Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals

Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals

CLAY-03986; No of Pages 18 Applied Clay Science xxx (2016) xxx–xxx Contents lists available at ScienceDirect Applied Clay Science journal homepage: ...

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CLAY-03986; No of Pages 18 Applied Clay Science xxx (2016) xxx–xxx

Contents lists available at ScienceDirect

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Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals A.A. Zaki ⁎, M.I. Ahmad, K.M. Abd El-Rahman Hot Lab. Center, Atomic Energy Authority of Egypt, P.O. 13759, Inshas, Cairo, Egypt

a r t i c l e

i n f o

Article history: Received 18 April 2016 Received in revised form 7 September 2016 Accepted 11 September 2016 Available online xxxx Keywords: Silty clay soil Landfill barriers Sorption mechanisms Kinetics Geohydrology

a b s t r a c t The sorption characteristics of landfill silty clay soil (LSC), taken from AL Hammam landfill site, as a retardation barrier of Zn2+, Cd2+ and Pb2+ ions were investigated. LSC consists mainly of kaolinite, quartz and calcite minerals, is a part of thin blanket of Miocene rocks forming a vast persistent limestone plateau of the north part of the western desert of Egypt. The results showed that the percentage equilibrium uptake of the metal ions by LSC are 94.8, 92.7, and 86.0 for Zn2+, Pb2+, and Cd2+, respectively. The coefficient of diffusion's value was found in the range (3.32–6.8)0.10−17 m2/s and increases with the increase in temperature. The distribution coefficient for Pb2+, Cd2+, and Zn2+ ranged from 404.9 to 568, 230.1 to 281.9, and 371.4 to 466.7 ml/g in the temperature range 298 to 333 ± 1 K, respectively. The value of retardation factor and the sorption affinity onto LSC took the order Zn2+N Pb2+ N Cd2+. The experimental investigation on ionic concentrations in sorption batches suggested that sorption behaviors of Zn2+, Pb2+, and Cd2+ ion metals onto LSC are mainly controlled by cation exchange. The wetting front of water movement in the LSC as an unsaturated soil reached to about 0.06, 0.19 and 0.25 cm after 6, 24 and 48 h of steady infiltration. The saturation hydraulic conductivity of the LSC (b2.0 μm) fraction is 2.18 × 10−10 m/s therefore, it matches the condition of suitability of soils as mineral liners for a landfill facility. The Pêclet number values (b 32) indicate dominance of dispersion over advection. The b2 μm fraction of LSC may be used to attenuate Zn2+, Cd2+ and Pb2+ ions presented in AL Hammam landfill leachate from reaching the shallow groundwater. © 2016 Elsevier B.V. All rights reserved.

1. Introduction Over the past thirty years, there was a lot of interest on the pollution of the environment and its effects on the public health especially the diseases. According to the world health organization (WHO), prolonged exposure to the polluted environment threatens the health of about the one fourth of the people (Kimani N. G., 2007). The presence of the heavy metals in the aquatic ecosystem has been of increasing concern because of their toxic properties and other adverse effects on receiving water use (Qaiser et al., 2009; Gammoudi et al., 2014; Wanlu et al., 2014). Under favorable pH and other conditions, metals became soluble in water, contaminating water bodies. Cd2 +, Zn2 +, and Pb2 + are among the toxic metals; Cd2+ is known to cause lung insufficiency, bon lesion, and hypertension (Mohan and Singh, 2002). Low levels of Pb2 + have been identified with anemia, as it causes injury to the blood-forming system, while high levels cause severe dysfunction of the kidneys, the liver, and the central and peripheral nervous system and high blood pressure (Sarkar, 2002). Cadmium finds its way to water bodies through wastewater from metal plating industries, cadmium-nickel batteries, ⁎ Corresponding author. E-mail address: [email protected] (A.A. Zaki).

phosphate fertilizer, mining, pigments, stabilizers and alloys (Zheng et al., 2007). On the other hand, waste containing zinc and its compounds arise from many industrial processes, such as acrylic fiber, rayon, cellophane, and special synthetic rubber (Bradl, 2005). Moreover, the national cyclotron at Inshas site produces 67Ga and 111In isotopes from irradiation of metallic zinc and cadmium by accelerated protons (Reddad et al., 2002). Alexandria Governorate of Egypt, generates about 3000 t/day of municipal waste (MW) of them about 46 t/day metal and heavy metals. Approximately 80% of the total amount of MW is disposed of in either Borg El-Arab or Al-Hammam landfills. Al- Hammam landfill site receives a daily amount of 1800 to 2600 t/day of waste (Integral Consult; 2005). Landfilling is environmentally favorable choice, even with an active and a successful waste reduction program. Various treatment technologies have been developed for the purification of water, wastewater and leachate contaminated by heavy metals. The most commonly used methods for the removal of metal ions from industrial effluents include solvent extraction, reverse osmosis, ultrafiltration, adsorption and ion exchange (Wilson et al., 2006; Khelifa, 2013). The geological barrier of heavy metals and municipal solid waste landfills is a key factor for the protection and safety issues in waste

http://dx.doi.org/10.1016/j.clay.2016.09.016 0169-1317/© 2016 Elsevier B.V. All rights reserved.

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

2

A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

disposal. The functions of the barrier are the confining of the wastes and the buffering or attenuation of the hazardous leachate during and after the lifetime of the landfill to reach the groundwater. For this purpose, geological materials rich in clay minerals are considered among the suitable geological barriers in landfills (Rowe et al., 1997; Visvanathan et al., 2007; Lehman, 1982). Sorption processes of various heavy metals on soils and clay minerals have been studied by several authors (El-Kamash, 2005; Deng et al., 2013; Abdel Aziz, 2005; HO and McKay, 1998). The sorption process of ions on soils as silty materials provides retardation of the migration of nuclides from the contaminated site to the biosphere (Lijuan et al., 2013; Takamitsu et al., 2013). The study of sorption of heavy metals is an important part of the overall investigation required for the assessment of potential sites for waste disposal (Vandenbossche et al., 2014; Benedicte et al., 2012). However, the application of these studies in a large scale requires knowledge of the equilibrium and kinetics of the metal adsorption process. In order to protect the 5 m shallow groundwater, the aim of this work was to study the feasibility and effectiveness of Al Hammam landfill silty clay soil (LSC) as a bottom liner barrier and as a hydraulic barrier and a sorbent for mobilization and retardation of heavy metals such as Pb2 +, Cd2 + and Zn2 + present in the leachate of the landfill. The transport,sorption and desorption kinetics, mechanisms, capacity of LSC for the studied metals, saturated and unsaturated hydraulic behavior, distribution and diffusion coefficients, and other parameters were addressed.

performed in a Micrometrics Gemini III 2375 Surface Area Analyzer device using the BET method. Reference standards, blanks and duplicates were used to obtain precise and accurate analytical data. The relative standard deviations of the analytical data were measured to b5%. 3.2. Soil samples and aggregates fractionation The LSC samples used in this study were collected from Al Hammam landfill site Alexandria Governorate, Egypt. The LSC samples of the layer 0–50 cm depth were collected from a site in the vicinity of the landfill site. After mixing, all the bulk samples were air dried at room temperature, roots and other large particles were removed, then ground and sieved through a 2 mm nylon screen. The samples were partitioned into four aggregate-size fractions: 2–0.25 mm, 0.25–0.05, 0.05–0.002 and b 0.002 mm. Before fractionation, 50 g of bulk soil was immersed in 150 ml DDW (in a conical flask, 250 ml) and dispersed with ultrasonic. The ultrasonic energy was set to 100 J/ml according to (Schmidt et al., 1999). Then the 2–0.25 mm fraction was obtained by wet sieving using a vibrator RETSCH AS 200, the 0.25–0.05 and 0.05–0.002 mm fractions were obtained by sedimentation and siphoning, during times determined by Stokes' law. The b 0.002 mm fraction was finally obtained after centrifuging. All the soil particles of four size fractions were collected respectively and freeze-dried for later experiments. 3.3. Physical and chemical analysis of the LSC

Al-Hammam landfill site is located approximately 80 km from the city of Alexandria Egypt, 35 km to the south of the coastal road. The coordinates of the site are 30°45′22.5″ N and 29°25′16.0″ E. The total area of the site is 1.190 km2 (1700 m × 700 m). The depth to groundwater at the landfill site was found to vary from b 5 m in the south and increases northwards to reach 10 m as shown in Fig. 1a (Abu-Zeid et al., 2009). The soil and subsoil profiles of a 20 m borehole at the Al Hammam landfill site is shown in Fig. 1b. The soil profile at the landfill site consists of the following main layers; silty clay extends from the ground surface to a depth of 8 m having a coefficient of permeability of 1.0 × 10−8 m/s. This silty clay was used as a bottom liner of the landfill and in leachate evaporating ponds. In addition to this clay, a 4 m thick from 8 to 12 m was a sandy silt of calcareous nature with some clay and limestone fragments existed. A whitish grey color limestone was encountered at a depth of about 17 to the 20 m bottom of the borehole (Integral Consult; 2005). The main source of leachate in Al-Hammam landfill site is liquid waste, refuse moisture content and rain water. Leachate waste contains high level of chlorides, ammonium, metals, inorganic and organic compounds, nutrients (Integral Consult; 2005).

The pH of AL Hammam LSC was measured in deionized water at a soil/solution ratio of 2:5 using HI 3221 pH meter (Hanna instruments Inc., USA).Soil organic carbon (SOC) was determined by K2Cr2O7 digestion method (Nelson and Sommers, 1982). Cation exchange capacity (CEC) of the LSC soil was determined by BaCl2 displaced method, free iron oxide was determined by Na2S2O4 - Na3C6O7 extraction method (Anderson and Jenne, 1970). The crystallinity, of the LSC were measured using X-Ray Diffraction (XRD) using a Shimadzu X-ray diffractometer, model XD610 (Japan) at room temperature, using Bragg-Brentano geometry, with a back monochromatized Cu Kα radiation. Sample was very lightly ground and mounted on a flat sample plate. The diffraction pattern was scanned over the range 8–90 (2θ) in step of 0.031(2θ) and a counting time of 10 s per step. The unit-cell parameters were refined by a least –square procedure. The chemical and mineralogical compositions of LSC was characterized by X-ray fluorescence (XRF) using AXIOS Analytical 2005. The measurements of the phase changes and weight losses of the LSC sample were studied using Shimatzu DTA-TGA system of type DTA50 and TGA50 Japan, respectively, at heating rate of 10 K/min in presence of nitrogen gas to avoid thermal oxidation of the sample. Total porosity (% pore space) of the AL Hammam LSC was calculated from the bulk density and soil bulk density was determined using the method described by (Braja, 1983).

3. Materials and methods

3.4. Determination the saturated hydraulic conductivity of LSC

3.1. Chemicals and reagents

The saturated hydraulic conductivity tests were carried out following the procedure described in ASTM D5084-1991 using a Tecnotest permeameter apparatus. The LSC sample was placed between two porous discs in a cell, and a standpipe was attached to the top of the sample. The dimensions of the sample were 60 mm in diameter and 25 mm in height. The initial head difference h1 at time t = 0 was recorded, and then the liquid was allowed to flow through the LSC in order to obtain h2 in the final head at time t. The saturated hydraulic conductivity (Ks) governed by the law of Darcy, was calculated by the following equation (Braja, 1983):

2. Location, groundwater and geology and leachate of Al- Hammam landfill site

All materials used were of analytical grade and used without further purification. All solutions were prepared using distilled, deionized water (DDW) with a resistivity of 18 MΩ/cm. All stock solutions were stored in polypropylene bottles prior to use. All glassware and plastic ware was soaked in 5 M HNO3 (Fisher Scientific) and rinsed several times with DDW prior to use. All experiments were carried out in triplicate and the mean values are presented. The measuring of sample solution concentrations was within ±1–2% statistical accuracy. The used heavy metals; zinc chloride, cadmium chloride, lead chloride and malic acid (MA), were BDH products. The surface area of LSC soil was obtained by N2 physisorption (BET method) at 77 K. These analyses were

K s ¼ 2:303 ða  L=A  t Þ logðh1 =h2 Þ

ð1Þ

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

3

(a)

(b)

Fig. 1. (a) Location, groundwater contour map and (b) lithologic log of a 20 m borehole of Al Hammam landfill site, Alexandria, Egypt.

where Ks is the saturated hydraulic conductivity (m/s), A is the crosssectional area of the sample (m2), a is the cross-sectional area of the standpipe (m2), and L is the length of sample (m). Once the sample

has been prepared in the test cell, the sample is consolidated and saturated for 24 h duration. The duration of hydraulic conductivity test using DDW was about 120 h.

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

3.5. Soil water retention curve (SWRC) It is a soil water characteristic graph, illustrates the relation between soil water tension and soil water content in the unsaturated zone, θ (h), h is soil suction pressure head (cm). It is used to determine an index of the available water in soil and to classify soils for irrigation, capillary conductivity, capillary and thermal conductivities, clay and organic matter contents, construction and other applications (Vitalis et al., 2014).To construct a SWRC for AL Hammam silty clay soil, the samples first were DDW saturated in a pressure plate extractor; 1500F1 Pressure Plate Extractor from The Netherlands (Sun et al., 2006). Then the LSC samples were weighed to get the saturated water content, θs. The LSC samples were placed on the saturated ceramic disk of the ceramic plate extractor, ensuring good contact between the LSC samples and the disk. Then the LSC samples were subjected to a stepwise increasing matric suction to obtain the desorption branch of the SWRC. The water out flow was monitored in the burette until equilibrium was achieved in about 95 h. The LSC samples were taken out the for weighing to get the residual water content, θr,. Brooks-Corey model (Brooks and Corey, 1964), a nonlinear curve fitting model was used to fit experimental data and get the parameters of SWRC. The Brooks-Corey functions can be defined as:

column was filled with about 2.7 g LSC of the fraction 50–200 μm particle diameter. The LSC was put into the column carefully so that no air remained between the particles, which helped attaining compact layers. Up stream flow, using a high pressure pump (Nihon Seimitsu Kagaku Co. Ltd. Japan, model SP-D-2502 U), was used to avoid the problem of clogging of the plate. The experiments were carried out at isothermal conditions (T = 298 ± 1 K). The sorption of Pb2 +, Cd2 + and Zn2 + metal ions by the LSC was carried out on leachate (10−4 M NaCl) with initial concentration of 100 mg/l of the metal ions. Solution flow rate was maintained at a constant value through the column (Q = 0.1 cm3/min). At selected time intervals, Pb2+, Cd2+ and Zn2+ metal ions concentrations were determined in the outlet solution (effluent). The process was stopped when the concentration of ions in the effluent became equal to the initial concentration in the influent. During extraction, the metal ion concentrations values were measured in the influent and the raw effluent using the AAS. In order to study the effect of particle size on the dynamic sorption process, the column was packed with about 2.9 g of the total range of LSC particle size distribution including the b 2 μm, under the same conditions. 3.9. Sorption and desorption parameters estimation

where the inverse of air entry potential value (bubbling pressure), α, and the pore size distribution coefficient, λ (log slope of SWRC) estimated from the code RETC code (Van Genuchten et al., 1991).

Continuous sorption loading of Pb2 +, Cd2 + and Zn2 + metal ions (100 mg/l) in leachate solution to an initially solute free saturated LSC (θ = 0.48) column of length 5.0 cm and 0.50 cm inner diameter was achieved. Desorption of Pb2+, Cd2+ and Zn2+ metal ions from AL Hammam LSC was applied using solute free of 0.2 M malic acid (MA) at pH 5.0 ± 0.1 and at room temperature (298 ± 1).

3.6. Contact time and kinetic measurements under the saturation conditions

3.10. Error analysis

Batch sorption studies of zinc cadmium and lead ions were performed at different temperatures (298, 313, 323 and 333 ± 1 K) to obtain the sorption kinetics. In all kinetic and equilibrium experiments, 0.5 g of the LSC was equilibrated with 100.0 ml of a simulated leachate (leachate);10− 4 M NaCl containing the heavy metal at tracer level (100 mg/l for each metal ion) in stoppered polyethylene bottles. All the bottles were agitated on thermostat shaker adjusted at the desired temperature and the heavy metals were measured for different time intervals. All the solutions were centrifuged at 6000 rpm, MLW Zentrifugenban Co. (Germany), for 30 min. The supernatant was filtered and the concentration was measured using Atomic Absorption spectrometer (AAS), (Buck scientific model VGP 210). Each kinetic run provided the relative concentration of the solution at different elapsed time intervals. Temperature was adjusted over the range 298–333 ± 1 K and the pH at 5.0 ± 0.1. The heavy metal ion concentration retained in the solid phase (LSC) (mmol/g) was calculated using:

A normalized standard deviation (SD) was calculated in order to quantitatively compare the applicability of the kinetic models, by using Eq. (4) (Wang et al., 2015).

−λ

θðhÞ ¼ θr þ ðθs −θr Þ ðαhÞ

qe ¼

ðC o −C e Þ  V=M

ð2Þ

ð3Þ

SD ¼

X h 1=2  i2 qt; exp −qt;cal =qt; exp =ðn−1Þ

ð4Þ

where qt,exper and qt,calc are the experimental and calculated values for the amount of heavy metal ions (mmol/g) adsorbed per unit LSC mass at time t (min) and n is the number of data points (Wang et al., 2015). If the data from the model are similar to experimental data, SD approaches zero. A mean square error (MSE), was used to estimate the concentration, diffusion coefficients, the dispersivity and average particle velocity of saturated LSC column. n  2 X MSE ¼ 1=n ðC=C o Þexp −ðC=C o Þcal

ð5Þ

i¼1

where C0 and Ce are the initial and equilibrium concentrations of metal ion in solution (mmol/l), V is the volume (l) of the leachate and M is the weight (g) of the LSC.

As shown in Eq. (5) MSE is the difference between a relative concentration profile obtained from column test results and the corresponding values calculated from theoretical equations.

3.7. Sorption mechanism

4. Results and discussions

To understand the sorption mechanism(s) for the retardation of Zn2+,Cd2+, and Pb2+ heavy metal ions by LSC, Ca2+, Mg2+, Na+, and K+ metal ions were analyzed in addition to target heavy metal in the solutions using ICP-AES during and under the conditions of the kinetic experiments.

4.1. Characterization of LSC

3.8. Dynamic sorption and the effect of the particle size distribution The experiments were carried out in a pyrex glass column provided with a porous glass plate of inner diameter 0.5 cm, and 5 cm height. The

4.1.1. Physical, chemical and hydrological analysis of the LSC LSC was characterized before interaction with Pb2+, Cd2+ and Zn2+ metal ions. Fig. 2a shows the resulting diffractogram of X-ray diffraction for the bulk LSC. XRD diffractogram shows that the LSC is formed from group of non clay minerals such as calcite (basal spacing 3.03 Å) and Quartz (basal spacing 4.27 Å, 3.34 Å, 2.46, and 1.54 Å) and kaolinite clay (basal spacing 7.10 Å) mineral. It was observed that the principal components of LSC are quartz, kaolinite and calcite which are in good

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

1100

Q

900

Intensity

Q= quartz K=Kaolinite C=Calcite

(a)

700

C Q

500

(b)

Q

K K

C K K C

5

H2O

C

K Q=quartz

Q

K=Kaolinite 300

K

K

C=Calcite K

100

8

K

K

16 24 32 40 48 56 64 72 80 88

2 (deg)

Percent Finer (%)

100

Experimental Fredlund-Xing fitting Equation

(d)

80 Sedimentation 60

Sieving Siphoning

40 Centri 20 fugation

0 0.001

0.01

0.1

1

Particle size (mm) Volumetric water content (-)

0.5 0.10 0

(e)

0.4

0.20

0.25

0.30

0.35

0.40

0.2

6 hrs

-0.1

0.3

0.1

0.15

s

Vertical depth (cm)

Soil water content, θ

θ

Brooks-Corey fitting Experimental

-0.2

24 hrs

-0.3 -0.4

(f) 48 hrs

-0.5

θr

0

0.5

1.0 1.5 Log suction (h) (cm)

2.0

2.5

-0.6 0.10

0.15

0.20

0.25

0.30

0.35

0.40

Fig. 2. (a) X-ray diffractogram. (b) FTIR Spectrum, (c) DTA-TGA thermogram, (d) particle size distribution, (e) soil water retention curve (SWRC) and (f) Infiltration of water through the AL Hammam LSC landfill of unsaturated zone beneath the disposal facility.

agreement with the chemical and mineral analysis; clay (43%), silt (42%), and sand (15%). Some basic properties of LSC such as particle size (mm), soil organic carbon (g/kg) (SOC), cation exchange capacity (mol(+)/kg) (CEC), free ion oxide (g/kg) (FIO), surface area (m2/g) (SA) and saturated hydraulic conductivity (m/s)(Ks) are presented in Table 1. It is clear from the table that the b0.002 mm fraction has the highest SOC, CEC, FIO, and SA. On the other hand it has the lowest hydraulic conductivity (2.18 × 10−10 m/s).This value of saturated hydraulic conductivity is b 1.0 × 10− 9 m/s and therefore, matches the condition of

suitability of soils as mineral liners for disposal facility (Mitchell, 1993). This fraction (b2.0 μm) was used for all the upcoming experiments; otherwise is mentioned. The bulk density and the total porosity of LSC were found to be 1.53 g/cm3and 0.48, respectively (Table 1). The chemical compositions of the studied LSC obtained from XRF analysis are presented in Table (2). The results show that kaolinite is the sole clay mineral found in the used soil. The Fourier-transform infrared spectra of LSC is shown in Fig. 2b. The main characteristic are hydroxyl stretching bands attributed to the inner-surface hydroxyls oriented towards the interlayer at 3692 cm−1.

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

Table 1 Physicochemical and hydraulic properties of bulk AL Hammam LSC and its various fractions. Particle size (mm)

SOCa (g/kg)

CECb(mol(+)/kg)

FIOc(g/kg)

SAd(m2/g)

Porosity

Kse (m/s) × 1010

Bulk LSC 2–0.25 0.25–0.05 0.05–0.002 b0.002

50.45 48.51 45.36 46.25 55.32

40.60 41.23 39.82 43.62 45.65

10.23 11.60 9.34 11.57 20.75

20.12 28.34 30.56 35.25 38.74

0.48

5.32f 5.15 4.62 3.46 2.18

a b c d e

Soil organic carbon. Cation exchange capacity. Free ion oxide. Surface area. Saturated hydraulic conductivity, f:reads 5.32 × 10−10 m/s.

The bands at 3623 and 917 cm−1 are assigned to those hydroxyl groups oriented towards the vacant sites in the external layers of the kaolinite structure. Additional broad stretching bands of kaolinite at 3473, 1000, 917, and 510 cm−1 are attributed to associated water sorbed on the external surface (Paul and Daniel, 2013; Wada, 1965). These characteristic may attributed to the presence of kaolinite in the LSC. On the other hand, quartz exists at 700 and 800 cm−1, and calcite exists at 1637 cm− 1 (Paul and Daniel, 2013). The bands situated at lower wavenumbers, due to the vibrations of Si-O-Al and Si-O-Mg, were located at 683 and 780.5 cm−1 (Vicente-Rodriguez et al., 1996). 4.1.2. Thermal and Differential analysis. The thermal analysis of the LSC was performed in the range 50.0 °C– 1100 °C using DTA-TG analysis. Fig. 2c shows DTA/TG thermogram of LSC. The DTA peak temperatures are characteristic for each mineral and DTA curves are applicable for the identification and determination of many clays (Mackenzie, 1970; Mackenzie, 1957; Kakali et al., 2001).The LSC shows four endothermic peaks at 123.76 (396.92 K), 550.0 (823.0 K), 650.0 (923 K) and 709.71 °C (982.87 K) and one exothermic peak at 1001.6 °C(1274.76 K). The first endothermic peak represents the removal of free water from the interlayer structure (Faieta and McColm, 1993). The second peak is indicative for complete dehydroxylation and dissociation of kaolinite clay mineral (Smykatz-Kloss, 1978). The third endothermic peak is related to calcination of kaolinite to non hydrous metakaolinite. The fourth endothermic peak is related to the decomposition of Calcite (Ondruska et al., 2015). The exothermic Peak at 1001.6 °C (1274.76 K) in the DTA curve belongs to the transformation of metakaolinite into Al\\Si spinel (Mackenzie, 1970). The percentage summation of weight losses is about 28% (Fig. 2c). 4.1.3. Particle size distribution of EL Hammam LSC The particle size distribution of EL Hammam silty cly soil is illustrated in Fig. 2d. The experimental data were fitted to get the soil characteristic parameters. Since the modified (Frediund and Xing, 1994) equation allows independent control over the lower end of the curve, it was selected as the basis for the development of a grain-size distribution equation. It was used to fit the experimental data of particle size distribution:  P p ðdÞ ¼

ln

h

expð1 þ ga=d

gn igm −1

:

h

1−½ð ln ð1 þ dr =dÞ=ð ln ð1 þ dr =dm Þ

7

i

ð6Þ where Pp(d), ga, gn, gm, d, dr, dm are percent passing a particular grainsize, fitting parameter corresponding to the initial break in the grainsize curve, fitting parameter corresponding to the maximum slope of

grain-size curve, fitting parameter corresponding to the curvature of the grain-size curve, particle diameter (mm), residual particle diameter (mm) and minimum particle diameter (mm), respectively (Frediund and Xing, 1994). 4.1.4. Weathering in the source area Geochemical parameters obtained from the analysis of sediments and sedimentary rocks are widely used to infer weathering and paleoweathering conditions in source areas (Eduardo and Alberto, 2016; Ohta and Arai, 2007). The weathering indices used to examine the decomposition of unstable mineral are the chemical index of alteration (CIA), the chemical index of weathering (CIW), plagioclase index of Alteration (PIA) and mineralogical index of alteration (MIA). Table (2) illustrated the calculation of the chemical index of alteration (Schneider et al., 2016), the chemical index of weathering (Sener, 2015), the plagioclase index of alteration (Madukwe et al., 2016), the mineralogical index of alteration (Voicu et al., 1997; Babechuk et al., 2014) were defined as in Eqs. (7),(8),(9),(10). CIA ¼ Al2 O3 =ðAl2 O3 þ CaO þ Na2 O þ K 2 OÞ  100

ð7Þ

CIW ¼ Al2 O3 =ðAl2 O3 þ CaO þ Na2 O Þ  100

ð8Þ

PIA ¼ ½ðAl2 O3 −K 2 OÞ=ðAl2 O3 þ CaO þ Na2 O−K 2 OÞ  100

ð9Þ

MIA ¼ 2: ðCIA−50Þ

ð10Þ

In case of the values of the chemical index of alteration (CIA) are 100% means complete weathering of a primary material into its equivalent weathered product (Eduardo and Alberto, 2016; Voicu and Bardoux, 2002). The low CIA values of approximately 50 imply an unweathered upper crust or weak weathering, but high CIA values (i.e. 76–100) indicate moderate weathering with a complete removal of alkali and alkaline earth elements and an increase in Al2O3, which is the case of LSC; 84.57 (Su et al., 2016; Etemad et al., 2011; Dupuis et al., 2006). The CIA value is 84.57% suggesting moderate weathering either of the original source or during transport before deposition, while the CIW value is in the 88.70% as in Table 2. The ranges of mineralogical index of alteration (MIA) values indicate incipient (0–20%), weak (20–40%), moderate (40–60%), and intense to extreme (60–100%), so LSC is an intense weathered soil; 69.15% (Table 2) (Hossain et al., 2014). The degree of the chemical weathering estimated using the plagioclase index of alteration (PIA) have value of 88.12% indicating moderate weathering at the source.

Table 2 Chemical composition, chemical index of alteration (CIA), chemical index of weathering (CIW), plagioclase index of alteration (PIA) and mineralogical index of alteration (MIA) of Al Hammam LSC. Chemical composition %

SiO2

Al2O3

CaO

Fe2O3

TiO2

MgO

K2O

Na2O

Organic carbon

Loss of ignition

CIA %

PIA %

CIW %

MIA %

30.2

45.5

3.5

0.8

0.02

0.3

2.5

2.3

3.88

11

84.57

88.12

88.70

69.15

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4.2. Unsaturated hydraulic properties of LSC The movement of water at the macroscopic scale, through unsaturated porous medium is governed by Richards' equation (Richards, 1931).This equation may be derived from the laws of soil physics; Darcy's law and conservation of mass or the continuity equation. The vertical, 1D, θ -based, Richards' equation is: ∂θðZ; t Þ=∂t ¼ ∂=∂Z ððDðθÞÞ  ∂θ=∂Z Þ−dK ðθÞ=dθ  ∂θ=∂Z

ð11Þ

where θ is the volumetric water content as a function of depth and time, z is depth (positive downwards), K is hydraulic conductivity as a function of θ, and D is the soil water diffusivity as a function of θ (Warrick, 2003). The van Genuchten's soil retention relationship was used for calculating hydraulic conductivity based on saturation content (Van Genuchten, 1980): h  m i2 K ðθÞ ¼ K s S1=2 1− 1−S1=m e e

ð12Þ

and =α:λ:ðθs −θr Þ DðθÞ ¼ K s  S2þ1=λ e

ð13Þ

where α and λ are experimentally determined parameters equal 0.027 (cm−1) and 0.2 from the fitting of SWRC. Also, the residual (θr) and saturated (θs) water contents are 0.11 and 0.48, respectively estimated from fitting SWRC using RETC code (Fig. 2e) (Van Genuchten et al., 1991) Ks (m/s) is the vertical saturation hydraulic conductivity which is a function of volumetric water content (θ(h)), m = 0.19, is the van Genuchten retention parameter that is related to the uniformity of pore-size distribution and is usually specified by the soil type and Se is the effective saturation which is expressed by(Richards, 1931) as: Se ¼ θ−θr =θs −θr

ð14Þ

in which θs and θr are respectively the saturated, and residual water contents. The boundary and initial condition are: t ¼ 0:0; zN0:0 : θ ¼ θr

ð15Þ

and t ≥0:0; z→ ∞ : θ ¼ θr

ð16Þ

7

Zn(OH)−3, Zn(OH)2−4, Cd(OH)2−4, Cd(OH)−3, Pb(OH)3− appeared in pH range 9–14, 11–14, 12–14, 10.5–14, and 9.8–13.5 of predominance at pH 12.0, 14.0,13.2,14.0, and 13.2, respectively. 4.4. Removal uptake, distribution coefficient and retardation factor 4.4.1. Effect of pH and M/V on metal ion uptake Under the saturation contions, the effect of hydrogen ion concentration and m/V on the uptake of the heavy metals Pb2+ and Cd2+ or Zn2+ were studied for the pH range from 2 to 9 and from 0 to 5.5 m/V, respectively (Fig. 4). Fig. 4a shows the percentage uptake of metal ions Pd2+ Cd2+ or Zn2+ ions as a function of the pH. A significant effect of the solution pH was observed. The uptake (%) of metal ions Pd2+ Cd2+ onto LSC was increased with increasing pH from 2 to 9 (Fig. 4a). This increase in the uptake (%) of metal ions may be attributed to the positively charged Zn(OH)+, Cd(OH)+, and Pb (OH)+ species and the precipetation of Zn(OH)2, Cd(OH)2, Pb (OH)2 (Fig. 3). The effect of M/V on the percentage uptake of the studied metal ions from aqueous solutions by LSC is illustrated in Fig. 4b. This increase in the uptake (%) of metal ions is attributed to the increase in the sorbent materials and the decrease in the distances between the metal ions and the active surfaces of the sorbent. The amount of metal ions sorbed onto the LSC particles at any time, qt (mmol/g), the percentage uptake P (%) were calculated from the expressions: qt ¼ ðC o −C t Þ  V=M

ð17Þ

P ð%Þ ¼ ½ðC o −C e Þ=C o   100

ð18Þ

4.4.2. Effect of pH and V/M on the distribution coefficient, Kd Migration rates of heavy metals are significantly influenced by their distribution between LSC and leachate, which is commonly expressed in terms of the distribution coefficient; Kd values. The Kd is the ratio of the metal ion concentration adsorbed by LSC (qe) and the concentration in the leachate (Ce), given by: K d ¼ qe =C e

ð19Þ

K d ¼ ½ðC o −C e Þ=C e   V=M

ð20Þ

where: qe ¼ ðC o −C e Þ  V=M

ð21Þ

Governing Eq. (11) with relationships Eqs. (12)–(14), boundary and initial condition Eqs. (15) and (16) was solved numerically using the UNSAT-H computer code (Fayer, 2000). Details of the numerical implementation can be found elsewhere (Burden and Douglas, 2011; Zaki, 2007). Fig. 2f shows changes in the soil volumetric water content profile in the unsaturated LSC during steady water input rate = 0.31 × 10− 10 m/s (14% Ks). The wetting front reached to about 0.14, 0.19 and 0.26 cm after 6, 24 and 48 h of steady infiltration.

The Kd values of Zn2+ Pb2+ and Cd2+ at 298 K plotted as a function of pH and V/M are presented in Fig. 4. It is clear from the data trends that there is a linear relationship between the Kd and both the pH and V/M. Fig. 4a illustrates that pH is the major factor controlling Kd, however V/ M is also has a principal effect (Fig. 4b).The chemical retardation factor Rf is related to the distribution coefficient Kd by:

4.3. Speciation of zinc, cadmium, and lead

R f ¼ 1 þ ðK d  ρb =ωÞ

The relationship between the relative amounts of zinc, cadmium, and lead ionic species and solution pH was calculated at 298 K by visual MINTEQ software (Gustafsson, 2009) and presented in Fig. 3. The figure shows that these metal ions present in aqueous solution were mainly in Zn2+ (Fig. 3a), Cd2+ (Fig. 3b), Pb2+ (Fig. 3c) form up to pH 6, 7, and 5 respectively. The positively charged Zn(OH)+, Cd(OH)+, and Pb (OH)+ appeared in pH range 7–10.5, 8–12, and 5.5–11.5 and their predominant reached at pH 8.5,10.21 and 7.7, respectively. Moreover, the neutral Zn(OH)2, Cd(OH)2, Pb (OH)2 appeared and precipitated in pH range 7.5–13, 9–14, and 7.8–13 of maximum at pH 9.8,11.6 and 10.85, respectively. On the other hand the negatively charged species

where Ct is the time interval concentration (mg/l) of metal ion in solution, Kd (ml/g) is the distribution coefficient, ρb is the bulk density (g/cm3) and ω is the porosity of LSC at saturation zone (groundwater). It is observed that the percentage uptake of metal ions Pd2 + Cd2 + or Zn2 + from lechate solutions increases rapidly up to 4 M/V (g/ml). The percentage uptake at equlibrium onto LSC takes the order Zn2+ N Pd2+ N Cd2+ with the percentage uptake 94.8, 92.7 and 86 respectively. On the other hand the Kd and Rf values increased with increase in the temperature with the same order of metal ions (Table 3). The sorption retardation of LSC against the mobility of used ion metals is maximum for Zn2+ and minimum for Cd2+ (Table 3).

ð22Þ

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8

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100

(a)

2Zn(OH)4

2+

Zn

% Zn(II) species

75

Zn(OH)2(aq)

50 -

Zn(OH)3 25 +

Zn(OH) 0

4

6

8

10

12

14

pH

100

(b) 2+ Cd

2-

Cd(OH)4

% Cd(II) species

75 Cd(OH)2(aq)

50

-

Cd(OH)3

4.5. Sorption kinetic modeling under saturation conditions

+

CdOH

25

0

6

8

10 pH

12

14

100 (c)

2+

Pb3(OH)4

2+

3-

Pb(OH)

Pb 75 % Pb(II) species

4.4.3. Sorption preference of Zn2+, Pb2+ and Cd2+ onto LSC The sorption preference exhibited by soils for one element over others may be due to the hydrolysis constant, the atomic weight, the ionic radius, and subsequently hydrated radius, and its Misono softness value. In this respect, (Shaheen, 2009) studied the sorption of Cd2+ and Pb2+ in different soils and reported that Kd values of Pb2+ were obviously higher than those of Cd2 +, indicating that Pb2 + was retained more strongly by the soils than Cd2+. This is usually attributed to differences in element characteristics and the resulting affinity for sorption sites (Appel and Ma, 2002; Appel et al., 2008; McBride, 1994). On the other hand, Zn2 + is adsorbed to a greater extent than Cd2 + onto ALHammam LSC, this agrees with the results of (Schwertmann and Taylor, 1989), and observations by (Gomes et al., 2001) for the retention of these two trace elements in Brazilian soils and (Shaheen et al., 2012), indicated that the Kd values of Zn2+ were higher than those of Cd2+ in the studied soils, indicating that Zn2+ was more strongly retained by the soils than Cd2+ (Figs. 4 and 5). This order indicates the higher tendency of Cd2+ to remain in solution compared to Zn2+. This is usually attributed to differences in element characteristics and the resulting affinity for sorption sites which can be summarized as follows: ionic radii Cd2+ (0.97 Å) N Zn2+ (0.74 Å); atomic weight Cd2+ (112.41) N Zn2+ (65.38) (Wilman et al., 2015); electronegativity Cd (1.7) N Zn (1.6) (Gomes et al., 2001) hydrolysis constant Cd (9.0) N Zn2+ (10.1) (Baes and Mesmer, 1976) and softness Cd2 + (3.04) N Zn2 + (2.34) (Misono et al., 1967). The retention of Zn2+ is higher than Pb2+ could be attributed to ionic radius of Zn (0.74 Å) b that of Pb2+ (1.2 Å), the electronegativity of Zn2 + (1.6) b that of Pb2 + (1.8). On the other hand the strongest bond should be formed by the metal with the greatest charge-to-radius ratio, which increases in the order of Zn2 + N Pb2 + (Péter et al., 2008) (Figs.4 and 5).

Pb(OH)2(aq) 50

4.5.2. The pseudo-second-order rate model The pseudo-second-order rate model (McKay and Ho, 1999a; McKay and Ho, 1999b; Weber and Morris, 1963) is expressed as:

4+

3+

Pb2OH

4.5.1. Effect of time on the uptake zinc, cadmium, and lead Preliminary investigations on the rate of uptake of Zn2+,Cd2+, and Pb2+ ions onto LSC indicated that the processes are quite rapid and typically 60–70% of the ultimate sorption of each ion occurs within the first 10 min of contact. The initial rapid sorption subsequently gives way to a slow approach to equilibrium, and saturation is reached in about 30 to 40 min. The amounts of metal ions sorbed after each interval time, for a fixed concentration of 100 mg/l and at different studied temperatures, are plotted in Fig. 5. The data showed that the amount of Zn2+, Cd2+ and Pb2+ sorbed at equilibrium increases with the increase in temperature indicating an endothermic nature of the process and the time required to reach saturation remained practically unaffected. Fig. 5d, e and f show the comparison of the amounts of metal ions sorbed onto LSC after each interval time at 298, 313 and 333 ± 1 K. It was found that the removal sorption affinity took the sequence of Zn2+ N Pb2+ N Cd2+ at all different temperatures.

Pb4(OH)4

+

PbOH 25

t=qt ¼ 1=k2 q2e þ t=qe

ð23Þ

or

0

4

6

8

10

12

14

t=qt ¼ 1=h þ t=qe

ð24Þ

pH Fig. 3. Percentage species of (a) zinc, (b) cadmium, and (c) lead in water at 298 K.

where k2 (g · mg−1 · hr−1) is the rate constant of second-order adsorption h is k2qe. The kinetic plots of t/qt vs. t for Zn2+, Cd2+, Pb2+ removal at different temperatures are presented in Fig. 6a–c. The relationship is

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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9

(a)

Zn2+ Pb2+ Cd2+

600 2+

Zn 2+ Pb 2+ Cd

525

(d)

Kd (ml/g)

450 375 300 225 150 Zn2+ Pb2+ Cd2+

75 0

0

50

100

150 200 V/M (ml/g)

250

300

Fig. 4. Percentage uptake of Zn2+ Pb2+ and Cd2+ as a function of (a) pH, (b) M/V and Kd values plotted as a function of (c) pH and (d) V/M at 298 K.

linear, and the correlation coefficient R2 and SD suggest a strong relationship between the parameters and also explains that the process of adsorption of each ion follows pseudo-second-order kinetics (Table 4). As illustrated in Table 4, the values of the initial adsorption rate h and rate constant k2 are increased with increase in temperature. The R2 has an extremely high value (N 0.99), and its calculated equilibrium adsorption capacity, qe, is consistent with the experimental data. These results suggest that the Pseudo-second order adsorption mechanism is predominant and that the overall rate constant of each ion adsorption process appears to be controlled by the chemisorption process (McKay and Ho, 1999b; Weber and Morris, 1963).

4.5.3. Morris and Weber model Other simplified models are also used to test the sorption mechanism. It is also known that at an intensive stirring of the sorptive system, the intraparticle diffusion of the solute sorbed from the solution into the sorbent pores could be a limiting step. Morris and Weber model (Weber and Morris, 1963) and that suggested by (Helfferich, 1962) were also used. The Morris–Weber equation is written as:

qt ¼ K ad ðt Þ1=2

ð25Þ

Table 3 Retardation factor (Rf), distribution coefficient (Kd), and percentage uptake (%P) of Zn2+, Cd2+, and Pb2+ ions from aqueous solutions onto AL Hammam LSC, at different temperatures and at pH 5.0. Temperature (K)

2+

298 313 323 333

Kd (ml/g)

P (%) 2+

2+

Zn

Cd

Pb

67.0 69.7 72.5 73.9

53.5 55.0 57.0 58.5

65.0 67.2 69.9 70.0

2+

Rf 2+

2+

Zn

Cd

Pb

428.2 460.3 526.7 568.0

220.5 244.4 265.1 281.9

333.9 409.8 464.2 466.7

Zn2+

Cd2+

Pb2+

950.98 1080.96 1236.70 1333.64

540.86 564.41 622.98 662.39

872.38 962.47 1090.11 1095.97

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

Fig. 5. Effect of contact time on the amount sorbed of (a) Zn2+, (b) Cd2+, (c) Pb2+ and the sorption order at (d) 298 ± 1 K, (e) 313 ± 1 K, and (f) 333 ± 1 K onto AL Hammam LSC at pH 5.0.

where, K ad is the rate constant of intraparticle transport (mmol · g−1 · min−1/2). According to this model, plotting a graphic of qt versus t1/2, if a straight line passing through the origin is obtained, it can be assumed that the involved mechanism is a diffusion of the species. In this case the slope of the linear plot is the rate constant of intraparticle transport. As can be seen in Fig. 6d–f, for times up to 12.25 min, the Morris-Weber relationship holds good and the values of Kad were calculated, from the slope of the linear plots obtained, and presented in Table 5. The values of SD for pseudo- second- order rate and Morris-Weber kinetic models are given in Table 4 and Table 5. Pseudo- second- order kinetic rate model can well predict the adsorption kinetics of Zn2+,Cd2+, and Pb2+ onto LSC with high correlation coefficients and

low SD. In all cases, Pseudo- second- order kinetic rate model represents the best fit of experimental data than the Morris-Weber kinetic models. Moreover, based on SD values, the error in the case of pseudo- secondorder is about 10% that of Morris-Weber kinetic model (Table 4 and Table 5).The comparison of the experimental and the values of qt obtained from the two models is shown in Fig. 6.

4.5.4. Homogeneous particle diffusion model (HPDM) The rate-determining step of sorption in this model, is normally described by either ions diffusion through the liquid film surrounding the particle, called film diffusion, or diffusion of ions into the sorbent

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

4.0

11

(a)

3.5

(d)

t/qt (min.g/mg)

3.0 2.5 2.0 1.5

298 K 313 K 323 K 333 K

1.0 0.5 0

0 4 8 12 16 20 24 28 32 36 40 44 Time (min)

5.5

(b)

5.0

(e)

4.5

t/qt(min.g/mg)

4.0 3.5 3.0 2.5 2.0 1.5

298 K 313 K 323 K 333 K

1.0 0.5 0

0

4

8 12 16 20 24 28 32 36 40 44 Time( min)

3.5

(c) (f)

3.0

t/qt(min.g/mg)

2.5 2.0 1.5 1.0

298 K 333 K 313 K 323 K

0.5 0

0

4

8 12 16 20 24 28 32 36 40 44 Time (min)

Fig. 6. Pseudo-second order kinetic plots for the sorption of (a) Zn2+, (b) Cd2+, and (c) Pb2+ and Morris–Weber kinetic plots for the sorption of (d) Zn2+, (e) Cd2+, and (f) Pb2+ ions from aqueous solutions onto AL Hammam LSC at pH 5.0 and at different temperatures.

particles, called particle diffusion mechanism. Nernst-Plank equation (Helfferich, 1962), which takes into account both concentration and electrical gradients of exchanging ions into the flux equation, was used to establish the HPDM equations. If the diffusion of ions from the solution to the sorbent particles is the slowest step, rate-determining step, the liquid film diffusion model controls the rate of sorption. So, the following equation can be used to calculate the coefficient of diffusion: − ln ð1−X Þ ¼ ð3  D  C e =r o  δ  C r Þ  t

ð26Þ

in which Cr and Ce are the equilibrium concentrations of the ion in solid and solution phases, respectively, X is the equilibrium fraction attainment, D is the diffusion coefficient in the liquid phase, ro is the radius of the adsorbent particle; δ is the thickness of the liquid film. Depending on the slowest step, the adsorption mechanism can determined; film or particle diffusion. The particle diffusion model could be applied to calculate the diffusion coefficients and the rate formula becomes:    − ln 1−X 2 ¼ 2  Dr  π2 =r 2o  t

ð27Þ

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

Table 4 The calculated parameters of pseudo second-order sorption kinetic models for Zn2+, Cd2+, and Pb2+ ions from aqueous solutions onto LSC, at different temperatures and at pH 5.0. Temp. (K)

298 313 323 333

h × 10−3 (mg/g · hr)

qe (mg/g) Pb2+

Zn2+

Cd2+

Pb2+

Zn2+

Cd2+

Pb2+

Zn2+

Cd2+

Pb2+

Zn2+

Cd2+

Pb2+

Zn2+

8.47 8.69 9.01 9.26

13.53 13.97 14.23 14.73

13.74 14.43 14.88 15.15

0.40 0.39 0.37 0.11

0.60 0.64 0.74 0.74

0.70 0.76 1.06 1.2

5.440 5.164 4.510 1.27

3.266 3.281 3.636 3.411

3.70 3.65 4.65 5.23

0.991 0.995 0.999 0.994

0.992 0.993 0.992 0.997

0.991 0.998 0.995 0.999

0.09 0.05 0.08 0.04

0.02 0.03 0.08 0.07

0.01 0.08 0.09 0.09

4.6. Sorption mechanism under saturation conditions To understand the sorption mechanism(s) for Zn2+, Cd2+, and Pb2+ onto LSC, all cations in the leachate solutions were measured. Fig. 8a shows the concentrations (meq/l) of Zn2+, Cd2+, and Pb2+ and their sum, in addition to their sum with the released cation metals Ca2 +, Mg2 +, Na+, and K+ from LSC in solution at contact time (min). The Zn2+, Cd2+, and Pb2+ concentration in leachate Fig. 8a shows the concentrations (meq/l) of Zn2+, Cd2+, and Pb2+ and their sum, in addition gradually decreased with increasing contact time, indicating their removal through sorption. On the other hand, Ca2 + + Mg2 ++ Na++ K+ concentration increased progressively due to the release of these cations from weakly bound sites on LSC. It is also noteworthy that the concentration of total cations in solution was constant throughout the contact time. The amount of Ca2+, Mg2+, Na+, K+ added was equivalent to that of the removed Zn2+, Cd2+, and Pb2+. The concentrations (meq/l) of the sorbed heavy metals versus the sum of concentrations of released cations Ca2+, Mg2+, Na+, and K+ within all the kinetic experiments is shown in Fig. 8b. It is clear from the figure that the 1:1

Table 5 The calculated parameters of Morris Weber sorption kinetic models for Zn2+, Cd2+, and Pb2+ ions from leachate solutions onto AL Hammam LSC, at different temperatures and at pH 5.0.

298 313 323 333

DS

Cd2+

where, Dr. is the particle diffusion coefficient. The two previous model equations (Eqs. (26), (27)) were tested against the kinetic rate data of Zn2 +, Cd2 + and Pb2 + ions sorbed onto LSC by plotting the functions of − ln (1 − X2)against contact time. The straight lines of the plots of − ln (1 − X2) versus contact time, as shown in Fig. 7a–c, pass through the origin for all metal ions indicating that the particle diffusion model control the sorption processes at all studied temperatures. The slope values of these plots were used to calculate the effective diffusion coefficients (Dr) as shown in Fig. 7d (Yakout and Hassan, 2014) using Eq. (27). These calculated values together with the R2 for the studied metal ions are presented in Table 6. The magnitude of the diffusion coefficient is dependent upon the nature of the sorption process. For physical adsorption, the value of the effective diffusion coefficient ranges from 10−6 to 10−9 m2/s and for chemisorption, the value ranges from 10−9 to 10−17 m2/s (Helfferich, 1962). The difference in the values is due to the fact that in physical adsorption the molecules are weakly bound and therefore there is ease of migration, whereas for chemisorption the molecules are strongly bound and mostly localized. Therefore, from this work, the most likely nature of sorption is chemisorption since the values of Dr. were in the order 10−17 m2/s for all metal ions. This is in agreement with the pseudo second - order kinetic model (Table 6).

Temperature (K)

R2

k2 (g/mg · hr)

Kad (mg/g · min1/2)

R2

Cd2+

Pb2+

Zn2+

Cd2+

Pb2+

Zn2+

Cd2+

Pb2+

Zn2+

2.73 3.74 3.21 4.28

4.24 4.86 5.35 5.85

4.67 4.93 5.22 5.50

0.969 0.976 0.950 0.975

0.956 0.965 0.982 0.985

0.978 0.989 0.987 0.985

0.15 0.38 0.55 0.28

0.36 0.65 0.52 0.48

0.58 0.49 0.37 0.75

DS

linear trend indicates that there is a strong correlation between the sorbed Zn2+,Cd2+ and Pb2+ cation metals onto LSC and the released Ca2+, Mg2+, Na+, and K+ cation metals from it, this likely dominated by cation exchange process could be expressed by Arshadi et al. (2014): BMe þ HMðsÞ2þ ⇄BHMðsÞ þ Með2Þþ Xh

i X h i Δ HM ðsÞð2Þþ Með2Þþ ¼

ð28Þ ð29Þ

where Me and HM(s) represent weakly bound major cations Ca2 +, Mg2+, Na+, and K+ and sorbate metals Zn2+,Cd2+ and Pb2+, respectively, and the squared brackets denote the concentration (meq/l) in the solution. This agrees with the results of kinetic results of this study; second order chemical reaction. However, for each metal, the difference between the adsorption isotherm and the cations desorption isotherm could be related to the extent of a complexation/chelation mechanism completing the ion exchange one, where the extent of these mechanisms depended on the nature of the metal (Arshadi et al., 2014). 4.7. Effect of temperature There are lots of parameters affecting the sorption rates like structural properties of the sorbent, metal ion properties, initial concentration of metal ions, pH, temperature, or presence of competing ions (Venkatrajan et al., 2016; Byungryul et al., 2013). It is known that sorption kinetics is dependent or controlled by different kind of mechanisms like mass transfer, diffusion control, chemical reactions, and particle diffusion. The values of Dr calculated at different studied temperatures for the studied metal ions are presented in Table 5. Plotting of ln Dr. versus 1/T gave a straight line, as shown in Fig. 8c, proves the validation of Arrhenius equation. Dr ¼ D0 expð−Ea =R  T Þ

ð30Þ

The energies of activation for Zn2+, Cd2+ and Pb2+, Ea, were calculated from the slope of the straight lines in Fig. 7d and the obtained values were presented in Table 7. Such a low value of the activation energies for the studied metal ions suggests that the temperature has a limited effect and the sorption process is controlled by particle diffusion mechanism. The Arrhenius equation would be also used to calculate D0 which in turn is used for the calculation of entropy change, ΔSo of the sorption process using:    2 D0 ¼ 2:72 d KT=h exp ΔSo =R

ð31Þ

where, h is the Plank's constant, K is the Boltzmann's constant, T is the absolute temperature, d is the distance between two adjacent active site in the solid matrix, and R is the ideal gas constant (8.31 × 10−3 kJ/K/mol). It was assumed that the value of d is 5 × 10−8 cm (Nirmalya et al., 2014; Petrou and Eliopoulos, 2009), the values of entropy, ΔSo the enthalpy, ΔHo and the Gibbs free energy of activation, ΔGo for Zn2+, Cd2+ and Pb2+ were calculated and presented in

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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13

3.0

(a)

2.7

-37.5

2.4

333 K 323 K 313 K 298 K

-1 2

1.8

(d)

-38.0 ln(Dr), m .s

2

-ln(1-x )

2.1

1.5 1.2

-38.5

-39.0

0.9 0.6

2+

Pb 2+ Zn 2+ Cd

-39.5

0.3 -40.0

0

0

2

4

6

8

10

12

14

16

0.0030

0.0031

Time (min)

2.1

0.0032 1/T, K

0.0033

0.0034

-1

(b)

1.8 333 K 323 K 313 K 298 K

2

-ln(1-x )

1.5 1.2 0.9 0.6 0.3 0

0

2

4

6

8

10

12

14

16

Time(min) 3.9

(c)

3.6 3.3

333 K 323 K 313 K 298 K

3.0 2.4

2

-ln (1-x )

2.7 2.1 1.8 1.5 1.2 0.9 0.6 0.3 0

0

2

4

6

8

10

12

14

16

Time (min)

Fig. 7. Plots of –ln(1 − χ2) as a function of time for the diffusion of (a) Zn2+, (b) Cd2+, and (c) Pb2+ ions from aqueous solutions onto AL Hammam LSC and (d) diffusion coefficient, Dr., for the sorption of Zn2+, Cd2+ and Pb2+ ions from lechates at pH 5.0 and at different temperatures.

Table 6 Diffusion coefficients and reaction activation energy calculated based on HPDM for Zn2+, Cd2+, and Pb2+ ions sorbed from leachate solutions onto AL Hammam LSC, at different temperatures and at pH 5.0. Element

Temperature (K)

Dr (m2/s)

R2

Do (m2/s)

Ea (kJ/mol · K)

Cd2+

298 313 323 333 298 313 323 333 298 313 323 333

3.32 × 10−17 3.91 × 10−17 4.20 × 10−17 4.58 × 10−17 3.59 × 10−17 3.91 × 10−17 4.30 × 10−17 4.68 × 10−17 4.51 × 10−17 5.11 × 10−17 5.99 × 10−17 6.80 × 10−17

0.998 0.995 0.984 0.987 0.986 0.986 0.991 0.979 0.988 0.992 0.983 0.986

6.97 × 10−16

7.5

4.23 × 10−16

6.25

2.31 × 10−16

9.8

Pb2+

Zn2+

Table 6. The Gibbs free energy of activation ΔGo was calculated from the well-known equation (Venkatrajan et al., 2016; Byungryul et al., 2013):

ΔGo ¼ ΔH o −T ΔSo ¼ Ea −RT−T ΔSo

ð32Þ

The values of enthalpy change (ΔHo) and entropy change (ΔSo) calculated from the slope and intercept of the plot of ΔGo versus T (Fig. 8c) are also given in Table 7. The change in ΔHo for Zn2+, Cd2+ and Pb2+ ion metals was found to be positive confirming the endothermic nature of the sorption process. So, with the increase in the temperature of a MSW barrier system as a result of biological and chemical reactions, this could lead to retardation and immobilization of heavy metals into LSC.

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

Concentration of cation metals in solution, (meq/l)

14

(a)

2.4

2+ 2+ 2+ 2+ 2+ 2+ + ∑Zn +Cd +Pb +Mg +Ca +Na +K

(b)

2.0 2+

1.6 ∑

1.2

2+

2+

Cd 2+ Zn 2+ Pb

2+

Zn +Cd +Pb

0.8 0.4 0

2

6

10

14

18 22 26 Time ,(min)

30

34

38

42

-1.2

(c) -1.5

Δ G (kJ/mol)

-1.8 -2.1 -2.4 2+

Cd 2+ pb 2+ Zn

-2.7 -3.0 290

300

310

320

330

340

T(K) Fig. 8. Concentration of cations in solution versus time and (a) adsorption concentration of Zn2+, Cd2+, and Pb2+ ions onto AL Hammam LSC and (b) the Ca2+, Mg2+, Na+, and K+ ions released in solution based on the sorption kinetics data and (c) Gibbs free energy change as a function of sorption temperature of Zn2+, Cd2+, and Pb2+ ions from aqueous solutions onto Al Hammam LSC, at pH 5.0.

The Gibbs free energy change, ΔGo, is the fundamental criterion of spontaneity. Reactions occur spontaneously at a given temperature if ΔGo is a negative quantity. The ΔGo of the reaction is given by: ΔGo ¼ −RT lnK c

ð33Þ

K c ¼ F e =ð1− F e Þ

ð34Þ

where, Kc is the sorption equilibrium constant and Fe is the fraction attainment of metal ion sorbed at equilibrium. These parameters are important to understand the adsorption mechanism. The variation of Kc with temperature, as summarized in Table 7, showed that Kc values

increase with increase in sorption temperature, thus implying a strengthening of adsorbate–adsorbent interactions at higher temperature. It is clear to see that the values of Gibbs free energy change (ΔGo) at all tested temperatures were negative. This confirms that the sorption extraction process is spontaneous and thermodynamically favorable. The positive values of ΔHo confirms the endothermic nature of sorption extraction process, which is supported by the increase in the adsorption capacity of LSC (Table 7). Furthermore, the positive value of ΔSo shows the freedom of adsorbate heavy metal ions during the adsorption extraction and the high degree of randomness at solid-liquid interface during the adsorption removal of Zn2 +, Pb2 +, and Cd2 + by LSC from waste solutions (Ali et al., 2016; Elvan and Atun, 2006).

Table 7 The thermodynamic parameters of Zn2+, Cd2+, and Pb2+ ions sorbed from leachate solutions onto AL Hammam LSC at different temperatures and at pH 5.0. Temp. K

ΔGo (kJ/mol) 2+

298 313 323 333

ΔHo (kJ/mol) 2+

Zn

Cd

−1.60 −2.17 −2.61 −2.89

−1.32 −1.52 −1.71 −1.85

2+

Pb

−1.54 −1.87 −2.14 −2.35

2+

ΔSo (J/mol · K) 2+

2+

Zn

Cd

Pb

9.17

3.1

5.39

2+

Kc = Fe/(1 − Fe) 2+

2+

Zn

Cd

Pb

40.0

15.0

23.0

Zn2+

Cd2+

Pb2+

1.94 2.33 2.63 2.84

1.15 1.23 1.35 1.42

1.87 2.05 2.22 2.33

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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15

4.8. Sorption and desorption transport of Zn2+, Pb2+, and Cd 2+ through leachate saturated LSC

is one of the most general and widely used models in column performance theory (Cruz et al., 2013), it is represented by:

Sorption and desorption transport of Zn2+, Pb2+, and Cd2+ in leachate saturated LSC was investigated using experimental fixed bed column, where the breakthrough curves can be predicted. Thomas model

C ðt Þ=C o ¼ ð1 þ exp½ðK Th ðqo  M=Q −C o  t ÞÞ−1

1.0

(a)

(b)

5.0

2+

2+

Zn 2+ Pb 2+ Cd

2.5

Cd 2+ Pb 2+ Zn

Sorption

0.8

0.6 C/Co

ln(Co/C-1)

ð35Þ

0

Desorption

0.4 -2.5

0.2 -5.0 40

60

80

100 120 140 Time (hr)

160

0

180

0

50

100 150 200 250 300 350 400 Time (hr)

1.0

(c)

0.8

0.6

2+

C/Co

Cd onto 50-200μm LSC 2+ Cd onto total range LSC 2+ Pb onto 50-200μm LSC 2+ Pb onto total range LSC 2+ Zn onto 50-200 μm LSC 2+ Zn onto total range LSC

0.4

0.2

0

0

50

100 Time (hr)

150

200

1.0 2+

Cd 2+ Pb 2+ Zn

0.8

0.6 C/Co

(d)

0.4 CXTFIT Fitting

0.2

0

0

20

40

60

80 100 120 140 160 180 Time (hr)

Fig. 9. (a) Experimental and linear Thomas model fitting of the onto the 50–200 μm fraction (b) Thomas sorption breakthrough curve fitting and desorption of Zn2+, Cd2+, and Pb2+ from the 50–200 μm fraction of AL Hammam LSC using malic acid (c) comparison of dynamic sorption onto total range of particle size and onto the 50–200 μm fraction at pH 5.0 and at 289 K (d) experimental and predicted breakthrough curves by the CXTFIT code for sorption of Zn2+, Cd2+, and Pb2+ onto the 50–200 μm fraction of AL Hammam LSC at pH 5.0 and at 289 K.

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

Table 8 Thomas model fitting transport parameters of Zn2+, Cd2+, and Pb2+ ions breakthrough curves in leachate saturated AL Hammam LSC column at pH 5.0 and 298 K ± 1. Metal ion

Kth × 102 (ml/mg · min)

qo (mg/g)

R2

MSE × 103

Zn2+ Cd2+ Pb2+

1.25 1.08 1.23

27.50 20.55 23.83

0.985 0.987 0.988

5.32 3.49 4.28

where, C(mg · l− 1), KTh (cm3 · min−1 · mg− 1), qo (mg/g), Q (cm3 · min−1) and t (min) are the effluent metal ion concentration at time t, Thomas rate constant, the adsorption capacity of the bed, volumetric flow rate and the time, respectively. The kinetic coefficient KTh and qo can be determined from a plot of ln [(Co/C) − 1] against t at a given condition (Fig. 9a). ln ðC o =C ðt Þ−1Þ ¼ K Th ðqo  M=Q−C o  t Þ

ð36Þ

It is clear from Fig. 9 that the model gave a good fit to the experimental data at initial metal ion concentration. The sorption experimental breakthrough curves were very close to calculated breakthrough curves according to Thomas model. Based on the values of R2 and maximum deviation, that the qo values calculated from the Thomas model is close to the qo, exp. values obtained experimentally (Table 8). Malic acid (MA), a dicarboxylic acid, of pKA1 and pKA2 3.4 and 5.2, respectively was used in desorption of Zn2 +, Pb2 +, and Cd2 + ions from LSC at pH 5.0 ± 0.1, at 298 ± 1 K, and at 0.1 cm3/min flow rate (Fig. 9b). It is clear from the data that the desorption of Cd2+ from LSC is faster than that of both Pb2+ and Zn2+. This may be attributed to different sorption mechanisms of Zn2+, Pb2+, and Cd2+ onto both LSC and MA. 4.9. Effect of the particle size fraction b2 μm on the sorption behavior of LSC Adsorption loading amounts of Zn2 +, Cd2 + and Pb2 + onto the total range of particle size range of LSC and the 50–200 μm fraction of LSC from 10− 4 M NaCl containing the heavy metal at tracer level (100 mg/l for each metal ion, were calculated from an adsorbate mass balance as follows (Wilman et al., 2015): Zt qðt Þ ¼ Co  Q =M

ð1−C ðt Þ=CoÞ  dt

ð37Þ

0

The percentage difference, Δq%, due to sorption onto the total range of particle size including the fraction b2 μm and that sorbed on the 50–200 μm fraction of the LSC was calculated according to the Eq. (38): 0t 1 Z Zt @ Δq% ¼ ð1−C ðt Þ=CoÞTot  dt− ð1−C ðt Þ=CoÞ50−200μm  dtA 0

ð38Þ

0

Zt = ð1−C ðt Þ=CoÞTot  dt  100 0

Table 9 Estimated transport parameters using CXTFIT for Zn2+, Cd2+, and Pb2+ ions breakthrough curves in saturated leachate LSC column at pH 5.0 and 298 K. Ion metal

vp (cm/h) × 102

D (cm2/h) × 102

P

λ

R2

MSE × 103

Zn2+ Cd2+ Pb2+

4.45 6.11 5.11

1.0 1.0 1.0

22.25 30.54 25.55

0.22 0.16 0.20

0.981 0.991 0.983

4.64 1.37 2.49

where ∫ (1 − C(t)/Co)Tot is the sorption of Zn2+, Cd2+ and Pb2+ onto the total range of particle size range of LSC and ∫ (1 − C(t)/Co)50–200 μm is the sorption on the 50–200 μm fraction of LSC. Fig. 9c shows the breakthrough curves of the dynamic sorption of Zn2+, Cd2+ and Pb2+ onto the total range of particle size range of LSC and the 50–200 μm fraction of LSC at room temperature. The Δq% for Cd2+, Pb2+ and Zn2+ are 7.4, 6.3 and 2%, respectively. It is clear from Fig. 9c that the area over the curves (sorption onto the Al Hammam LSC) for the total range of LSC particle size are larger than that of the fraction 50–200 μm. This could be attributed to the increase of the surface area due to the presence of the fraction b 2 μm fraction of high surface area. 4.10. Hydrodynamic transport properties of Zn2+, Pb2+, and Cd2+ The experimental transport data of Zn2+, Pb2+, and Cd2+ from the LSC column are modeled with CXTFIT software (Tang et al., 2010). Assuming equilibrium, water saturated, steady state convection dispersion equation (CDE), no degradation of the solute in the waste water, and no production of the solute, the parameters are estimated using within the simulation domain is given by the following equation (Freeze and Cherry, 1979): 2

R f  ∂C=∂t ¼ D ∂ C=∂x2 −νp ∂C=∂x−λC

ð39Þ

where Rf is retardation factor, C is the concentration, t is time, x is distance from the inlet, νp average particle velocity(cm/h), D dispersion coefficient (cm2/h) and λ the dispersivity (D/vp). where θ and ρb are the porosity and bulk density of LSC, respectively. The experimental data set from a 5 cm column length (h) give full transport profiles for concentration profile at 100 mg/l. The experimental breakthrough curves for different ion metals are largely coinciding with the corresponding curves fitted by CXTFIT model, giving goodness-of-fit values R2 N 0.99 (Fig. 9d, Table 9). In addition to Pêclet number (P) (vp · h/D), the fitted transport parameters D, νp, and λ, within mean square error (MSE) (b 0.2 × 10−3), are reported in Table 9. The calculated hydraulic Pêclet number values of Zn2+, Cd2+ and Pb2+ in leachate were 22.25, 30.54 and 25.55, respectively (Table 9). This high Pêclet number (b 32) indicates dominance of dispersion over advection and vice versa (Freeze and Cherry, 1979). The fitted transport parameters D, νp, and λ are for each depth were almost identical (Table 9). These results show that the CDE is an appropriate model for describing transport of Zn2+, Pb2+, and Cd2+ in saturated porous LSC media. The fitted and derived transport parameters within R2 N 0.98 and MSE b 4.7 × 10–3 are summarized in Table 9. 5. Conclusions The results showed that the percentage equilibrium uptake of the metal ions by LSC are 94.8, 9.27, and 86.0 for Zn2+, Pb2+, and Cd2+, respectively. The predominant kinetic model was found to be the Pseudosecond order and the overall rate constant of each metal ion adsorption is controlled by the chemisorption process. The diffusion coefficient value was found in the range (3.32–6.8)0.10−17 m2/s and increases with increase in the temperature. The distribution coefficient for Pb2+, Cd2 +, and Zn2+ ranged from 404.9 to 568, 230.1 to 281.9, and 371.4 to 466.7 ml/g in the temperature range 298 to 333 ± 1 K, respectively. The value of retardation factor and the sorption affinity took the order Zn2+ N Pb2+ N Cd2. The experimental investigation on ionic concentrations in sorption batches suggested that sorption behaviors of Zn2 +, Pb2+, and Cd2+ ion metals onto LSC are mainly controlled by cation exchange. The wetting front of water movement in the LSC unsaturated zone reached to about 0.06, 0.19 and 0.25 cm after 6, 24 and 48 h of steady infiltration. The saturation hydraulic conductivity of the LSC (b 2.0 μm) fraction is b1.0 × 10−9 m/s (2.18 × 10−10 m/s) therefore, it matches the condition of suitability of soils as mineral liners for a landfill facility. The Pêclet number values (b32) indicates dominance of

Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

A.A. Zaki et al. / Applied Clay Science xxx (2016) xxx–xxx

dispersion over advection. Based on the results of this work, it is recommended to use the b2.0 μm fraction of LSC as a sorptive and hydraulic barrier to attenuate Zn2 +, Cd2 + and Pb2 + ions presented in landfill leachate to protect the shallow groundwater beneath AL Hammam landfill facility. Acknowledgement The authors acknowledge the journal anonymous reviewers, for their very constructive and helpful comments as well as for editorial comments, which helped to improve the manuscript. The authors are grateful to the HLWMC of Atomic Energy Authority of Egypt for the financial and academic support throughout this work. References Abdel Aziz, H.M., 2005. Sorption equilibria of lead (II) on some Palestinian soils the silty ion exchangers. Colloids Surf. A Physicochem. Eng. Asp. 264, 1–5. Abu-Zeid, K., El Arabi, N., Fekry, A., Abdel Meneum, M., Taha, K., El Karamany, S., 2009. Assessment of Groundwater Potential in Alexandria Governorate. Cedar, Egypt. 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Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016

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Please cite this article as: Zaki, A.A., et al., Sorption characteristics of a landfill clay soil as a retardation barrier of some heavy metals, Appl. Clay Sci. (2016), http://dx.doi.org/10.1016/j.clay.2016.09.016