in a p recipitation and common adsorption (LaGrega e t v ariety of wastewaters stemming from catalyst, elec- a l., 1994a). t rical apparatus, painting and coating, extractive B oth powdered (Sorg e t al., 1978) and granular m etallurgy, antibacterials, insecticides and fungi- a ctivated carbon (Huang and Smith, 1981) have c ides, photography, pyrotechnics, smelting, metal b een used for the adsorptive removal of Pb, Cd and e leetroplating, fertilizer, mining, pigments, stabil- s imilar "soft" heavy metals, especially when associ- i zers, alloy industries, electrical wiring, plumbing, a ted with common organic particulate matter in h eating, roofing and building construction, piping, w ater. Activated carbon from cheaper and readily w ater purification, gasoline additive, cable covering, a vailable sources, such as coal, coke, peat, wood, a mmunition and battery industries (Buchauer, 1973; n utshell (Freeman, 1989) and rice husk (Srinivason, L ow and Lee, 1991; Periasamy and Namasivayam, 1986), may be successfully employed for the 1994) and sewage sludge (Bhattacharya and r emoval of heavy metals from aqueous solutions. V enkobachar, 1984). The acute toxicity of these H ydrous oxides such as alumina, iron oxides h eavy metals have caused various ecological cata- ( hematite and goethite) (Cowan e t al., 1991; Gerth s trophes in human history, such as the "itai-itai" a nd Bruemmer, 1983) manganese (IV) oxide d isease due to cadmium (Riley and Skirrow, 1975). ( Hasany and Chaudhary, 1986) and titanium (IV) P rolonged effect may cause other chronical dis- o xide (Koryukova e t al., 1984) have also been used o rders (Huang and Ostovic, 1978). f or the adsorption of the indicated heavy metals. T he cost of the adsorptive metal removal process *Author to whom all correspondence should be addressed. is relatively high when pure sorbents (either acti- 430
431 S orption of heavy metal cations t ion based on surface-complex formation, where v ated carbon or hydrated oxides) are used. m etal ions are usually removed as uncharged hy- T herefore, there is an increasing trend for substitut- d roxides condensed on to surfaces of - O H group i ng pure adsorbents with natural by-product or b earing adsorbents (Lieser, 1975), i.e. aluminium s tabilized solid waste materials for the development o xide, silica gel, ferric and titanium oxides, existing o f cost-effective composite sorbents capable of a s components of the utilized composite sorbents; t reating a variety of contaminants. For example, ( iv) ion exchange, where the acid-pretreated sor- r ecent evidence on the combined use of lime, ferric b ents may function as synthetic cation exchangers. a nd aluminium coagulants has shown that these O f these mechanisms, surface precipitation and s ubstances are more effective in combination than c hemical adsorption are believed to play the domi- i ndividually (Harper and Kingham, 1992). A num- n ant role in heavy metal ions removal (Apak and b er of metallurgical solid wastes such as bauxite U nseren, 1987). w aste red muds and coal-fired thermal plant fly T he aim of the present study is to develop cost- a shes have been screened in this regard to serve as e ffective unconventional sorbents, preferably metal- v ersatile and cost-effective sorbents for heavy metals l urgical waste solids, for heavy metal removal from ( Apak and (0nseren, 1987; Apak e t al., 1993) and c ontaminated water. The heavy metal (Pb, Cd and r adionuclides (Apak e t al., 1995; 1996). The ability C u) removal capacity as well as sorption modelling o f fly ash to remove metal cations from water has o f red muds and fly ashes will be evaluated in this a lso been demonstrated in the literature r egard. The irreversible nature of sorption needs to ( Bhattacharya and Venkobachar, 1984; Panday and b e shown so as to guarantee non-leachability of S ingh, 1985; Yadava e t al., 1987) for a limited num- m etals from the metal-loaded sorbents. b er of metals. T he alternative mechanism for heavy metal r emoval by red muds and fly ashes (either natural EXPEilIMENTAL o r in activated form) are assumed to comprise four s teps (Gregory, 1978; Apak a n d Llnseren, 1987). (i) M aterials and methods s urface precipitation (sweep flocculation), where A ll heavy metal solutions (divalent cations Of Pb, CA m ost hydrolysable heavy metals are removed via co- a nd Cu) were prepared in stock solutions up to 10000 p pm 0a g/ml) of metal from the corresponding nitrate p recipitation of their insoluble hydroxides forming salts. No further pH adjustment of these solutions was s uccessive layers on the sorbent surface; (ii) floccu- m ade as their natural acidity due to hydrolysis of metals l ation by adsorption of hydrolytic products, where (i.e. to form MOIl + and H +) prevented the precipitation m ulti-nuclear hydrolysis products (formed on the o f the corresponding metal hydroxides. All chemicals (E. Merck, Darmstadt, Germany) were of analytical reagent a dsorbent surface as kinetic intermediates) including grade. [ Fe2(OH)4 ]2+, [Fe3(OH)415+, [AI4(OH)s]4+ and O f the metallurgical solid wastes used as sorbents, the [AIs(OH)20]4 + act as more effective flocculants than red muds were supplied from Etibank Seydi~ehir t heir parent ions due to their higher charge and A luminum plant, Konya, Turkey and coal fly ashes s trong specific adsorptivities; (iii) chemical adsorp- were from TEK Af~in-Elbistan Thermal Power Plant, Table 1. Saturation capacitiesof the sorbents for the metals from column and batch experimentsand Langmuirparameters of equilibrium modelling LangmuirParametersb Qo (rag/8) Equilibrium Qexp.(rag/g) Qexp.(rag/g) Theoretical Metal iona Adsorbent pH Column capacity Batch capacity capacity b (litre/mg) Corr.coeff. (r) C d (II) F 7,2 220 198.2 374.3 1.14.10-3 0.957 C d (II) Fw 6.7 -- 195,2 223.2 I.17.10-3 0.970 C d (II) Fa 6.6 122 180,4 217.2 6.07.10-3 0,953 Cd OI) Rw 6.0 160 66.8 113.7 0.57.10-3 0.958 Cd (II) gab 5.9 115 66.8 112.0 0.65.10-3 0.994 Cd (II) Ra 4.2 105 46.9 107.5 0.11.10-3 0.989 Cu (If) F 6.0 -- 207.3 335.2 0.94,10-3 0.968 Cu (II) Fw 5.8 264 205.8 328.2 0.75,10-3 0.961 Cu (II) Fa 5.7 187 198.5 283.9 0.73.10-3 0.960 Cu (II) Rw 6.0 110 75.2 90.0 0.96.10-3 0.958 Cu (II) 5.7 100 65.2 87.8 0.79.10-~ 0.956 Rah Cu (II) Ra 4.5 63 35.2 65.4 1.00.10-3 0.964 P b (II) F 6.2 530 444.7 526.0 I.I1.10-3 0.948 P b (II) Fw 6.0 -- 483.4 490.7 1.10.10-3 0.976 Pb (II) Fa 6.0 -- 437.0 483.0 0.84.10-3 0.958 Pb (II) Rw 6.0 161 165.8 158.9 0.66.10-3 0.960 Pb (II) R.a 5.7 164 138.8 137.2 1.17.10-3 0.970 Pb (II) R, 4.4 123 117.3 118.5 1.56.10-3 0.956 "The initial aqueous metal concentrationsfor different metal/sorbentcombinationswere as follows: Cu (II) 50 mM (mmol/litre)for red muds and 90 mM for fly ashes; Pb (If) 50 mM for red muds and 65 mM for fly ashes; Cd (II) 35 ram for red muds and 40 mM for fly a shes, bCalculatedby the aid of iinearizedLangmuirequation (4),
432 Re,at Apak e t ,7l. .0 , , , , , , 5 .0 t A] [ CdF CdFw 4 .0 ~ caF, • I Jo.'~ ~]~ Cd~ 3 .0 I 1.ot • "1 0 .0 ~ • • I ~ ~ w 0 .0 1.0 2.0 3.0 4.0 5.0 6.0 C e ( mi/mL) Fig. 1. Distribution coefficient of Cd (II) as a function of equilibrium aqueous concentration on fly ashes and red muds. K ahramanmara~, Turkey. The red muds, obtained as alka- a dsorbent from red muds in phosphorus (Shiao and line leaching wastes of bauxite in the Bayer process of A kashi, 1977) and heavy metal (Apak and Unseren, 1987; A pak e t al., 1995, 1996) removal. However, acid treatment a lumina manufacture, had the following chemical compo- o f red mud sorbents had the drawback of the partial loss sition by weight: Fe203 37.3%, A1203 17.6%, SiO2 16.9%, TiO2 5.6%, Na20 8.3%, CaO 4.4%, loss on ignition o f acid-soluble fractions like hematite. The Ra fraction was further subjected to heat treatment at 600°C for 4 h 7.2%. Red muds, being multicomponent systems, are com- t o obtain the Rah sorbents. The red muds (partly agglom- p osed of sodium aluminosilicates, kaolinite, chamosite, e rated due to relative humidity) could not be classified i ron oxides (hematite) and hydroxides. Basically, Fe is in with respect to true grain size as most were of 200 mesh t he form of hematite, Ti is in the form of Fe-Ti oxides size in wet sieving. a nd AI is in the form of ahiminosilicates. 94% of red m uds have less than 10/an grain size. T he specific areas of Rw, Ra and Rah samples were 14.2, 20.7 and 28.0 m2/g, respectively, measured by the BET/Nz T he red muds were thoroughly washed with water to a m ethod (Brunauer e t al., 1938). neutral pH, dried and sieved (R,) prior to adsorption tests. The red muds were also acid-treated (R~). The acid C oal fly ash was recovered from the cyclones and elec- t reatment was carried out according to a modified version t rostatic precipitators of the power plant and had the fol- o f Shiao's procedure (Shiao and Akashi, 1977) by boiling lowing average composition: CaO 42.5%, SiO2 21.9%, I00 g of water-washed and dried red mud in 2 dm 3 of SO3 13.6%, A1203 11.8%, Fe.zO3 2.4%, MgO 1.3%, K20 10% (by weigh0 HCI solution for 2h, filtering off, 1.1%, Na20 0.9%, loss on ignition 4.4%. Almost 99% of t horoughly washing with water, drying and sieving to t he fly ash could pass through a 200-mesh sieve. The raw o btain the Ra-sorbents. The acid-treatment technique, fly ash (F) was washed with 10-fold distilled water for sev- which has also been applied by Wahlberg e t al. (1964) to e ral (5-6) times, filtered and dried (Fw). A part of the Fw c lay minerals for improving their surface properties, has was further treated several times with acid using 2% (by been demonstrated with success in synthesizing a better weight) HCI in boiling solution for 2 h. Higher acidity (as 8 .0 [] oaf 5 .0 A c~Fw • 0.% 4.0 2.o3°1.o • 0.0 0 .0 0.2 0.4 0.6 0.8 1.0 1.2 C e l ing/roLl Fig. 2. Distribution coefficient of Cu (II) as a function of equilibrium aqueous concentration on fly ashes and red muds.
S orption of heavy metal cations 433 6.0 5.0 A ~'~w PbF,, 4 .0 • ~'Rw 1.0 ~ 0 .0 0 .0 1.0 2.0 3.0 4.0 5.0 8.0 7.0 8.0 9.0 (m mLl Ce F ig. 3. Distribution coefficient of Pb (II) as a function of equilibrium aqueous concentration on fly ashes and red muds. in the activation of red mud) was avoided due to severe line phase. Elemental analysis of selected spots in the het- losses of fly ash components by solubilization. The solid e rogenous amorphous slag particles by the XRF technique p roduct was thoroughly washed with water, filtered, and ( Apak e t al., 1996) yielded 41-52% CaO, 27% SiO2, 13% o ven dried at 100 + 5°C to produce the acid-treated (Fa) A1203, 2-5% FeO, 1-4% MgO and up to 2% other ox- sorbent. ides. T he X-ray diffractogram (Apak e t al., 1996) of the Fw T he BET/N2 surface area of fly ash were 10.2 and identified 51% calcite (CaCO3), 32% anhydrite (CaSO4), 9 % quartz (SiO2) and 3% hematite (Fe203) in the crystal- 14.3 m2/g for Fw and Fa, respectively. 2 50 [] C dF A c a% • cd% 0 ,' o8 o 150 g @ [] ) 100 • • 50 0 0 0 ! 2.0 4.0 6.0 0.0 Co(msI L) Fig. 4. Isotherm of Cd (II) adsorption onto fly ashes and red muds.
434 Re,at Apak e t al. 160 [] C uF [] © C.F a [] ® c ,e,, c ,e, 120 [] © [] © 80 A © 40 ml, O r • 0.4 0.8 0.0 1 .2 C e(mg/mL) Fig. 5. Isotherm of Cu (II) adsorption onto fly ashes and red muds. P oint of zero charge (PZC) measurements by potentio- Ko = q o / G (2) m etric titration of the sorbent suspensions at different i onic strengths (Apak e t aL, 1995, 1996) yielded approxi- w here KD is the empirical distribution ratio of the metal m ate P Z C v alues of 6.4 and 8.3 for fly ash and red mud c ation M ((mg/g)/(mg/litre)= litres/g) determined on the s orbents, respectively. a pproximately linear portion of the corresponding adsorp- W hen 1 g of sorbent was equilibrated with 50 ml dis- t ion isotherm. tilled water, the indicated sorbents showed the following B atch desorption tests were carried out by agitating I g a pproximate final pH in their aqueous leachates: o f metal loaded sorbent with 50 ml of the desired solution u ntil equilibrium (8 h). Rw Ra R~ F Fw Fa T he saturation capacities of the sorbents for the uptake pH 8.1 4.8 5.3 12.0 10.8 9.3 o f indicated metals were determined by both batch and c olumn tests. For the latter, 40 g of adsorbent was filled at T he acid-treated sorbents contained no free HCI but 15- a h eight of 8-11 cm in a thermostatic (25 + 0.1°C) column 20 mg Cl-/g. o f dimensions (h = 30 cm, ~b = 3 cm), and the adsorbate B atch sorption tests were carried out by agitating a sus- s olution was fed (counter to gravity) by a peristaltic pump p ension of I g sorbent in 50 ml metal nitrate solution for t hrough the fixed bed of sorbent at a constant rate of 8 h ( equilibration period) at room temperature 0.5 ml/min. The metal concentration of the eluate was (25 +0.1°C) in stoppered flasks placed on a thermostatic r ecorded against throughput volume. The dynamic metal w ater-bath/shaker. After centrifugation, the remaining u ptake capacities of the sorbents were calculated by the in- m etal concentration in the filtrate was determined by t egration technique (Apak e t al., 1996), i.e. the area above f lame AAS (Perkin Elmer 300, Norwalk, CT, U.S.A.) and t he curve up to the line on which the eluate concentration t he equilibrium pH was measured by a pH-meter was equalized with the initial concentration of metal was ( Metrohm E-512 Herisau, Switzerland) equipped with a c alculated. The total amount of retained metal was divided glass electrode. b y the mass of sorbent to yield the saturation capacity T he metal concentration retained in the sorbent phase (t.)col. x (qe, mg/g) was calculated by (1) q e = (Co - G ) V / m R ESULTS AND DISCUSSION w here Co and C~ are the initial and final (equilibrium) con- c entrations of the metal ion in solution (mg/litre), V is the I n weakly acidic-neutral suspensions whose p H s olution volume (litres) and m is the mass of sorbent (g). w as attained naturally by equilibrating aqueous T he solid/water distribution ratios (at equilibrium) of m etal n i t r a t e - s o r b e n t mixtures, the distribution m etals for both sorption and desorption were calculated r atios generally increased with initial aqueous by
Sorption of heavy metal cations 435 500 [] [] [] [] I :bF A 0 400 0 @ ~ aa, [] [] O0 h a,, 0 & 0 300 0 0 2 00 • • •1 i 1~ .i , n II , |1 II I • 0.0 2.0 4.0 6.0 0.0 10.0 CelmllnnU Fig. 6. Isotherm of Pb (II) adsorption onto fly ashes and red muds. a dsorbate concentration at equilibrium up to a lim- m ay form at the surface of the hydrous oxide sor- i ting value where the batch capacities of the sor- b ent prior to its formation in bulk solution and b ents for the metals were calculated /~exp. ~. The 'nbatch~ t hus contribute to the total apparent sorption. The s aturation capacities found by both batch and c ontribution of surface precipitation to the overall d ynamic column experiments (the latter symbolized s orption increases as the sorbate/sorbent ratio is a s Q~L.) are listed in Table 1. The variation of dis- i ncreased (Stumm and Morgan, 1996). It should be t ribution ratio (KD) with equilibrium concentration a dded that raw fly ash (F) cannot be considered as o f the adsorbate in solution is shown in Figs 1, 2 a n EPA-acceptable sorbent (LaGrega, 1994b) as it a nd 3, where KD vs C~ on semi-logarithmic scale i ntroduces new contaminants to water in untreated g ave roughly linear plots. A gradual decrease of dis- ( either water- or acid-washed) form. t ribution ratios with aqueous concentration was T he adsorption isotherms (q~ vs Ce) of metal n oted, due to increased occupation of active surface u ptake at 25°C (see Figs 4, 5 and 6) essentially sites of sorbent with metal loading in the aqueous s howed BET (Branauer e t aL, 1938) (type II, V) s olutions (Apak e t al., 1995, 1996). As long as a c haracter curves pointing out to the heterogeneity s trict differentiation between true adsorption and o f the sorbents containing hydrous oxides, silicates p recipitation- masked sorption (McKay e t al., 1985) a nd sulfates, resulting in various combinations of is not made, both red muds and fly ashes may be l inear and nonlinear isotherms (Weber e t al., 1996). visualized as effective sorbents capable of removing I t is known from the literature that BET type IV-V t he studied heavy metal ions from solution with i sotherms are quite common for the porous hydrox- h igh distribution ratios, /~D ranging up to 10-2- as', ides (xerogels) such as siligagel or iron hydroxide 10-1 litres/g. Naturally the slightly alkaline charac- ( Gregory, 1978). t er of aqueous leachate obtained from fly ashes in A lthough Langmuir and Frenndlich approxi- c onjunction with the CaO and CaSO4 constituents m ations of the observed adsorption data in the line- o f this material should account for hydrolytic metal a rized forms gave satisfactory correlation p recipitation reactions (Burgess, 1978; Freeman, coefficients (r > 0.95) for most of the covered con- 1988) as well as counter-ion adsorption at pH c entration range, the Langmuir model had more a bove PZC accompanying chemical adsorption p ractical utility for representing the limiting sorp- ( Apak and Unseren, 1987; Apak e t al., 1993). t ion capacities of the sorbents than the exponen- G enerally a metal hydroxide may precipitate and t ially increasing Freundlich isoterm (McKay e t aL,
436 Re,at Apak et al. 0.08 e cde w II cue w ~ew 0.06 Q 0 0.04 0.02 0 .00 V 4 .0 6.0 8.0 0 .0 2.0 C e ( mg/mL) Fig. 7. Selected isotherms linearized with respect to the Langmuir model (red mud). 1985) in spite of the invalidity of the classical i n Table 1 (three runs made per isotherm) together L angmuir assumptions, i.e. site-specific and uni- w ith the experimental saturation capacities of batch f ormly energetic adsorption confined to monolayer rn~t~h~ and column (Q~exp.) tests. After screening of ol. c overage (Weber and DiGiano, 1995; Weber e t aL, t hose results where metal hydroxide precipitation 1996). Heavy metal adsorption on heterogenous c ould have been effective in metal removal (e.g. s orbents has been interpreted by the aid of the m odelling of Cu(H) sorption has been made up to L angmuir isotherm on various occasions in the en- t he concentration edge of Cu(OH)2 precipitation at v ironmental literature (Szymura, 1990; Prasad and t he studied pH), Langmuir modelling has been A garwal, 1991). q uite successful in predicting the experimental satur- A L angmuir equation for adsorption may be a tion capacities of the sorbents, especially those w ritten as o btained from dynamic column tests (see Table 1), a lthough its basic assumptions were not fulfilled Q ° b C~ ( Weber and DiGiano, 1995; Weber e t al., 1996), (3) qe = 1 + bCe d ue to heterogeneity of the multicomponent sorbent w hich transforms to the linearized form; s urfaces. Moreover, the presence of a hydrated o xide-type sorbent may delay the precipitation of a C e/qe = ( Q°)-lCe + ( Q°b)-I (4) m etal hydroxide in a saturated solution as, for w here the Langmuir parameters, Q0 (rag/g) and b e xample, in a suspension containing a silica sorbent ( L/rag), relating to monolayer adsorption capacity w here the binding of Cu (II) ions to the SiO2 sur- a nd energy of adsorption, respectively (Periasamy f ace would be preferred over precipitation (Park e t a nd Namasivayam, 1994), are found from the slope al., 1995). a nd intercept of C , /qe v s Ce linear plot such that T he capacities determined by column experiments QO= slope-t and b = intercept -t slope. Several line- w ere generally greater than those by batch tests, i.e. a rized isotherms with respect to the Langmuir QCOL> [ )batch , due to a number of reasons: e xp.- ~xp. m odel are shown in Figs 7 and 8. T he Langmuir parameters computed for all metal (i) the sorbent column consists of several transfer i on-sorbent combinations at 2 5°C a re summarized u nits, and the height equivalent to one theoreti-
Sorption of heavy metal cations 437 0 .020 i © O d Fw Cu Fw lib Fw 0 .015 o" 0 .010 Q 0 0 .005 A O.OOO ~ 8.0 0 .0 2.0 4.0 8.0 Ce(mWmL| Fig. 8. Selected isotherms linearized with respect to the Langrnuir model (fly ash). cal plate (HETP) may take quite low values in t o red muds and fly ashes (e.g. acid activation and efficient columns; s ubsequent heat treatment) did not significantly (ii) metal cations are partly held by ion-exchange increase the metal loading capacities unlike those of while passing through the column causing a C s + (Apak e t al., 1995) and orthophosphate (Shiao n atural pH gradient to develop across the col- a nd Akashi, 1977) adsorption by red mud. The urrm height, whereas pH is a rather conserved increased surface area of the pretreated sorbent was p roperty in batch tests; n ot reflected in sorption capacities. The only advan- (iii) a part of the sorbent surface may be covered tage of acid activation in this study seems to be the w ith a hydrous oxide gel containing the heavy p roduction of clean sorbents compatible with EPA m etal hydroxide as the elution proceeds, and r egulations (LaGrega e t aL, 1994b). this layer may promote further binding of the T he order of hydrolysable divalent metal cation metals enhancing sorption. r etention on the selected sorbents (which actually c ontained a mixture of hydrated oxides) were as fol- lows in terms of saturation capacities (mmoi/g): Cu G enerally very high limiting capacities have been > P b > Cd for fly ashes and Cu > Cd > Pb for achieved for metal sorption on to the selected r ed muds (see Table 1), with Pb (II) replacing Cd u nconventional sorbents giving rise to their possible ( II) in the sequence for the two sorbents. The utility in heavy metal removal from contaminated degree of the insolubility of the metal hydroxides water. All the observed metal cations sorption (expressed as the pKsp of the corresponding metal (except Cd (II) uptake by fly ash) took place at pH hydroxide) approximately followed the same order: values below the PZC of sorbents indicating specific a dsorption by the hydrous oxide gel layer as the M etal(II): Cu > Pb > Cd d ominant mechanism of adsorptive uptake (Apak pKsp o f M(OH)2 : 1 9 . 7 14.9 (13.7) 13.6 a nd Unseren, 1987; Apak e t aL, 1993, 1995, 1996) where (13.7) is the pKsp of Pb(OH)CI, probably r ather than electrostatic binding. The extremely high capacities of fly ash for Cu (II) and Pb (II) showing the role of heavy metal hydrolysis and hy- m ay be attributed to the contribution by surface drolytic precipitation in the observed uptake p recipitation. The pretreatment procedures applied sequence (Apak e t al., 1 993; TQtem and Apak,
438 R e,at Apak e t al. Table 2. The distributioncoefficientsof the metalsobtainedby batch tests for sorptionand desorption pH 4.75 (H2CO~)KdD ~" pH 7 (H2CO3,NaHCO3) fro ~' Metal Absorbent /t~D' (litres/g) (litres/g) (litres/g) Cd (II) F 0.372 249.7 624.3 Cd (11) Fw 0.329 148.8 111.6 Cd (II) Fa 0.090 144.8 217.2 Cd (II) Rw 0.026 I 13.8 227.6 Cd (II) Rah 0.026 112.0 86.2 Cd (II) Ra 0.016 107.4 82.6 Cu (II) F 0.132 223,5 335.2 Cu (II) Fw 0.128 218.8 328.2 Cu (II) Fa 0.125 189.3 656.8 Cu (II) Rw 0.045 150.0 69.2 Cu (II) Rah 0.035 146.3 125.4 Cu (II) Ra 0.014 109.0 65.4 Pb (II) F 0.126 52.6 10.5 Pb (II) Fw 0.127 49.1 377.4 Pb (II) Fa 0.093 48.3 966 Pb (II) Rw 0.024 317.8 -- Pb (II) 0.018 342.8 -- Rah Pb (II) Ra 0.015 395.0 -- b e leached out once retained in changing ground- 1995). Hydroxo-metal complexes and hydroxides w ater pH conditions, e.g. by CO2 injection. f ormed at a pH just below the precipitation limit T hus, these sorbents may serve as effective and t end to sorb on hydrated oxide-type sorbents with a lmost priceless fixation agents for heavy metal h igher affinity due to energetic reasons (Reed and r emoval from water prior to a more sophisticated C line, 1994). The correlation between the stability p rocedure such as solidification and stabilization as c onstant of the surface complex and that of t he means of the ultimate disposal. F o r example, h ydroxo-complex is linear, especially on a silica sur- w hen metal-loaded solid waste was added up to f ace (Park e t al., 1993). The much stronger adsorp- 2 0% by mass to Portland cement-based formu- t ion of Cu (II) on TiO2 (s) than of Cd (II) or Zn l ations, the fixed metals did not leach out from the ( II) has been attributed to the much lower solubility s olidified concrete blocks over extended periods, p roduct of Cu (OH)2 vs Cd(OH)2 or Zn(OH)2 w ith the exception of Cu (II), which reached a con- ( Zang e t al., 1994). Thus, there is a natural corre- c entration of 0.4 ppm after 8 months in a water lca- l ation as observed in this work between adsorbabil- c hate of pH 8-9 (Klhnqkale e t aL, 1997). A double- i ty of the metal and the pKsp of its hydroxide. The f old aim o f heavy metal fixation and metallurgical h igh capacity of fly ash for Pb (II) may have been s olid waste disposal would then be achieved with a dditionally affected by PbSO4 formation on the t he constraint that fly ashes better serve the purpose s orbent surface containing sulphate. o f heavy metal fixation than red muds. I f the utilized sorbents are suggested for use in r estricting the expansion of a metal contaminated CONCLUSIONS p lume in soil, then it will be necessary to show the l eachability of the retained metals from the sorbents I n investigation of the possibility of usage of met- u nder changing groundwater conditions. The poss- a llurgical solid wastes as cost-effective sorbents in i ble p H changes in groundwater have been modelled h eavy metal removal from contaminated water, red b y saturated aqueous carbonic acid (pH 4.75) and m uds and especially fly ashes have been shown to H 2CO3/NaHCO3 buffer (pH 7.0) solutions, the lat- e xhibit a high capacity for heavy metals with the s orption sequence Cu > Pb > Cd in accordance t er being prepared by bubbling CO2 through a w ith the order of insolubility of the corresponding 4 .0 x 10-3 M NaHCO3 solution until the solution m etal hydroxides. An empirical Langmuir approach b ecame neutral (pH 7.0). The distribution coeffi- c ould approximate isotherm modelling of metal c ients obtained by batch tests for limiting adsorp- s orption. The metals were essentially held irreversi- t ion (/~v ds') and for desorption (/~v~') with both b ly, and would not leach out into carbonic acid or c arbonic acid and pH 7,0 buffer solutions at room b icarbonate buffered solutions. The metal-loaded t emperature are listed for comparison in Table 2. s olid wastes could be solidified to an environmen- T he fact that the K~D values were in general 3-4 s" t ally safe form, thereby serving the double-fold aim o rders of magnitude higher than the /~vdS' values o f water treatment and solid waste disposal. c onfirmed the essential irreversible character of m etal adsorption (Park e t al., 1992; Apak e t al., 1995) on to the selected sorbents. Therefore, the R EFERENCES s uggested unconventional sorbents may be used in A pak V. and I:lnsvren E. (1987) Treatment of waste water c onfining a subsurface metal contaminant plume in and effluents with solid industrial wastes for the adsorp- tive removal of heavy metal contaminants. In a r estricted zone, and the retained metals would not
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