Heavy metal cation retention by unconventional sorbents (red muds and fly ashes)

Pergamon
PII: S0043-1354(97)00204-2

IVat. Res. Vol. 32, No. 2, pp. 430-440, 1998
© 1998 ElsevierScienceLtd. All rights reserved
Printed in Great Britain
0043-1354/98 $19.00 + 0.00

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

Al~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 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 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 saturation capacities of the sorbents. The sorption data have been analysed and fitted to linearized adsorption 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
Key words---cadmium (II), lead (II), copper (II), sorption, red muds, fly ashes

INTRODUCTION
Cadmium (II), lead (II) and copper (II) are wellknown toxic heavy metals which pose a serious

threat to the fauna and flora of receiving water
bodies when discharged into industrial wastewater.
In spite of strict regulations restricting their careless
disposal, these metal cations may still emerge in a
variety of wastewaters stemming from catalyst, electrical apparatus, painting and coating, extractive
metallurgy, antibacterials, insecticides and fungicides, photography, pyrotechnics, smelting, metal
eleetroplating, fertilizer, mining, pigments, stabilizers, alloy industries, electrical wiring, plumbing,
heating, roofing and building construction, piping,
water purification, gasoline additive, cable covering,
ammunition and battery industries (Buchauer, 1973;
Low and Lee, 1991; Periasamy and Namasivayam,
1994) and sewage sludge (Bhattacharya and
Venkobachar, 1984). The acute toxicity of these
heavy metals have caused various ecological catastrophes in human history, such as the "itai-itai"
disease due to cadmium (Riley and Skirrow, 1975).
Prolonged effect may cause other chronical disorders (Huang and Ostovic, 1978).
*Author to whom all correspondence should be addressed.

Various treatment technologies have been developed for the removal of these metals from water.
The hydrometallurgical technology extracts and
concentrates metals from liquid waste using any of
a variety of processes, such as ion exchange, electrodialysis, reverse osmosis, membrane filtration,
sludge leaching, electrowinning, solvent stripping,
precipitation and common adsorption (LaGrega et
al., 1994a).
Both powdered (Sorg et al., 1978) and granular
activated carbon (Huang and Smith, 1981) have
been used for the adsorptive removal of Pb, Cd and
similar "soft" heavy metals, especially when associated with common organic particulate matter in
water. Activated carbon from cheaper and readily

available sources, such as coal, coke, peat, wood,
nutshell (Freeman, 1989) and rice husk (Srinivason,
1986), may be successfully employed for the
removal of heavy metals from aqueous solutions.
Hydrous oxides such as alumina, iron oxides
(hematite and goethite) (Cowan et al., 1991; Gerth
and Bruemmer, 1983) manganese (IV) oxide
(Hasany and Chaudhary, 1986) and titanium (IV)
oxide (Koryukova et al., 1984) have also been used
for the adsorption of the indicated heavy metals.
The cost of the adsorptive metal removal process
is relatively high when pure sorbents (either acti-

430


431

Sorption of heavy metal cations
vated carbon or hydrated oxides) are used.
Therefore, there is an increasing trend for substituting pure adsorbents with natural by-product or
stabilized solid waste materials for the development
of cost-effective composite sorbents capable of
treating a variety of contaminants. For example,
recent evidence on the combined use of lime, ferric
and aluminium coagulants has shown that these
substances are more effective in combination than
individually (Harper and Kingham, 1992). A number of metallurgical solid wastes such as bauxite
waste red muds and coal-fired thermal plant fly
ashes have been screened in this regard to serve as

versatile and cost-effective sorbents for heavy metals
(Apak and (0nseren, 1987; Apak et al., 1993) and
radionuclides (Apak et al., 1995; 1996). The ability
of fly ash to remove metal cations from water has
also
been
demonstrated
in
the
literature
(Bhattacharya and Venkobachar, 1984; Panday and
Singh, 1985; Yadava et al., 1987) for a limited number of metals.
The alternative mechanism for heavy metal
removal by red muds and fly ashes (either natural
or in activated form) are assumed to comprise four
steps (Gregory, 1978; Apak a n d Llnseren, 1987). (i)
surface precipitation (sweep flocculation), where
most hydrolysable heavy metals are removed via coprecipitation of their insoluble hydroxides forming
successive layers on the sorbent surface; (ii) flocculation by adsorption of hydrolytic products, where
multi-nuclear hydrolysis products (formed on the
adsorbent surface as kinetic intermediates) including
[Fe2(OH)4 ]2+, [Fe3(OH)415+, [AI4(OH)s]4+ and
[AIs(OH)20]4 + act as more effective flocculants than
their parent ions due to their higher charge and
strong specific adsorptivities; (iii) chemical adsorp-

tion based on surface-complex formation, where
metal ions are usually removed as uncharged hydroxides condensed on to surfaces of - O H group
bearing adsorbents (Lieser, 1975), i.e. aluminium
oxide, silica gel, ferric and titanium oxides, existing

as components of the utilized composite sorbents;
(iv) ion exchange, where the acid-pretreated sorbents may function as synthetic cation exchangers.
Of these mechanisms, surface precipitation and
chemical adsorption are believed to play the dominant role in heavy metal ions removal (Apak and
Unseren, 1987).
The aim of the present study is to develop costeffective unconventional sorbents, preferably metallurgical waste solids, for heavy metal removal from
contaminated water. The heavy metal (Pb, Cd and
Cu) removal capacity as well as sorption modelling
of red muds and fly ashes will be evaluated in this
regard. The irreversible nature of sorption needs to
be shown so as to guarantee non-leachability of
metals from the metal-loaded sorbents.

EXPEilIMENTAL
Materials and methods

All heavy metal solutions (divalent cations Of Pb, CA
and Cu) were prepared in stock solutions up to 10000
ppm 0a g/ml) of metal from the corresponding nitrate
salts. No further pH adjustment of these solutions was
made as their natural acidity due to hydrolysis of metals
(i.e. to form MOIl + and H +) prevented the precipitation
of the corresponding metal hydroxides. All chemicals (E.
Merck, Darmstadt, Germany) were of analytical reagent
grade.
Of the metallurgical solid wastes used as sorbents, the
red muds were supplied from Etibank Seydi~ehir
Aluminum plant, Konya, Turkey and coal fly ashes
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

Metal iona

Adsorbent

LangmuirParametersb
Qo (rag/8)
Equilibrium
Qexp.(rag/g)
Qexp.(rag/g) Theoretical
pH
Columncapacity Batch capacity capacity
b (litre/mg)

Corr.coeff. (r)

C d (II)
C d (II)
C d (II)

F
Fw
Fa

7,2
6.7
6.6

220
-122

198.2
195,2
180,4

374.3
223.2
217.2

1.14.10-3
I.17.10-3
6.07.10-3

0.957
0.970
0,953

Cd OI)
Cd (II)
Cd (II)
Cu (If)
Cu (II)
Cu (II)
Cu (II)
Cu (II)
Cu (II)

Rw

gab
Ra
F
Fw
Fa
Rw
Rah
Ra

6.0
5.9
4.2
6.0
5.8
5.7
6.0
5.7
4.5

160
115
105
-264
187
110
100
63

66.8
66.8

46.9
207.3
205.8
198.5
75.2
65.2
35.2

113.7
112.0
107.5
335.2
328.2
283.9
90.0
87.8
65.4

0.57.10-3
0.65.10-3
0.11.10-3
0.94,10-3
0.75,10-3
0.73.10-3
0.96.10-3
0.79.10-~
1.00.10-3

0.958
0.994

0.989
0.968
0.961
0.960
0.958
0.956
0.964

Pb (II)
Pb (II)

F
Fw

6.2
6.0

530
--

444.7
483.4

526.0
490.7

I.I1.10-3
1.10.10-3

0.948

0.976

Pb (II)
Pb (II)
Pb (II)
Pb (II)

Fa
Rw
R.a
R,

6.0
6.0
5.7
4.4

-161
164
123

437.0
165.8
138.8
117.3

483.0
158.9
137.2
118.5

0.84.10-3
0.66.10-3
1.17.10-3
1.56.10-3

0.958
0.960
0.970
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
ashes,

bCalculatedby the aid of iinearizedLangmuirequation (4),


432

Re,at Apak et ,7l.
.0

,

,

,

,

,

,

5.0 t

A
[]

CdF
CdFw

4.0

~

caF,

3.0

Jo.'~

1.ot

0.0 ~

0.0

•

•

I

1.0

2.0

3.0

•

•

I

~]~ Cd~

I

~

~

4.0

5.0

"1

w

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.
Kahramanmara~, Turkey. The red muds, obtained as alkaline leaching wastes of bauxite in the Bayer process of
alumina manufacture, had the following chemical composition by weight: Fe203 37.3%, A1203 17.6%, SiO2 16.9%,
TiO2 5.6%, Na20 8.3%, CaO 4.4%, loss on ignition
7.2%. Red muds, being multicomponent systems, are composed of sodium aluminosilicates, kaolinite, chamosite,
iron oxides (hematite) and hydroxides. Basically, Fe is in
the form of hematite, Ti is in the form of Fe-Ti oxides
and AI is in the form of ahiminosilicates. 94% of red
muds have less than 10/an grain size.
The red muds were thoroughly washed with water to a
neutral pH, dried and sieved (R,) prior to adsorption
tests. The red muds were also acid-treated (R~). The acid
treatment was carried out according to a modified version
of Shiao's procedure (Shiao and Akashi, 1977) by boiling
I00 g of water-washed and dried red mud in 2 dm 3 of
10% (by weigh0 HCI solution for 2h, filtering off,
thoroughly washing with water, drying and sieving to
obtain the Ra-sorbents. The acid-treatment technique,
which has also been applied by Wahlberg et al. (1964) to
clay minerals for improving their surface properties, has
been demonstrated with success in synthesizing a better

adsorbent from red muds in phosphorus (Shiao and
Akashi, 1977) and heavy metal (Apak and Unseren, 1987;

Apak et al., 1995, 1996) removal. However, acid treatment
of red mud sorbents had the drawback of the partial loss
of acid-soluble fractions like hematite. The Ra fraction
was further subjected to heat treatment at 600°C for 4 h
to obtain the Rah sorbents. The red muds (partly agglomerated due to relative humidity) could not be classified
with respect to true grain size as most were of 200 mesh
size in wet sieving.
The specific areas of Rw, Ra and Rah samples were 14.2,
20.7 and 28.0 m2/g, respectively, measured by the BET/Nz
method (Brunauer et al., 1938).
Coal fly ash was recovered from the cyclones and electrostatic precipitators of the power plant and had the following average composition: CaO 42.5%, SiO2 21.9%,
SO3 13.6%, A1203 11.8%, Fe.zO3 2.4%, MgO 1.3%, K20
1.1%, Na20 0.9%, loss on ignition 4.4%. Almost 99% of
the fly ash could pass through a 200-mesh sieve. The raw
fly ash (F) was washed with 10-fold distilled water for several (5-6) times, filtered and dried (Fw). A part of the Fw
was further treated several times with acid using 2% (by
weight) HCI in boiling solution for 2 h. Higher acidity (as

8.0

[]

oaf

5.0

A

c~Fw

4.0

•

0.%

2.o3°1.o

•

0.0

0.0

0.2

0.4

0.6

0.8

1.0

1.2

C e ling/roLl
Fig. 2. Distribution coefficient of Cu (II) as a function of equilibrium aqueous concentration on fly
ashes and red muds.

Sorption of heavy metal cations

433

6.0
5.0

A

4.0

•

~'~w
PbF,,

~'Rw

1.0 ~
0.0
0.0

1.0

2.0

3.0

4.0

Ce

5.0

8.0

7.0

8.0

9.0

(m mLl

Fig. 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
losses of fly ash components by solubilization. The solid
product was thoroughly washed with water, filtered, and
oven dried at 100 + 5°C to produce the acid-treated (Fa)
sorbent.
The X-ray diffractogram (Apak et al., 1996) of the Fw
identified 51% calcite (CaCO3), 32% anhydrite (CaSO4),
9% quartz (SiO2) and 3% hematite (Fe203) in the crystal-

line phase. Elemental analysis of selected spots in the heterogenous amorphous slag particles by the XRF technique
(Apak et al., 1996) yielded 41-52% CaO, 27% SiO2, 13%
A1203, 2-5% FeO, 1-4% MgO and up to 2% other oxides.
The BET/N2 surface area of fly ash were 10.2 and

14.3 m2/g for Fw and Fa, respectively.

250

,'

o8

o

0

[]

CdF

A

ca%

•

cd%

150

g
@

)

[]

100

•

•

50

0
0
0
0.0

!

2.0

4.0
Co(msI L)

Fig. 4. Isotherm of Cd (II) adsorption onto fly ashes and red muds.

6.0


434

Re,at Apak et al.

160
[]
[]

120

[]

CuF

â

C.F a

đ c,e,,
c,e,

[]

â

[]
â
80

A

â

40

ml,

O
r

ã

0.4

0.0

0.8
Ce(mg/mL)

1.2

Fig. 5. Isotherm of Cu (II) adsorption onto fly ashes and red muds.
Point of zero charge (PZC) measurements by potentiometric titration of the sorbent suspensions at different
ionic strengths (Apak et aL, 1995, 1996) yielded approximate P Z C values of 6.4 and 8.3 for fly ash and red mud
sorbents, respectively.
When 1 g of sorbent was equilibrated with 50 ml distilled water, the indicated sorbents showed the following
approximate final pH in their aqueous leachates:

Rw Ra R~
pH

8.1

4.8

5.3

F
12.0

Fw

Fa

10.8 9.3

The acid-treated sorbents contained no free HCI but 1520 mg Cl-/g.
Batch sorption tests were carried out by agitating a suspension of I g sorbent in 50 ml metal nitrate solution for
8 h (equilibration period) at room temperature
(25 +0.1°C) in stoppered flasks placed on a thermostatic
water-bath/shaker. After centrifugation, the remaining
metal concentration in the filtrate was determined by
flame AAS (Perkin Elmer 300, Norwalk, CT, U.S.A.) and
the equilibrium pH was measured by a pH-meter
(Metrohm E-512 Herisau, Switzerland) equipped with a
glass electrode.
The metal concentration retained in the sorbent phase
(qe, mg/g) was calculated by
qe = (Co - G ) V / m

Ko = q o / G

(2)

where KD is the empirical distribution ratio of the metal
cation M ((mg/g)/(mg/litre)= litres/g) determined on the
approximately linear portion of the corresponding adsorption isotherm.
Batch desorption tests were carried out by agitating I g
of metal loaded sorbent with 50 ml of the desired solution
until equilibrium (8 h).
The saturation capacities of the sorbents for the uptake
of indicated metals were determined by both batch and
column tests. For the latter, 40 g of adsorbent was filled at
a height of 8-11 cm in a thermostatic (25 + 0.1°C) column
of dimensions (h = 30 cm, ~b = 3 cm), and the adsorbate
solution was fed (counter to gravity) by a peristaltic pump
through the fixed bed of sorbent at a constant rate of
0.5 ml/min. The metal concentration of the eluate was
recorded against throughput volume. The dynamic metal
uptake capacities of the sorbents were calculated by the integration technique (Apak et al., 1996), i.e. the area above
the curve up to the line on which the eluate concentration
was equalized with the initial concentration of metal was
calculated. The total amount of retained metal was divided
by the mass of sorbent to yield the saturation capacity
(t.)col. x

(1)

where Co and C~ are the initial and final (equilibrium) concentrations of the metal ion in solution (mg/litre), V is the
solution volume (litres) and m is the mass of sorbent (g).
The solid/water distribution ratios (at equilibrium) of
metals for both sorption and desorption were calculated

by

RESULTS AND DISCUSSION

In weakly acidic-neutral suspensions whose p H
was attained naturally by equilibrating aqueous
metal n i t r a t e - s o r b e n t mixtures, the distribution
ratios generally increased with initial aqueous


Sorption of heavy metal cations

435

500

[]

[]

400

[]
A

[]

0

0

[]

[]
&

300

I:bF

@

O0

~aa,
ha,,

0
0

0
0

200

•

•1

0.0

i

1~

2.0

.i

,

n

II

,

|1

4.0

II

I

•

•

6.0

CelmllnnU

0.0

10.0

Fig. 6. Isotherm of Pb (II) adsorption onto fly ashes and red muds.
adsorbate concentration at equilibrium up to a limiting value where the batch capacities of the sor'nbatch~
bents for the metals were calculated /~exp.
~. The
saturation capacities found by both batch and
dynamic column experiments (the latter symbolized
as Q~L.) are listed in Table 1. The variation of distribution ratio (KD) with equilibrium concentration
of the adsorbate in solution is shown in Figs 1, 2
and 3, where KD vs C~ on semi-logarithmic scale
gave roughly linear plots. A gradual decrease of distribution ratios with aqueous concentration was
noted, due to increased occupation of active surface
sites of sorbent with metal loading in the aqueous
solutions (Apak et al., 1995, 1996). As long as a
strict differentiation between true adsorption and
precipitation- masked sorption (McKay et al., 1985)
is not made, both red muds and fly ashes may be
visualized as effective sorbents capable of removing
the studied heavy metal ions from solution with
high distribution ratios, /~Das', ranging up to 10-210-1 litres/g. Naturally the slightly alkaline character of aqueous leachate obtained from fly ashes in
conjunction with the CaO and CaSO4 constituents
of this material should account for hydrolytic metal
precipitation reactions (Burgess, 1978; Freeman,
1988) as well as counter-ion adsorption at pH
above PZC accompanying chemical adsorption

(Apak and Unseren, 1987; Apak et al., 1993).
Generally a metal hydroxide may precipitate and

may form at the surface of the hydrous oxide sorbent prior to its formation in bulk solution and
thus contribute to the total apparent sorption. The
contribution of surface precipitation to the overall
sorption increases as the sorbate/sorbent ratio is
increased (Stumm and Morgan, 1996). It should be
added that raw fly ash (F) cannot be considered as
an EPA-acceptable sorbent (LaGrega, 1994b) as it
introduces new contaminants to water in untreated
(either water- or acid-washed) form.
The adsorption isotherms (q~ vs Ce) of metal
uptake at 25°C (see Figs 4, 5 and 6) essentially
showed BET (Branauer et aL, 1938) (type II, V)
character curves pointing out to the heterogeneity
of the sorbents containing hydrous oxides, silicates
and sulfates, resulting in various combinations of
linear and nonlinear isotherms (Weber et al., 1996).
It is known from the literature that BET type IV-V
isotherms are quite common for the porous hydroxides (xerogels) such as siligagel or iron hydroxide
(Gregory, 1978).
Although Langmuir and Frenndlich approximations of the observed adsorption data in the linearized
forms
gave
satisfactory
correlation
coefficients (r > 0.95) for most of the covered concentration range, the Langmuir model had more
practical utility for representing the limiting sorption capacities of the sorbents than the exponentially increasing Freundlich isoterm (McKay et aL,

436

Re,at Apak et al.

0.08

e
II

cde w
cue w

~ew
0.06

Q

0

0.04

0.02

0.00 V
0.0

4.0

2.0

6.0

8.0

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
Langmuir assumptions, i.e. site-specific and uniformly energetic adsorption confined to monolayer
coverage (Weber and DiGiano, 1995; Weber et aL,
1996). Heavy metal adsorption on heterogenous
sorbents has been interpreted by the aid of the
Langmuir isotherm on various occasions in the environmental literature (Szymura, 1990; Prasad and
Agarwal, 1991).
A Langmuir equation for adsorption may be
written as
Q° b C~
qe = 1 + bCe

(3)

which transforms to the linearized form;
Ce/qe = (Q°)-lCe + (Q°b)-I

(4)

where the Langmuir parameters, Q0 (rag/g) and b
(L/rag), relating to monolayer adsorption capacity
and energy of adsorption, respectively (Periasamy
and Namasivayam, 1994), are found from the slope

and intercept of C, /qe vs Ce linear plot such that
QO= slope-t and b = intercept -t slope. Several linearized isotherms with respect to the Langmuir
model are shown in Figs 7 and 8.
The Langmuir parameters computed for all metal
ion-sorbent combinations at 25°C are summarized

in Table 1 (three runs made per isotherm) together
with the experimental saturation capacities of batch
rn~t~h~ and column (Q~exp.)
ol. tests. After screening of
those results where metal hydroxide precipitation
could have been effective in metal removal (e.g.
modelling of Cu(H) sorption has been made up to
the concentration edge of Cu(OH)2 precipitation at
the studied pH), Langmuir modelling has been
quite successful in predicting the experimental saturation capacities of the sorbents, especially those
obtained from dynamic column tests (see Table 1),
although its basic assumptions were not fulfilled
(Weber and DiGiano, 1995; Weber et al., 1996),
due to heterogeneity of the multicomponent sorbent
surfaces. Moreover, the presence of a hydrated
oxide-type sorbent may delay the precipitation of a
metal hydroxide in a saturated solution as, for
example, in a suspension containing a silica sorbent
where the binding of Cu (II) ions to the SiO2 surface would be preferred over precipitation (Park et
al., 1995).
The capacities determined by column experiments
were generally greater than those by batch tests, i.e.
QCOL
> [)batch

exp.~xp. , due to a number of reasons:
(i) the sorbent column consists of several transfer
units, and the height equivalent to one theoreti-


Sorption of heavy metal cations

0.020

437

i

©

O d Fw

Cu Fw
lib Fw

0.015

o"
Q

0.010

0

0.005

A

O.OOO ~

0.0

2.0

4.0

Ce(mWmL|

8.0

8.0

Fig. 8. Selected isotherms linearized with respect to the Langrnuir model (fly ash).
cal plate (HETP) may take quite low values in
efficient columns;
(ii) metal cations are partly held by ion-exchange
while passing through the column causing a
natural pH gradient to develop across the colurrm height, whereas pH is a rather conserved
property in batch tests;
(iii) a part of the sorbent surface may be covered
with a hydrous oxide gel containing the heavy
metal hydroxide as the elution proceeds, and
this layer may promote further binding of the
metals enhancing sorption.
Generally very high limiting capacities have been

achieved for metal sorption on to the selected
unconventional sorbents giving rise to their possible
utility in heavy metal removal from contaminated
water. All the observed metal cations sorption
(except Cd (II) uptake by fly ash) took place at pH
values below the PZC of sorbents indicating specific
adsorption by the hydrous oxide gel layer as the
dominant mechanism of adsorptive uptake (Apak
and Unseren, 1987; Apak et aL, 1993, 1995, 1996)
rather than electrostatic binding. The extremely
high capacities of fly ash for Cu (II) and Pb (II)
may be attributed to the contribution by surface
precipitation. The pretreatment procedures applied

to red muds and fly ashes (e.g. acid activation and
subsequent heat treatment) did not significantly
increase the metal loading capacities unlike those of
Cs + (Apak et al., 1995) and orthophosphate (Shiao
and Akashi, 1977) adsorption by red mud. The
increased surface area of the pretreated sorbent was
not reflected in sorption capacities. The only advantage of acid activation in this study seems to be the
production of clean sorbents compatible with EPA
regulations (LaGrega et aL, 1994b).
The order of hydrolysable divalent metal cation
retention on the selected sorbents (which actually
contained a mixture of hydrated oxides) were as follows in terms of saturation capacities (mmoi/g): Cu
> Pb > Cd for fly ashes and Cu > Cd > Pb for
red muds (see Table 1), with Pb (II) replacing Cd
(II) in the sequence for the two sorbents. The
degree of the insolubility of the metal hydroxides

(expressed as the pKsp of the corresponding metal
hydroxide) approximately followed the same order:
Metal(II):
pKsp of M(OH)2 :

Cu > Pb >
1 9 . 7 14.9 (13.7)

Cd
13.6

where (13.7) is the pKsp of Pb(OH)CI, probably
showing the role of heavy metal hydrolysis and hydrolytic precipitation in the observed uptake
sequence (Apak et al., 1993; TQtem and Apak,


438

Re,at Apak et al.
Table 2. The distributioncoefficientsof the metalsobtainedby batch tests for sorptionand desorption

Metal

Absorbent

Cd (II)
Cd (11)
Cd (II)
Cd (II)
Cd (II)

Cd (II)
Cu (II)
Cu (II)
Cu (II)
Cu (II)
Cu (II)
Cu (II)
Pb (II)
Pb (II)
Pb (II)
Pb (II)
Pb (II)
Pb (II)

F
Fw
Fa
Rw
Rah
Ra
F
Fw
Fa
Rw
Rah
Ra
F
Fw
Fa
Rw

Rah

Ra

/t~D' (litres/g)
0.372
0.329
0.090
0.026
0.026
0.016
0.132
0.128
0.125
0.045
0.035
0.014
0.126
0.127
0.093
0.024
0.018
0.015

1995). Hydroxo-metal complexes and hydroxides
formed at a pH just below the precipitation limit
tend to sorb on hydrated oxide-type sorbents with
higher affinity due to energetic reasons (Reed and
Cline, 1994). The correlation between the stability
constant of the surface complex and that of

hydroxo-complex is linear, especially on a silica surface (Park et al., 1993). The much stronger adsorption of Cu (II) on TiO2 (s) than of Cd (II) or Zn
(II) has been attributed to the much lower solubility
product of Cu (OH)2 vs Cd(OH)2 or Zn(OH)2
(Zang et al., 1994). Thus, there is a natural correlation as observed in this work between adsorbability of the metal and the pKsp of its hydroxide. The
high capacity of fly ash for Pb (II) may have been
additionally affected by PbSO4 formation on the
sorbent surface containing sulphate.
If the utilized sorbents are suggested for use in
restricting the expansion of a metal contaminated
plume in soil, then it will be necessary to show the
leachability of the retained metals from the sorbents
under changing groundwater conditions. The possible p H changes in groundwater have been modelled
by saturated aqueous carbonic acid (pH 4.75) and
H2CO3/NaHCO3 buffer (pH 7.0) solutions, the latter being prepared by bubbling CO2 through a
4.0 x 10-3 M NaHCO3 solution until the solution
became neutral (pH 7.0). The distribution coefficients obtained by batch tests for limiting adsorption (/~vds') and for desorption (/~v~') with both
carbonic acid and pH 7,0 buffer solutions at room
temperature are listed for comparison in Table 2.
The fact that the K~Ds" values were in general 3-4
orders of magnitude higher than the /~vdS' values
confirmed the essential irreversible character of
metal adsorption (Park et al., 1992; Apak et al.,
1995) on to the selected sorbents. Therefore, the
suggested unconventional sorbents may be used in
confining a subsurface metal contaminant plume in
a restricted zone, and the retained metals would not

pH 4.75 (H2CO~)KdD
~"
(litres/g)

pH 7 (H2CO3,NaHCO3) fro~'
(litres/g)

249.7
148.8
144.8
I 13.8
112.0
107.4
223,5
218.8
189.3
150.0
146.3
109.0
52.6
49.1
48.3
317.8
342.8
395.0

624.3
111.6
217.2
227.6
86.2
82.6
335.2

328.2
656.8
69.2
125.4
65.4
10.5
377.4
966
----

be leached out once retained in changing groundwater pH conditions, e.g. by CO2 injection.
Thus, these sorbents may serve as effective and
almost priceless fixation agents for heavy metal
removal from water prior to a more sophisticated
procedure such as solidification and stabilization as
the means of the ultimate disposal. F o r example,
when metal-loaded solid waste was added up to
20% by mass to Portland cement-based formulations, the fixed metals did not leach out from the
solidified concrete blocks over extended periods,
with the exception of Cu (II), which reached a concentration of 0.4 ppm after 8 months in a water lcachate of pH 8-9 (Klhnqkale et aL, 1997). A doublefold aim o f heavy metal fixation and metallurgical
solid waste disposal would then be achieved with
the constraint that fly ashes better serve the purpose
o f heavy metal fixation than red muds.
CONCLUSIONS
In investigation of the possibility of usage of metallurgical solid wastes as cost-effective sorbents in
heavy metal removal from contaminated water, red
muds and especially fly ashes have been shown to
exhibit a high capacity for heavy metals with the
sorption sequence Cu > Pb > Cd in accordance
with the order of insolubility of the corresponding

metal hydroxides. An empirical Langmuir approach
could approximate isotherm modelling of metal
sorption. The metals were essentially held irreversibly, and would not leach out into carbonic acid or
bicarbonate buffered solutions. The metal-loaded
solid wastes could be solidified to an environmentally safe form, thereby serving the double-fold aim
of water treatment and solid waste disposal.
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