UvA-DARE (Digital Academic Repository) Characterization of ... · Refaey Mohammed, Y. B. (2016). Characterization of clay deposits from Egypt and assessment of their potential application

UvA-DARE is a service provided by the library of the University of Amsterdam (http://dare.uva.nl)

UvA-DARE (Digital Academic Repository)

Characterization of clay deposits from Egypt and assessment of their potential application forwaste water treatment: How dissolved organic matter determines the interaction of heavymetals and clay minerals

Refaey Mohammed, Y.B.

Link to publication

Creative Commons License (see https://creativecommons.org/use-remix/cc-licenses):Other

Citation for published version (APA):Refaey Mohammed, Y. B. (2016). Characterization of clay deposits from Egypt and assessment of their potentialapplication for waste water treatment: How dissolved organic matter determines the interaction of heavy metalsand clay minerals.

General rightsIt is not permitted to download or to forward/distribute the text or part of it without the consent of the author(s) and/or copyright holder(s),other than for strictly personal, individual use, unless the work is under an open content license (like Creative Commons).

Disclaimer/Complaints regulationsIf you believe that digital publication of certain material infringes any of your rights or (privacy) interests, please let the Library know, statingyour reasons. In case of a legitimate complaint, the Library will make the material inaccessible and/or remove it from the website. Please Askthe Library: https://uba.uva.nl/en/contact, or a letter to: Library of the University of Amsterdam, Secretariat, Singel 425, 1012 WP Amsterdam,The Netherlands. You will be contacted as soon as possible.

Download date: 15 Sep 2020

Characterization of clay deposits from Egypt and assessment

of their potential application for waste water treatment

How dissolved organic matter determines the interaction

of heavy metals and clay minerals

Yasser Refaey

Characterization of clay deposits from Egypt and assessment of their

potential application for waste water treatment

How dissolved organic matter determines the interaction of heavy metals and

clay minerals

Yasser Refaey

Characterization of clay deposits from Egypt and assessment of their

potential application for waste water treatment

How dissolved organic matter determines the interaction of heavy metals and

clay minerals

ACADEMISCH PROEFSCHRIFT

Ter verkrijging van de graad van doctor

aan de Universiteit van Amsterdam

op gezag van de Rector Magnificus

Prof. dr. ir. K.I.J. Maex

Ten overstaan van een door het College voor Promoties ingestelde

commissie, in het openbaar te verdedigen in de Agnietenkapel

op donderdag 22 december 2016, te 16:00 uur

door

Yasser Baeoumy Refaey Mohammed

geboren te Sohag, Egypte

Promotiecommissie

Promotor: Prof. dr. W.P. de Voogt, Universiteit van Amsterdam

Promotor: Prof. dr. K. Kalbitz, Technische Universität Dresden

Copromotor: Dr. Boris Jansen, Universiteit van Amsterdam

Copromotor: Dr. John R. Parsons, Universiteit van Amsterdam

Overige leden: Prof. dr. ir. W. Bouten, Universiteit van Amsterdam

Prof. dr. J. Sevink, Universiteit van Amsterdam

dr. W. D. Gosling, Universiteit van Amsterdam

Prof. dr. E. Smolders, Universiteit Leuven

Prof. dr. A.H. El-Shater, Sohag University

Prof. dr. P.C. de Ruiter, Universiteit van Amsterdam

Faculteit der Natuurwetenschappen, Wiskunde en Informatica

This research was carried out at the Institute for Biodiversity and Ecosystem Dynamics (IBED), Faculty of Science, University of Amsterdam (Amsterdam, The Netherlands). This study was financially supported by

the Egyptian higher Education ministry and by IBED, University of Amsterdam.

ISBN: 978-94-91407-46-8

Printed by: Ipskamp Drukkers B.V.

Cover design: Sara Mohamed Samir

Content layout: Yasser Mohammed

Available as pdf via the Universiteitsbibliotheek Amsterdam at http://hdl.handle.net/11245/1.546657

Copyright © Yasser Refaey 2016

All rights reserved. No part of this publication may be reproduced or transmitted in any form or by any means, electronic or mechanical, including photocopy, recording or otherwise, nor may it be retained in any

information and retrieval system, without prior written permission from the author.

| 5

Table of contents

Chapter 1 7

General introduction

Chapter 2 15

Clay minerals of Pliocene deposits and their potential use for the purification of polluted wastewater in

the Sohag area, Egypt.

Chapter 3 37

The role of dissolved organic matter in adsorbing heavy metals in clay-rich soils.

Chapter 4 61

Effects of clay minerals, hydroxides, and timing of dissolved organic matter addition on the competitive

sorption of Copper, Nickel and Zinc: A column experiment.

Chapter 5 87

The influence of organo-metal interactions on regeneration of exhausted sorbent materials loaded with

heavy metals

Chapter 6 103

Synthesis

References 109

Summary in English 127

Samenvatting in het Nederlands 133

Summary in Arabic 139

Acknowledgements 145

List of Abbreviations 147

List of papers used in this thesis 148

Curriculum Vitae 149

6 |

GENERAL INTRODUCTION | 7

Chapter 1

General introduction

1. Background

1.1. Hazards of heavy metal contamination

With ongoing rapid industrialization and economic development, heavy metals

(HMs) continue to be introduced into the environment. Water contamination

with toxic HMs is a serious environmental issue and represents a hazard to

public health (Järup, 2003; Qin, 2006). The sources of toxic HMs include

domestic and industrial effluents (Lin and Juang, 2002; Jamil et al., 2010). The

main anthropogenic sources of HM contamination are disposal of untreated and

partially treated effluents containing toxic metals from mining and industrial

activities as a result of metal refinishing by products, as well as the use of HM-

containing fertilizer and pesticides in agricultural fields (Martin, 2000; Macklin

et al., 2006; Nouri et al., 2008; Reza and Singh, 2010).

Therefore, contamination with HMs is still an environmental problem today in

both developing and developed countries throughout the world (Inglezakis et

al., 2003; Dan’azum and Bichi, 2010; Momodu and Anyakora, 2010). In Egypt,

many of the industries discharge their wastewater either on the open desert area

or in surrounding surface water bodies (Fig. 1.1). These effluents cause

contamination with HMs of the soil, surface water and groundwater.

8 | CHAPTER 1

In addition, industrial wastewater is often used for the irrigation of non-

contaminated arable land in certain areas. Since this wastewater contains a

considerable amount of toxic HMs, it can negatively affect soil and groundwater

quality (Radwan and Salama, 2006). Metals including Cu, Zn and Ni are

considered the most hazardous and found to be common groundwater

contaminants in Egypt (e.g., Ayman and Mohamed, 2011; Zaki et al., 2015;

Chen et al., 2012; El-Badry, 2016). As result, a strong relationship was recorded

between drinking water contaminated with HMs and the incidence of chronic

diseases such as renal failure, liver cirrhosis, hair loss and chronic anemia

(Salem et a., 2000; Johri et al., 2010; Unisa et al., 2011).

Fig.1.1: A beverage factory discharge partially treated effluent into the Nile River at Aswan city, Egypt.

1.2. Using low-cost local materials as potential sorbents for removal of heavy metals

Numerous methods are commonly used to remove HMs from wastewater

solutions, including solvent extraction, precipitation, ion exchange,

phytoextraction, ultrafiltration, reverse osmosis, electrodialysis, and adsorption

(Donat et al., 2005). Recently, the use of alternative low-cost materials as

potential sorbents for the removal of HMs has been introduced to minimize the

problem of high costs decreasing the use of activated carbon despite its

effectiveness (Ali and Gupta 2007; Gupta et al., 2009; Akpomie and Dawodu,

2015). As the cost factors play a major role in treatment technology, efforts

have been directed towards looking for low-cost adsorbents for water

purification over the past years. Low-cost alternative adsorbents can be

prepared from a wide variety of local raw materials, which are abundant and

CHAPTER 1 | 9

cheap such as: agricultural waste and biomass materials, zeolites, and bentonite

(Moreno-Castilla and Rivera-Utrilla, 2001; Bhattacharyya et al., 2008; Vaghetti

et al., 2008; Dawodu et al., 2012). In developing countries, the optimization of

water and wastewater purification processes requires work on new operations

based on using low-cost local raw materials with high pollutant-removal

efficiency. In particular in Egypt, HM pollution can be severe and their removal

from wastewaters is crucial to protect public health (Mellah and Chegrouche,

1997; Jeon et al., 2001). In recent years the interest in application of hydroxides

such as Mn- and Fe-oxides and clay minerals as adsorption technology in

environmental clean-up, has increased (Colombani et al., 2015). The adsorption

of HMs on a variety of substances, such as activated carbon (Kadirvelu et al.,

2001) and clay minerals (Bhattachrayya and Gupta, 2008; Motsi et al., 2009), is

generally considered as the most powerful approach for wastewater cleanup.

However, as previously indicated, using activated carbon for the removal of

HMs at trace quantities is not suitable in developing countries because of the

high costs associated with production and regeneration of spent carbon (Panday

et al., 1985). In contrast, clay minerals are classified as low-cost adsorbents for

HMs from polluted wastewater. Given their high adsorption capacity, a very

interesting application of clay materials is to remove HMs from wastewaters,

particularly in developing countries such as Egypt, where more sophisticated

techniques are often not widely available (e.g., Ikhsan et al., 2005; Gu, et al.,

2010).

In the Egyptian Sohag area, a large reserve of Pliocene clay-rich deposits is

available which may offer great potential for removal of HMs from polluted

waters. The Egyptian Pliocene deposits are rich in smectite clays, a family of

common 2:1 phyllosilicates with a large permanent negative charge and a large

specific surface area (SSA) resulting in a large cation exchange capacity (CEC)

(e.g., Ikhsan et al., 2005; Gu, et al., 2010). Other constituents of the Pliocene

mineral phase that are important for the adsorption of HMs include Fe- and Mn-

(hydr)oxides (e.g., Sprynskyy et al., 2011). When clay minerals are used to

remove HMs in a wastewater treatment application, regeneration and reusability

of the spent sorbent material must be taken into account when assessing the

effectiveness and feasibility of the treatment process. After the adsorbents are

exhausted, they are either to be discarded of or recovered for reuse. Alteration

10 | CHAPTER 1

in conditions of metal-loaded sorbent may result in release of the contaminants

into the soil solution, thereby causing pollution of groundwater (Karathanasis,

1999). For that reason, the used adsorbents should be released into the

environment only after recovery of the adsorbed HMs (Lata et al., 2015).

1.3. Influence of (dissolved) organic matter on the mobility/immobility of heavy

metals

Organic matter (OM) is an abundant component in most soils (Troeh et al.,

2005) and because of its high CEC has the potential of significantly influencing

the mobility of HMs (Lu and Xu, 2009). The efficiency of clay minerals to

remove HMs from solution might be affected by the presence of organic ligands

in the effluents. Sorption of dissolved organic matter (DOM) to mineral

surfaces is considered an important pathway for the retention and also the

stabilization of OM (e.g., Kaiser and Guggenberger, 2000; Kalbitz et al., 2005;

Mikutta et al., 2007). If DOM is sorbed to the solid phase it can serve as

additional adsorption medium for HMs. However, when bound to the mineral

phase, OM can also alter the physicochemical properties of clay minerals by

decreasing their SSA and thus their HM adsorption capacity (Kaiser and

Guggenberger, 2003; Wang and Xing, 2005). Because DOM is often present

either in the wastewater itself (e.g. industrial and agricultural effluents) or in the

soil (e.g. due to manuring), the effects of (D)OM on the interactions of HMs

with clay minerals are crucial factors to take into account in the context of HM

mobility in soils in general, and specifically when assessing the applicability of

clay minerals as a simple wastewater treatment method (Arshad et al., 2008;

Cecchi et al., 2008). However, such effects have received surprisingly little

research attention so far, in particular where kinetic effects such as differences

in the timing of the presence of DOM and HMs (concurrently or sequentially)

are concerned.

CHAPTER 1 | 11

2. Aim and Objectives of the study

The main aim of this study was to investigate the potential of using clay

minerals abundant in local soils in Egypt as low cost materials to reduce Cu, Ni

and Zn pollution of soil and groundwater originating from polluted wastewater;

specifically focusing on the influence of the interaction of clay minerals and

heavy metals with OM already present in the soil, or in the wastewater itself. To

achieve this aim, the following specific objectives were formulated:

To identify and characterize the different clay mineral types in the context

of their application in local wastewater treatment in Egypt.

To investigate the influence of the presence of DOM and the timing of its

application (before or concurrently with the HMs) on the mobility of Cu, Ni

and Zn in clay-rich deposits using a static batch approach.

To unravel the effect of the timing of the addition of DOM on the

competitive adsorption of Cu, Ni and Zn onto different sorbent

compositions in a kinetic system using a dynamic column approach.

To investigate the role of the presence and timing of addition of DOM

during loading of clay mineral-based wastewater treatment columns on the

subsequent removal of the HMs from the columns, focusing both on

regenerating clay minerals used in wastewater treatment, and

(re)mobilization of HMs previously immobilized in clay rich soils.

3. Outline and structure of the thesis

The thesis is divided into six chapters to achieve the aim and the specific goals

mentioned above.

Chapter 1 contains a general introduction to the topic under study, its aims and

objectives, and its scientific and societal background and relevance.

Chapter 2 investigates the composition of the clay fraction of the Pliocene clay

deposits in the Sohag area, Egypt to obtain insights into the origin of the clay

12 | CHAPTER 1

deposits, and to assess their potential for use in low-cost wastewater purification

technology. The rationale for the latter was that both industrial wastewater and

irrigation water in Egypt are often polluted with HMs. In this chapter, the

physico-chemical analyses with detailed X-ray diffraction (XRD) mineralogical

investigations and SSA measurements of samples from 16 locations in four

areas containing Pliocene clay deposits were identified. To identify the different

clay minerals groups, the clay-sized fractions of the Pliocene deposits were

treatment using different cation saturation (K+ and Mg

2+) and different heat

treatments (25C, 300C, and 550C) on the XRD patterns of oriented

aggregates of the clay-sized fractions. Furthermore, analyses of both grain-size

and clay minerals assemblage were employed to estimate the source area and

climatic conditions during the period of deposition.

Chapter 3 represents the first step towards examining and assessing the

potential of using the Pliocene clay deposits as scavenger materials for HMs

using batch experiments. Batch adsorption experiments were performed to

investigate the interactions of Cu, Zn and Ni with both Pliocene clay deposits

and DOM to predict the fate of the three HMs under three different scenarios.

The initial mass (IM) isotherm approach of Nodvin et al. (1986) was employed

to describe the adsorption processes. This chapter also evaluates the influence of

the timing of DOM addition (before or concurrently with the HMs) as DOM is

often present either in the wastewater itself (e.g., industrial and agricultural

effluents) or in the soil (e.g. due to manuring). Therefore, the presence of DOM

can exert a significant influence on the fate and transport of HMs in soil/sorbent

materials.

Chapter 4 builds upon the insights acquired in chapter 3 to study HM

adsorption and the influence of the presence and timing of addition of DOM in

dynamic column experiments to better approximate actual conditions in the

environment and/or during the application of clay minerals to clean up polluted

wastewater. To this end, well-defined soil samples were amended with three

different minerals: goethite, birnessite and smectite, and subjected to three

different solutions containing a mixture of Cu, Ni and Zn under various flow

scenarios (A, B and C): A) absence of DOM; B) sequential addition of first

DOM and then HMs; C) concurrent addition of HMs and DOM. Adsorption

CHAPTER 1 | 13

parameters were determined quantitatively using the modified dose-response

model.

Chapter 5 represents the next step after HMs have entered the soil environment

and/or have been adsorbed on clay minerals in a wastewater cleanup setup, and

focusses on the desorption of HMs adsorbed on the clay mineral rich soil

columns subjected to the different adsorption scenarios described in Chapter 4.

As such it sets out to fill the knowledge gap concerning regeneration or

reusability of sorbent materials after having been loaded with Cu, Ni and Zn. To

this end the columns previously loaded with HMs were leached with 0.001 M

CaCl2 dissolved in water as a control eluent and 0.001 M CaCl2 dissolved in

DOM as a treatment eluent. The removal efficiency (E%) of the HMs was

calculated from the numerical integration of the regeneration curves. The results

of this chapter have important consequences for the regeneration potential of

clay minerals used in wastewater treatment aimed at removal of HMs and to

assess the potential of (re)mobilization of HMs adsorbed in a clay rich soil

environment.

Chapter 6 synthesizes and integrates the main findings of the research presented

in the previous chapters. In this chapter the obtained results and their

implications are discussed in a coherent overview of the research.

14 |

CHAPTER 2 | 15

Chapter 2

Clay minerals of Pliocene deposits and their potential use for

the purification of polluted wastewater in the Sohag area, Egypt

Abstract

In our study we investigated the clay fraction composition of Pliocene clay deposits in

the Sohag area, Egypt. Our goal was to obtain insights into the origin of the deposits,

and to assess their potential for use in inexpensive wastewater purification. The

rationale for the latter was that in Egypt both industrial wastewater and irrigation water

are often polluted with heavy metals (HMs), the load of which can be significantly

reduced using the Pliocene clay. We combined physico-chemical analyses with detailed

X-ray diffraction (XRD) mineralogical investigations and Specific Surface Area (SSA)

measurements of samples from 16 locations in four areas containing Pliocene clay

deposits. The grain size distribution of the studied samples was dominated by silt (75-89

%) with lower quantities of clay (6-20 %) and sand (2-15%). Neither grain size

distribution nor the distribution of individual clay minerals varied between the tested

samples, suggesting that they all originate from a single source area. The effect of

differential cation saturation (K+ and Mg

2+) and differential heat treatments (25C,

300C, and 550C) on the XRD patterns of oriented aggregates of the clay-sized

fractions revealed 4 different clay mineral groups in the tested samples. The relative

abundances of the clay minerals were semi-quantified and revealed a dominance of

smectite (69-91% on average) with relatively low contents of kaolinite (9-29% on

average) and minor amounts of illite (1-7% on average) and chlorite (0 ≤ 1%). This

mineral assemblage suggests chemical weathering and indicates warm climatic

conditions of the source area during the period of deposition. The higher CEC values of

the Pliocene clay deposits (32.3-65.4 cmolc/kg) also pointed to the occurrence of

smectite in the soils. The SSA of the Pliocene clay fractions (26.25-128.97 m2/g)

correlated well with their exchangeable cation contents (K+ and Ca

2+, R

2 = 0.96 and 1.0,

respectively) and micropore volumes (R2

= 1.0). Micropore volumes and SSA of the

studied samples increased with the size of the exchanged cation: K+

> Ca2+

> Na+. The

mineralogical composition suggests that Pliocen smectite-rich deposits in the studied

area have great potential to be used as raw material for inexpensive, local purification of

wastewater polluted with HMs.

This chapter is based on: Refaey, Y., Jansen, B., El-Shater, A., El-Haddad, A., Kalbitz, K. 2015 Clay minerals

of Pliocene deposits and their potential use for the purification of polluted wastewater in the Sohag area,

Egypt. Geoderma Regional 5, 215-225.

16 | CHAPTER 2

1. Introduction

Clay materials (< 2 μm size) are an important, abundant and inexpensive natural

resource for many applications such as the process industries, agriculture,

environmental remediation and construction (Pedro, 1994; Murray, 1999;

Murray, 2000; Sanfeliu et al., 2002; Gomes and Silva, 2007). The main

properties that make clay materials such an important resource are, inter alia,

their high specific surface area, adsorptive capacity, rheological properties,

chemical inertness and, depending on their chemical composition, absence of

toxicity (Dixon and Weed, 1989; Lin et al., 2002; Carretero et al., 2006).

Given their high adsorptive capacity, a very interesting application of clay

materials is as a low cost agent to remove heavy metals (HMs) from

wastewaters, particularly in developing countries such as Egypt, where more

sophisticated techniques are often not widely available (e.g., Srivastava et al.,

1989; Ikhsan et al., 2005; Gu, et al., 2010; Refaey et al., 2014). In the Egyptian

Sohag area a large reserve of Pliocene clay deposits is present. In a recent pilot

study we showed that clay materials from several locations in this area offer

great potential for removal of HMs from polluted waters, both in the presence

and absence of naturally-occurring organic matter that might interfere with such

an application (Refaey et al., 2014). A next step towards application of clay

materials from the Sohag area for wastewater treatment is a thorough

characterization of their clay mineral assemblage.

The clay mineral assemblage determines properties related to the retention of

HMs, such as cation exchange capacity (CEC) and specific surface area (SSA),

but also properties like swelling ability and plasticity that are crucial for the

technical applicability of the clays in wastewater cleanup. With respect to CEC,

the clay mineral composition determines the amount of permanent negative

charge, and contributes to a larger or smaller extent to the amount of variable

charged sites present (McBride 1994). For instance, kaolins have only a modest

amount of permanent negative charge due to limited isomorphic substitution

and only a modest amount of residual charge at their edges and from exposed

basal hydroxyls (Grim 1968; Bolland et al., 1976; Murray, 1999). On the other

hand, due to extensive isomorphic substitution, smectites and vermiculites have

high permanent negative charge and thus are responsible for most of the high

CHAPTER 2 | 17

CECs generally found in soils (Aparicio et al., 2010). The SSA is mainly

influenced by the grain size distribution, CEC, geotechnical characteristics and

types and amounts of clay minerals and is considered one of the most important

parameters that quantify interaction processes at the liquid-solid interface (e.g.,

Yukselen-Aksoy and Kaya, 2010; Heister, 2014).

Clay mineral distribution in parent material of marine origin, like in the Sohag

region in Egypt, generally reflects varying climatic zones in the clastic source

areas in addition to means of transport (Biscaye, 1965; Griffin et al., 1968;

Rateev et al., 1969). It therefore has also been used for paleoclimatic

reconstructions (Singer, 1984; Chamley, 1989). Characterization of Sohag

region clay minerals therefore not only is important to assess their technical

applicability in local wastewater treatment, but will also shed light on the

climatic and environmental records of the sediments, yielding valuable insights

in the history and genesis of the region (El-Shahat et al., 1997).

Therefore, the main objectives of this study were: (i) discriminating the

Pliocene deposits from several locations in the Egyptian Sohag area on the basis

of their textural and mineralogical attributes, (ii) identifying and characterizing

the different clay mineral types in the context of their application in local

wastewater treatment, and (iii) shedding light on the paleoclimatic conditions

that prevailed during formation of the sediments and their influence on the

sediment’s composition. To this end, X-ray diffraction (XRD) analyses were

applied on air-dried, heated and glycerol-saturated oriented preparations with

prior saturation with K+ and Mg

2+.

1. Materials and methods

2.1. Area of study and sedimentary successions

The study area is presented in Fig. 1.2 and is referred to as the Sohag area given

the central location of this town in the study area. The area is bordered from

both the east and west by a higher relief Eocene limestone plateau. Fig. 1.2

presents a simplified geological map of the area.

18 | CHAPTER 2

The Sohag area is located at East and West Sohag governorate, Egypt, in the

middle part of the Nile Valley that is situated south of Cairo (460 km),

represented by the Nile basin stretch extending between 26 19′ 87′′ to 26 33’

08” N lat; 31 39′ 04′′ to 32 03′ 62′′ E long. The study area comprises various

sediments ranging in age from Lower Eocene to Recent (Said, 1990; Omer,

1996; Omer and Issawi, 1998; Hassan et al., 2005) as shown in Fig. 1.2. The

studied samples were collected from The Muneiha Formation (Pliocene)

deposits that are characterized by their high fine earth fraction and smectite

content.

Fig. 1.2: Simplified geological map of Sohag area (TEGPC and CONOCO, 1987) with indication of clay

mineral assemblages at each sampling area.

CHAPTER 2 | 19

The Muneiha Formation forms a single lithostratigraphic unit (Omer, 1996;

Omer and Issawi, 1998; Hassan et al., 2005), that is equivalent to the Madmoud

Formation of Said (1981). The Muneiha Formation includes estuarine fine

clastic sediments formed as a result of the rising and invasion of the

Mediterranean Sea through a long gulf extending from Cairo to Aswan in the

Pliocene (Issawi et al., 1978; Hassan et al., 2005). This Formation was divided

into two main divisions (lower and upper members) based on its deposition

environment and facies (Omer, 1996). Fig. 2.2 represents the deposit. The lower

part is composed mainly of bedded to massive dark brown (chocolate) clay to

dark grey marine clay with thin interbeds of fine sand and silt (Fig. 2.2). The

upper part is dominated by fluvial sediments consisting of fining upward cycles.

The sediments are sloped toward the cultivated flood plain covered with the

younger sediments and widespread in both the surface and subsurface in the

study area.

Fig. 2.2: Outcrop field photograph of the Al-Kola area showing an alternation of Pliocene clay-rich beds with

thin interbeds of fine sand and silt laminae.

20 | CHAPTER 2

2.2. Sampling

Sixteen samples were selected from the larger suite of twenty seven soils

sampled from the Muneiha formation in the Egyptian Sohag region. The

samples were collected from outcrops of this Pliocene deposit at the east and

west bank of the Nile River. The Pliocene clay deposits constitute the main

outcrops in the studied area. They lack diagnostic soil horizons and are capped

by poorly consolidated sand, gravel, and clay of Quaternary age. The wadi

deposits, including the Pliocene deposits, in the Sohag area are infertile and

generally classify as Calcaric Fluvisols according to the World Reference Base

for Soil Resources 2006 (Jones et al., 2013).

Selection of the sixteen samples was based on analyses of grain size

distribution, texture and clay mineral composition such that the samples

represented a variation in physico-chemical characteristics linked to potential

use for large scale application purposes revolving around treatment of

wastewater. Specifically, multiple spatially distributed samples were taken from

the Al-Kwamel (KW; 5 samples), Al-Kola (KO; 3 samples), Al-Ahaywa (AH; 2

samples) and Wadi Qasab (WQ; 6 samples) areas. The samples from KW, KO,

AH, and WQ were collected along the surface of vertical exposures, i.e. both

artificial and natural outcrops in the field at heights of 1 to up to 10 m (see Fig.

2.2 for an example). The geographic distribution of the sampling sites is

displayed in Fig. 1.2. The samples were transported from Egypt to The

Netherlands in sealed plastic bags and stored at 4C until analyzed.

2.3. Physico-chemical characteristics of the Pliocene clay deposits

The samples were first air dried, then gently crushed by means of an agate

mortar and pestle to pass through a 2-mm sieve. Total carbon (TC) and total

nitrogen (TN) contents in the soils were determined with a C/N analyzer

(Elementar Vario EL, Hanau, Germany). The total content of Fe-oxyhydroxides

was estimated as dithionite-citrate-bicarbonate extractable iron (Fed) (AAS,

Perkin Elmer, Waltham, Massachusetts, USA) using the method of Mehra and

Jackson (1960) and Holmgren (1967). Mn-oxide and short-range-order (oxalate

extractable) Fe- and Al- (hydr)oxide (Feo and Alo) contents were measured

using the method of Searle and Daly (1977). Field water content was

CHAPTER 2 | 21

determined by drying soil samples at 105◦C for 24 h. The soil pHH2O was also

measured (1:2.5 v/v ratio). The CEC of the soils was determined using the

method of Hendershot and Duquette (1986). Major cations (Ca2+

, Mg2+

, and K+,

and Na+) and major anions (Cl

-, SO4

2-, and PO4

3-) were measured using an ICP-

OES (Perkin Elmer-Optima 3000XL) and San+2

Automated Wet Chemistry

Analyzer-Continuous Flow Analyzer (CFA), respectively.

Before determining the grain size distribution of the studied soils, they were

treated with 1 M HCl and H2O2 (30%) to remove respectively carbonates and

organic matter contents. Grain size distributions were determined on the basis of

sieving for the coarse component (> 150 µm) and a SediGraph (model 5100

grain size analyzer) for the fine component (< 150 µm) according to Stein

(1985) and Jones et al. (1988). Silt and clay fractions (< 63 µm) were separated

from the remaining portion by sieving. Then, the clay fraction (< 2 μm) was

separated using the sedimentation-decantation technique according to Jackson

(1969).

SSA measurements were performed at the Van 't Hoff Institute for Molecular

Sciences, University of Amsterdam, The Netherlands, using CO2 at 273 K on

Thermo Scientific Surfer instrument. CO2 gas adsorption was used. Given the

higher temperatures it employs as opposed to N2 physisorption and it is

considered more suitable for SSA determination of sediments and soil material

(e.g., Echeverría et al., 1999; Kwon and Pignatello, 2005; Eusterhues et al.,

2011). Four samples (KW-2, KO-2, WQ-2, and AH-2) were selected for SSA

measurement from the soils sampled in the study area; representing one sample

for each area. The selection was such that the samples represented a variation in

physico-chemical characteristics. The SSA was calculated according to the

Dubinin-Radushkevich equation (Dubinin and Radushkevich, 1947). Prior to

the measurements, the samples were outgassed for at least 24 h at 200 °C in

vacuum to remove adsorbed water.

22 | CHAPTER 2

2.4. Qualitative and semi-quantitative analysis of clay assemblages using XRD

2.4.1. Sample preparation for XRD analysis

The separated clay fraction was divided into two parts; the first one was

saturated with K+ and the second one with Mg

2+. Demineralized water and

centrifugation were then used (2575 x g) to remove excess salts after saturation

(Whittig, 1965). Afterwards, the samples were freeze-dried and kept for mineral

identification.

To prepare oriented aggregates for XRD analysis about 25 mg from each freeze-

dried clay fraction sample was added to 10 ml demineralized water in a 25 ml

volume tube and mixed well ultrasonically (5 sec). The mixture was deposited

gravimetrically on a porous mounting medium (ceramic tile, 37.2 mm in

diameter and 6.2 mm thick) connected to a funnel under vacuum that provided

the preferred orientation. After subsequent air-drying, the samples were ready

for XRD analysis. For each sample, five X-ray diffractograms were taken; Mg-

saturated samples were X-rayed in the air-dried and glycerol solvated states.

The K-saturated samples were X-rayed after air drying and heating to 300 and

550C for 2 h (Bouchet et al., 1988).

2.4.2. XRD analysis

XRD analysis was performed at the Van der Waals-Zeeman Institute,

University of Amsterdam, The Netherlands, using a Philips (now PANalytical)

PW 1830 instrument, with a Philips PW 3710 control unit (Cu Kα radiation with

wavelength 1.54056 Å produced at 50 mA and 40 kV). Minerals were identified

by characteristic reflections as discussed in Brindley and Brown (1980) and

Moore and Reynolds (1997). The relative percentages (semi-quantitative) of

clay minerals were determined using empirically estimated weighting factors of

Biscaye (Biscaye, 1965). The low chlorite and illite contents in studied samples

were estimated from the relative peak height (Johns et al., 1954) because they

could not to be detected in the glycerol solvated states.

CHAPTER 2 | 23

a b c

3. Results

3.1. Physico-chemical properties of Pliocene clay deposits

Field observations showed that the Pliocene sediments have slickensides and

desiccated under formation of deep cracks in a roughly polygonal structure (Fig.

3.2 a;b;c). The studied samples had large CECs ranging from 32.3-65.4

cmolc/kg (Table 1.2). The CEC, water content of air-dried samples and amount

of clay fraction were strongly correlated (both R2 = 0.75). Crystalline iron-oxide

contents were small to moderate (3.7-17.6 g kg-1

), while soil organic carbon

(SOC) contents were low in all samples (0-5.35 g kg-1

; Table 1.2). The pH was

always slightly basic (Table 1.2). Na+ was the dominant exchangeable cation

with the highest abundance of Na+ in sediments sampled from the KW area,

while Ca+2

, Mg+2

and K+ provided minor contributions (Table 1.2).

Fig. 3.2:. a) Pliocene deposits upon swelling and shrinking in wet and dry conditions display slickenlines on slickensides (polished and striated) (WQ area), b) deep wide cracks forming wedge-shaped or parallel-sided

aggregates (KO area), and c) polygonal patterns (AH area).

3.2. Grain size distribution

Table 2.2 presents the grain size distribution, i.e. percentages of clay, silt and

sand in the various samples. The silt content along the study area did not vary

greatly in the studied samples (Table 2.2). The silt fraction dominated over the

other fractions in all studied samples (75-89%); the clay fraction varied from 6

to 20% and the sand content fluctuated between 2 and 15% (Table 2.2). As a

result, the grain size distribution of all studied samples classified as silt.

24 | CHAPTER 2

Table 1.2: Selected physical and chemical properties of soil samples

#CBD ext. (Citrate Bicarbonate Dithionite extraction).

Samples

pH

EC

µs cm-1

H2O

%

CEC

cmolc kg-1

SOC CBD ext. Oxalate ext. _________ Major cations _________ _ Major anions_

Fe-oxide MnO2 Na+ Ca+2 Mg2+ K+ Cl -1 SO4-

_____________________________ g kg-1___________________________ ____ g L-1____

KW-1 7.58 9.9 8.40 60.48 0.43 17.13 31.54 5.34 0.65 0.23 0.012 7.26 0.17

KW-2 7.53 8.8 8.88 64.23 0.28 14.63 35.08 5.44 0.33 0.11 0.010 6.50 0.45

KW-3 7.80 14.5 6.90 43.21 0.21 17.42 35.90 9.01 1.69 0.26 0.016 8.15 5.92

KW-4 8.21 3.1 4.02 32.28 5.35 3.67 5.64 1.48 0.17 0.02 0.006 2.06 0.47

KW-5 7.81 5.9 7.58 51.38 0.36 5.12 3.59 2.77 0.81 0.09 0.011 3.11 2.85

KO-1 7.77 3.3 8.43 56.72 0.00 17.64 23.30 1.22 0.43 0.16 0.014 2.26 0.36

KO-2 7.61 5.0 10.36 65.36 1.72 9.69 13.05 2.09 0.68 0.21 0.011 3.56 0.17

KO-3 7.20 7.9 9.99 51.26 0.69 14.34 17.06 2.77 2.32 0.60 0.017 5.35 3.47

WQ-1 7.50 5.8 9.44 58.76 1.24 17.46 12.12 2.41 0.97 0.08 0.022 4.18 0.09

WQ-2 7.38 9.3 9.44 62.52 0.00 17.53 33.22 3.60 2.44 0.18 0.026 6.54 0.12

WQ-3 7.61 3.0 7.14 48.08 0.35 8.83 4.74 1.14 0.34 0.08 0.019 2.46 0.10

WQ-4 7.30 4.1 7.27 47.53 0.41 16.05 5.07 1.31 0.74 0.16 0.030 3.38 0.19

WQ-5 7.67 4.1 7.47 50.09 0.00 4.49 70.51 1.28 0.72 0.14 0.032 2.89 0.11

WQ-6 8.05 1.3 8.90 55.40 0.30 11.54 51.66 0.49 0.10 0.05 0.008 1.67 0.43

AH-1 7.73 7.5 7.96 55.36 0.85 10.63 20.96 4.07 0.54 0.18 0.010 3.82 0.05

AH-2 8.07 3.9 8.22 59.07 0.80 9.18 10.75 2.09 0.13 0.05 0.004 3.16 0.15

CHAPTER 2 | 25

Table 2.2: Grain size distribution (% of total mass) of the studied samples

Samples

Depths

m

Clays V. F. silt F. silt M. silt C. silt V. C. silt V. F. sand F. sand C. sand

< 2 2-4 4-8 8-16 16-31 31-63 63-125 125-250 500-1000

______________________________________ µm ________________________________________

Location 1

KW-1 5.0 9.2 5.4 7.2 12.2 23.1 32.7 7.2 0.3 2.8

KW-2 3.0 17.6 8.9 12.5 16.6 22.6 17.6 2.6 0.1 1.5

KW-3 2.0 8.0 7.0 12.3 19.7 29.0 21.0 2.3 0.1 0.6

Location 2

KW-4 4.0 6.3 7.8 12.5 19.1 19.8 22.7 9.1 0.7 2.0

KW-5 3.5 6.7 6.9 10.9 14.8 18.5 26.9 9.4 0.5 5.4

Location 3

KO-1 3.0 7.2 6.5 9.3 15.8 27.6 27.0 5.5 0.3 0.8

KO-2 2.0 18.2 14.8 17.8 20.2 17.3 8.4 1.1 0.0 2.2

KO-3 1.0 14.7 12.5 17.8 21.2 19.2 12.7 1.3 0.1 0.5

Location 4

WQ-1 4.0 11.3 10.2 14.2 20.6 22.6 17.6 3.6 0.2 0.0

WQ-2 1.0 14.7 15.0 21.5 24.1 16.5 6.6 0.4 0.1 1.0

Location 5

WQ-3 4.0 7.2 6.4 8.9 15.9 23.6 28.5 7.2 0.4 2.0

WQ-4 3.0 12.9 9.3 15.1 21.0 23.8 14.4 1.3 0.2 2.1

WQ-5 2.0 14.8 13.0 17.7 19.4 18.5 11.8 2.7 0.3 2.1

Location 6

WQ-6 2.0 18.5 9.9 14.4 21.2 19.7 13.4 2.2 0.0 0.7

Location 7

AH-1 5.0 19.8 15.2 16.4 17.3 14.6 11.4 2.5 0.3 2.7

AH-2 1.0 18.3 14.9 15.7 18.5 16.9 10.0 1.8 0.2 3.8

# V.F. = Very fine; F. = Fine; M. = Medium; C. = Coarse; V.C. = Very coarse.

26 | CHAPTER 2

a b

c

3.3. Clay mineral composition

3.3.1. Smectite

The XRD patterns of Mg- and K-saturated clay fraction patterns confirmed that

smectite is present (Fig. 4.2). The basal spacing of smectite (001) in Mg-

saturated expanded from ~16 Å to ~18 Å after the glycerol solvation treatment.

Upon treatment with K+, the ~16 Å of smectite contracted to ~12.5 Å at room

temperature (25C), slightly collapsed to ~12 Å following heating to 300C, and

highly collapsed to ~11 Å at 550C (Fig. 4.2). Second-, third-, and fourth-order

basal reflections were identified in the XRD patterns of the untreated and

glycerol solvation treated samples. Only the first-, second-, and third-order

reflections were present in the heated samples (Fig. 4.2). The absence of

reflection in the glycerol solvated samples (Mg-Gl-treatments) between 6 and

92θ (Fig. 4.2), confirm that in all samples smectite was well crystallized and

did not contain interlayers of illite (Raigemborn et al., 2014).

Fig. 4.2: Selected XRD patterns of oriented mounts of the < 2 µm size soil fraction in the KW (a), KO (b), WQ (c), and AH (d) areas. Lower panels (Green and blue): Mg-saturated slides in air-dried state (Mg-AD) and after glycerol

solvation (Mg-Gl). Upper panels (violet, brown, and red): K-saturated slides in air-dried state (K-AD) after thermal

treatment (K-300 o C and K-500

o C). S: smectite; K: kaolinite; I: illite; Q: quartz; C: chlorite; F: Fe-oxide.

CHAPTER 2 | 27

The correlation between CEC vs. water content (R2 = 0.75) and CEC vs. amount

of clay fraction (R2 = 0.75) in the studied samples (Tables 1.2 and 2.2) confirms

once again the predominance of smectite clays and strongly influences the

ability of the smectite-rich sediments to retain water in interlayer sites (e.g.,

Henry, 1997; Eisenhour and Brown, 2009).

Na+ was the dominant exchangeable cation in all studied samples (Refaey et al.,

2014) and this indicates that the type of smectite is Na-smectite (Weaver, 1956;

Murray, 1999). Furthermore, the Na/Ca ratio of tested samples in the KW (3.4-

16.5), KO (2.1-3.1), WQ (1.5-4.9), and AH (7.5-16) sites revealed a high

expanding capacity of the Pliocene sediments (Karakaya et al., 2011). The

alternate swelling and shrinking of expanding smectite clays in the study area

resulted in deep cracks, slickensides and wedge-shaped structural elements (Fig.

3.2 a,b) during the dry season (Gray and Nickelsen, 1989; Youssef, 2008;

Ismaiel, 2013). The large content of Na+ over Ca

2+ and Mg

2+ cations in the

studied sediments which occupy most of the wadi terraces and part of the wadi

floors in the low desert zone of the study area probably was responsible for

swelling and cracking in the foundations in the new planning area (Mitchell,

1976; Youssef, 2008).

3.3.2. Kaolinite

Reflection peaks of kaolinite at ~7.15 Å and ~3.57 Å remained unchanged

when the samples were subjected to solvation with glycerol (Mg-saturated) and

upon heating to 300°C but they disappeared upon heating to 550°C (K-

saturated) as a result of the destruction of the structure. Such behavior is

characteristic of kaolinite (Moore and Reynolds, 1997; Refaey et al., 2008;

Hong et al., 2012; Tsao et al., 2013).

Kaolinite in the WQ and AH areas displayed narrow and sharp peaks (Fig. 4.2

c; d), indicating that it was highly crystalline (Brindley and Brown, 1980). Less

narrow but well-defined peaks (Fig. 4.2 a; b) were also present at infrequent

levels in the KW and KO areas, suggesting a slightly poorer degree of

crystallinity (Arslan and Aslan, 2006).

28 | CHAPTER 2

3.3.3. Illite and chlorite

Illite and chlorite were scarce or absent in most of the studied samples. The

presence of illite was confirmed in the current study by reflection peaks at

10.13, 4.53, and 3.33 Å d-spacing, which remained unchanged when the

samples were subjected to heating treatments (300°C and 500°C). Furthermore,

the 3.33 Å (003) reflection of illite was developed with quartz (010). The K+

content in all tested soil samples was the lowest among all exchangeable

cations, which is in line with the scarcity of illite as a mineral constituent of the

clay fractions (Hower and Mowatt, 1966; Inoue and Utada, 1983; Brusewitz,

1986; Velde, 1986; Refaey et al., 2014).

Chlorite minerals were detected only in the WQ and AH areas as a very weak

peak appeared in the position of the kaolinite peak after heating treatment to

550C (Fig.4.2 c). Also, a weak peak was recognized as a shoulder at 3.46 Å

that showed no change upon glycerol solvation and heating to 550C (Fig.4.2

c).

3.3.4. Non-clay minerals

Non-clay minerals such as quartz and hematite were recognized as weak

reflection peaks at 4.21 and 2.69 Å d-spacing, respectively, indicating that only

trace amounts of non-clay minerals were present. To exclude the possibility of

the quartz peaks having originated from the ceramic tile itself, we compared X-

ray analyses of the tile with and without sample material present. None of the

characteristic peaks found for the empty tiles (results not shown) appeared in

the analyses when sample material was present.

3.4. Qualitative and semi-quantitative description of experimental XRD patterns

The intensity of the clay mineral peaks changed notably in the different sampled

areas, indicating variations in relative proportions of clay species of the studied

samples (Fig. 4.2). On the one hand, in all studied samples smectite had the

highest, sharpest and most symmetrical peaks (001 reflection) indicating

predominance of smectite over the other clay minerals in the assemblage. On

the other hand, the illite and chlorite mineral groups, as well as the non-clay

CHAPTER 2 | 29

minerals such as quartz and Fe-oxide (hematite/goethite), were present as weak

peaks indicating trace amounts (Fig. 4.2). There was no indication of the

presence of any regularly stratified mixed-layer clay. Therefore, all expandable

clay is treated here as highly expandable smectite.

The clay mineral assemblages of the Pliocene clay deposits showed an overall

similar composition with smectite as the most abundant class, followed by

moderate amounts of kaolinite and scarce amounts of illite and chlorite (1-7%

and < 1% abundance, respectively) (Table 3.2). Therefore, representative XRD

patterns were selected to illustrate clay mineral assemblages with different

treatments as shown in Fig. 4.2. The average of the smectite content as

determined by the Biscay method (Biscay, 1965) and the peak intensity

methods (Johns et al., 1954) in the KW, KO, WQ, and AH areas was 91, 80, 74,

and 69%, whereas the proportion of kaolinite equaled 9, 18, 24, and 29%,

respectively (Table 3.2). The smectite proportional percentages in all studied

areas increased in the order KW > KO > WQ > AH (see Table 3.2).

Table 3.2: Relative clay mineral abundance in the clay fraction of sediment samples based on the

peak height (method according to Johns et al., 1954 and Biscay, 1965).

Samples Clay

fraction _____ Peak height method (Johns et al., 1954) _____ _Biscay’s method (1965)_

_Smectite_ _Kaolinite_ __ Illite __ _ Chlorite_ _Smectite_ _Kaolinite_

mg g-1 Rel.

% mg g-1

Rel. %

mg g-1

Rel. %

mg g-1

Rel. %

mg g-1

Wt. %

mg g-1

Wt. %

mg g-1

KW-1 92 94 86 4 4 2 2 - - 89 82 11 10 KW-2 176 87 153 10 18 3 5 - - 90 158 10 18

KW-3 80 85 68 12 10 3 2 - - 94 75 6 5

KW-4 63 93 59 7 4 1 1 - - 94 59 6 4 KW-5 67 89 60 10 7 2 1 - - 90 60 10 7

KO-1 72 77 55 18 13 5 4 - - 86 62 14 10

KO-2 182 72 131 22 40 6 11 - - 87 158 13 24 KO-3 147 71 105 24 35 5 7 - - 85 125 15 22

WQ-1 113 68 77 25 28 7 8 - - 84 95 16 18

WQ-2 147 62 91 35 52 3 4 - - 82 121 18 27 WQ-3 72 68 49 28 20 4 3 - - 84 61 16 12

WQ-4 129 59 76 35 45 5 6 < 1 - 81 104 19 25

WQ_5 148 66 98 30 44 4 6 < 1 - 81 120 19 28 WQ-6 185 67 124 29 54 4 7 < 1 - 81 150 19 35

AH-1 198 60 119 38 75 2 4 < 1 - 75 148 25 49

AH-2 183 66 121 30 55 4 7 < 1 - 76 139 24 44 # Rel. % = relatively peak height % of each clay mineral; Wt. % = relative weight percentages of each clay mineral.

30 | CHAPTER 2

3.5. The SSA of the clay fraction

The SSAs in the representative samples of KW-2, KO-2, W-2, and AH-2 were

34.0, 47.8, 129.0, and 26.2 m2/g, respectively, whereas the micropore volumes

were 6.02, 8.48, 22.86 and 4.65 cm3/g respectively. The results show SSA to be

related to the type of exchangeable cations, in the following order: K+ > Ca

2+ >

Na+. In all samples, strong correlations were observed of the SSA with the

amount of K+ (R

2 = 0.96), Ca

2+ (R

2 = 1.0), and micropore volumes (R

2 = 1.0).

4. Discussion

4.1. Origin and genesis of clay minerals

The grain size distribution and clay mineral assemblage in the Sohag area can

give an important first idea about the palaeoclimate conditions and weathering

processes at the source area (Velde and Meunier, 2008; Agha et al., 2013). The

fact that the grain size distribution of the studied samples exhibited no large

variation in sand, silt, and clay fraction contents nor distribution of individual

clay minerals suggests that the sediments derived from one source area.

Moreover, the high amounts of silt in comparison to sand in all studied samples

(Table 2.2) suggest that the sediments were deposited from suspension and were

formed under uniform conditions of slow moving water (Ghandour et al., 2004).

During the Pliocene, East Africa was characterized by seasonal, arid and warm

environments (Jacobs et al., 1999). The Late Pliocene climate, specifically in

Egypt, was arid to semi-arid with seasonal runoff that resulted in a prevalence

of grasslands (e.g., Griffin, 2002; Swezey, 2009; Talbot and Williams, 2009;

Agha et al., 2013). A study by El-Shahat et al. (1997) of Pliocene sediment

from the North western desert, Egypt, indicates an initial provenance of

metamorphic and acidic igneous rocks from the Red Sea highlands. The

weathered regolith of the Red Sea basement rocks must have been eroded by

several tributaries that fed a master stream (Paleo-Nile) during the late Pliocene

pluvial (Said, 1981; El-Shahat et al., 1997). The genesis of smectite is favored

by dry seasons alternating with less pronounced wet seasons (Singer, 1984),

poorly drained environments (Schaetzl and Anderson, 2005) as well as low-

lying topography such as in marine environments (Odoma et al., 2013). All

CHAPTER 2 | 31

these features, in particular the dominance of smectite, also indicative of

chemical weathering (Raigemborn et al., 2014), are present in the study area.

This suggests the Pliocene deposits in the Sohag region were deposited in a

marine environment under arid to semi-arid climatic conditions (Salman et al.,

2013), in particular in a warm climate with alternating pronounced dry and less

pronounced wet seasons (Lewis and Berry, 1988; Ghandour et al., 2004;

Ehrmann, et al., 2007). In general, the presence of abundant smectite in the

studied area is generally linked to a transgression of the sea in the Pliocene

(Tantawy et al., 2001). In addition, the presence of kaolinite as second abundant

clay mineral is indicative of chemical weathering of acidic igneous and

metamorphic rocks or their detrital weathering products under tropical to

subtropical humid climatic conditions (Hendriks, 1985; Marzouk, 1985;

Chamley, 1989; Refaey et al., 2008). The lower abundance of kaolinite relative

to smectite in the present study, especially in the KW area (West bank of Nile

River; Fig. 1.2), situated at lower altitude than the other sampling sites (East

bank of Nile River), further confirms the earlier mentioned deposition from

suspension as a primary sedimentation mechanism. Kaolinite tends to

concentrate in relatively near-shore shallow water settings, in line with its

tendency to flocculate as coarser grains than smectite that tends to settle as finer

particles in deeper offshore settings (Raucsik and Merenyi, 2000; Thiry, 2000).

The absence of mixed layer minerals (Illite-smectite) implies that the origin of

clay minerals in the Pliocene deposits is of detrital origin where there is no

evidence for the influence of burial diagenesis which lead to conversion of

smectite to illite (Hoffman and Hower, 1979; Chamley, 1989; Ghandour et al.,

2004). Furthermore, it is implausible that the clay mineral assemblages in the

studied area originated from deep burial diagenetic alteration due to the lower

overburden thicknesses in the studied area (Agha et al., 2013). Nor was any

evidence for hydrothermal alteration observed during field sampling.

32 | CHAPTER 2

4.2. Implications for HM retention

It is well known that smectites have higher CEC values than kaolinites (e.g.,

Babel and Kurniawan, 2003; Aparicio et al., 2010; Hong et al., 2012). In the

KW and KO areas, soil KW-2 and soil KO-2 had the highest CEC values in line

with their having the largest clay fraction and highest proportion of smectite

compared to kaolinite (Table 1.2, 2.2). However, in the WQ and AH areas, soil

WQ-2 and soil AH-2 had the highest CEC values (Table 1.2) in spite of the fact

that they did not have the largest absolute clay content of the samples from the

area (Table 2.2). Therefore, in the KW and KO areas, in contrast to the WQ and

AH areas, it is the clay mineralogical composition rather than the absolute

amount of clay that determined the CEC. This observation is in line with the

results from other studies and is linked to the previously mentioned large

difference in CEC between smectites and kaolins, which in soils with

appreciable kaolinite contents overrides absolute amounts of clay minerals as

dominant factor (e.g., Rice et al., 1985; Parfitta et al., 1995; Usman, 2008;

Youssef, 2008). As a result of these differences in CEC, although our previous

study showed the capacity for HM (Cu, Ni and Zn) adsorption was high in all

soils, a higher affinity was found in the soils from the KW and KO than from

the WQ and AH areas (Refaey et al., 2014).

The measured SSA values (26.25-128.97 m2/g ) of the Pliocene clay fraction all

fell within the normal range (33-130 m2/g) for pure Na

+, K

+, Ca

2+, and Mg

2+-

smectite as determined by gas adsorption (Volzone and Ortiga, 2004; Kaufhold

et al., 2010). Compared with the KW-2, KO-2, and AH-2 samples, the WQ-2

sample showed a significantly higher SSA and micropores volume in line with

its higher K+ and Ca

2+ contents (Volzone and Ortiga, 2004, Ayari et al., 2007).

This can be explained by the fact that the ionic radius of the exchangeable

cations has a strong effect on the CO2 gas adsorption and consequently on the

value of SSA in the interlayer positions of this sample (Rutherford et al., 1997;

Volzone and Ortiga, 2004). Also in line with previous findings, the measured

SSAs and micropore volumes of our samples both decreased in the following

sequence: K-clay > Ca-clay >> Na-clay (e.g., Rutherford et al., 1997; Volzone

and Ortiga, 2004; Afsin et al., 2009). The large primary surfaces of the

investigated clays explain the previously observed large adsorption capacity of

the deposits in the study area for the Cu, Ni, and Zn (Refaey et al., 2014).

CHAPTER 2 | 33

4.3. Application in the region

Pliocene smectite-rich deposits in the studied area can be used as a potential raw

material for purification of wastewater from toxic HMs because of their very

fine particle size and physical and chemical properties, including SSA, that are

related to their clay mineralogical composition. Particular for the Sohag region,

discharge of large amounts of wastewater (sewage) used as irrigation water for

wood production as well as the use of high amounts of fertilizers in new

reclamation areas causes infiltration and accumulation of many pollutants to

ground water reservoirs, including HMs (Ayman and Mohamed, 2011). The

Pliocene clay deposits in the study area might be used to reduce the load of

HMs as well as organic pollutants from such sources. Furthermore, crude water

purification using Pliocene clay deposits may extend the applicability of the

treated water from wood production alone to include the irrigation of crops

(Rashed and Soltan, 2002).

In addition to potential use in regional wastewater (pre)treatment, the studied

Pliocene deposits could be utilized as liners (barrier) in landfills to control the

seepage of HM containing leachate into the surrounding environment (Abollino

et al., 2003; Talaat et al., 2011), a particular problem in the desert areas of the

Sohag region. The suitability of soil material as liners relates to its particle size

distribution, Atterberg limits, swelling potential, CEC and hydraulic

conductivity (Taha and Kabir, 2005). Specifically for landfill liners, preferred

value ranges for such parameters that have been specified in the literature

include: percentage of clay (< 2µm) ≥ 20, percentage of fines (< 75 µm) ≥ 50%,

CEC ≥ 10 meq/100g , plasticity index (PI) ≥ 12% and activity (AC) ≥ 0.3

(Rowe et al., 1995; Daniel, 1998; Taha and Kabir, 2005). The studied Pliocene

clay deposits from the Sohag area contain clay percentages between 8-20% and

percentages of fines between 81-98% (Table 1.2). In particular the more clay

rich samples (i.e., locations KO, WQ and AH) thereby come close to the

mentioned preferred values with respect to clay percentages, while always

exceeding the threshold values for the percentage of fines. Moreover, a study by

Youssef (2008) of the Pliocene deposits in the studied area revealed that these

deposits had the following Atterberg limits: liquid limit (LL) = 44-62%, plastic

limit (PL) = 26-38%, plasticity index (PI) = 14-36% and activity (AC) = 0.41-

0.72. These index properties fall within the mentioned preferred value range and

34 | CHAPTER 2

are indicative of a high shrinkage limit that will result in little volume change

of the material when used as landfill liners (Taha and Kabir, 2005). Note that

the high AC value of the studied Pliocene deposits is an indication of the low

hydraulic conductivity as well as of the high SSA of the clay fraction (Benson

et al., 1994; Taha and Kabir, 2005). In addition, the high CEC and SSA values

of liner material made from Pliocene deposits will result in a greater amount of

inorganic contaminants being removed from the leachate (Kayabali, 1997; Taha

and Kabir, 2005). Altogether this makes that the studied clay deposits, in

particular the ones with the higher clay percentages, seem well suited for

regional application as cheap landfill lining. Our present study can serve as a

foundation on which to build further explorations into such regional

applications.

In general, the present work supports the potential of the studied sediments as a

natural and inexpensive material for removing toxic HMs that was tentatively

established in our previous study (Refaey et al., 2014), and established the

underlying mechanisms of such applications. Specifically, we recommended

materials from the KW area for treatment of wastewater rich in HMs such as Zn

and Ni, as they showed a high affinity for these materials. On the other hand, we

proposed clay-rich materials from the KO, WQ and AH areas to be most

effective in the treatment of wastewater rich only in Cu (Refaey et al., 2014)

which is known to form strong inner-sphere complexes with surface of Fe-

(hydro)oxides in sediments even when the smectite content is less. However,

the new information obtained from the present study shows that, given the

significantly higher SSA and microporosity of the WQ-2 sample, material from

this area warrants extra research attention to be used as starting point for further

exploration in wastewater treatment, not only for HM removal but possibly also

for toxic gas adsorption (Volzone, 2007).

CHAPTER 2 | 35

5. Conclusions

The grain size distribution of the studied samples was dominated by silts (75-89

%) with lower quantities of clays (6-20%) and sands (2-15%). XRD analysis

demonstrated that the clay mineral composition of the Pliocene clay deposits

studied was composed almost exclusively of smectites and kaolinite, with the

former always being the most abundant class of clay minerals. The presence of

a large amount of smectite in association with a low quantity of kaolinite

minerals in our study suggests an origin from chemical weathering conditions

under warm and semi-arid conditions. Furthermore, the absence of mixed layer

clays confirmed that the tested sediments were derived from transported

weathered materials. The physico-chemical properties of the studied sediments

as well as the type and amount of smectites indicate that they have a high

capacity to immobilize large amount of dissolved HMs. The present study has

shown that there are strong relationships between the SSA and soil chemical

properties such as exchangeable cations where K+ and Ca

2+-rich clay fraction

tends to have higher SSA and micropore volume values. This makes the studied

sediments potentially useful in high-value-added markets, e.g., as

environmentally friendly and inexpensive raw material for waste water

treatment. To further examine such application, additional research should focus

on unraveling the mechanisms of such interactions, specifically in quasi-

realistic operational and field settings such as in column experiments.

Particularly interesting is the exploration of the material from the WQ area

giving its significantly higher SSA and microporosity than material from the

other areas.

36 |

CHAPTER 3 | 37

Chapter 3

The role of dissolved organic matter in adsorbing heavy metals

in clay-rich soils

Abstract

Heavy metals (HMs) are toxic to human life and the environment when present in

excessive concentrations. Therefore, determining the interactions of HMs with soils and

dissolved organic matter (DOM) is essential to predict their fate. To find out the effect

of DOM and soil properties (Clay minerals, oxides, and bulk organic matter [OM]) on

the uptake of Cu, Ni, and Zn, batch adsorption experiments were conducted using five

soils sampled from Egypt. The sorption isotherms were well described by the initial

mass (IM) isotherm model. The amount and timing of DOM addition was found to play

a pivotal role in determining the affinity of the HMs for soil. When DOM and HMs

were added simultaneously, the affinity of Cu decreased in Fe-(hydr)oxide-rich soils (by

7%) and increased in soils poor in Fe-(hydr)oxide (by 6-10%). When DOM was added

first, followed by HMs, the affinity of Cu strongly increased. In contrast, affinity of

both Ni and Zn was enhanced (3-18%) in the presence of DOM, regardless of the timing

of DOM addition. The difference is explained by Cu binding to the solid phase and

DOM through strong inner-sphere complexes, whereas Ni and Zn adsorbed

predominantly through weaker electrostatic interactions. As a result, Cu was able to

bind more strongly to previously adsorbed DOM on the solid phase in case of smectite,

while this effect was counteracted by the coating of available specific binding sites on

Fe-(hydr)oxides. The study has revealed that Egyptian soils hold great potential to

remove HMs from aqueous solutions.

This chapter is based on: Refaey, Y., Jansen, B., El-Shater, A., El-Haddad, A., Kalbitz, K. 2014. The role of dissolved organic matter in adsorbing heavy metals in clay-rich soils. Vadose Zone J., Vol. 13 No. 7.

38 | CHAPTER 3

2. Introduction

Industrial activities release significant quantities of different pollutants,

including HMs. In particular in developing countries, like Egypt, such pollution

can be severe, and the removal of HMs from (waste)waters is important to

protect public health (Mellah and Chegrouche, 1997).

Methods that are commonly used to remove HMs from wastewater solutions

include precipitation, ion exchange, solvent extraction, phytoextraction,

ultrafiltration, reverse osmosis, electrodialysis, and adsorption onto activated

carbon (Donat et al., 2005). Of these the adsorption of HMs on a variety of

substances, such as activated carbon (Kadirvelu et al., 2001) and clay minerals

(Erdem et al., 2004; Chen et al., 2007; Bhattachrayya and Gupta, 2008; Motsi et

al., 2009), are generally seen as the most powerful tools for wastewater cleanup.

Soils have been shown to play a pivotal role in mitigating the environmental

and health effects of HM pollution originating from (waste)waters, in particular

when polluted water sources are used to irrigate crops (Zeid, 2013). On the one

hand, adsorption processes in the soil may remove HMs from the dissolved

phase. On the other hand, infiltrating water may stimulate the release of HMs

that were previously adsorbed (Ahlberg et al., 2006). Clay minerals play an

important role in retaining HMs through adsorption. As a result, the types and

amounts of clay minerals present are of considerable importance in determining

the distribution and mobility of HMs in soils (Spark et al., 1995; Doula et al.,

1999; Al-Qunaibit et al., 2004; Srivastava et al., 2005). Because of their

adsorption potential for HMs, many clay minerals are even actively used to

remove HMs from polluted wastewater. For example, consider the use of

smectite clays, a family of common 2:1 phyllosilicates with a large permanent

negative charge through isomorphic substitution, and a large specific surface

area resulting in a large cation exchange capacity (CEC) (e.g., Ikhsan et al.,

2005; Gu, et al., 2010). Other constituents of the mineral soil phase that are

important for the adsorption of HMs include Fe-, Al-, and Mn-(hydr)oxides

(e.g., Sprynskyy et al., 2011).

Due to its high specific surface area and CEC, soil organic matter (SOM) also

plays a significant but complicated role in affecting the mobility of (heavy)

metals in soils (Stevenson, 1982). When present as part of the solid phase, SOM

CHAPTER 3 | 39

can serve as adsorption medium for HMs. However, when bound to the mineral

phase, SOM can also alter the physicochemical properties of clay minerals by

decreasing their specific surface area (SSA) and thus their HM adsorption

capacity (Kaiser and Guggenberger, 2003; Wang and Xing, 2005). When

present as dissolved organic matter (DOM), SOM influences the mobility of

(heavy) metals by forming soluble organo-metal complexes with organic

ligands (e.g. Chairidchai and Ritchie, 1990; Kalbitz and Wennrich, 1998; Alvim

Ferraz and Lourenco, 2000). The resulting organo-metal complexes may remain

dissolved, be weakly adsorbed to the soil surface, or form strong inner-sphere

complexes that are bound even more strongly than would the free metal ion

(e.g., Benjamin and Leckie, 1982; Udom et al., 2004).

In general, a complex interplay of different inorganic and organic soil

constituents is involved in the adsorption of HMs in soils through a continuum

of reactive sites, ranging from weak physical (Van der Waals) forces and

electrostatic outer-sphere complexes (e.g. ion exchange) to the formation of

strong chemical bonding (inner-sphere complexation) and precipitation

(Sposito, 1989). Previous work has shown that the complex interplay between

(i) clay minerals and metal oxides in the solid mineral phase, (ii) organic matter

in the solid phase, and (iii) DOM together determine the mobility and

translocation of Al and Fe in podzols (Jansen et al., 2004; 2005). However, the

timing of the addition of DOM is an aspect that has received little research

attention so far. When a metal binds to solid phase OM, it is immobilized. In

contrast, when it is bound to DOM, its immobilization on the solid soil phase

may be actively prevented if the resulting dissolved organic metal complex

remains in solution. Multidentate coordinative binding of multicharged HMs,

such as Zn, Cu and Ni to oxygen, containing functional groups on both

dissolved and solid phase organic matter, is often largely irreversible

(Stevenson, 1982; Karlsson et al., 2006). Consequently, one could expect that

the timing of when metal-rich (waste)waters pass through soils (before,

concurrently with, or after application of DOM-rich solutions) can significantly

affect mobility of such HMs. The net influence of adding DOM to a certain soil

system depends on a complex interplay of interactions with the DOM and soil

properties in the solid phase (Harter and Naidu, 1995). For example, several

mechanisms exist through which complexation of HMs with DOM leads to

40 | CHAPTER 3

immobilization of the HMs. Apart from precipitation of the complex, which

only occurs upon its saturation (Jansen et al., 2004), a first important

mechanism is cation bridging, that is, formation of S-HM-DOM complexes,

where S represents the adsorption site on soil surface and HM is the metal ion

(Stumm, 1992). This will not only lead to immobilization of DOM itself, but

can also result in enhanced binding of metals when metals subsequently bind to

the adsorbed DOM in a second layer as: S-HM-DOM-HM. This will be

particularly important for metals, such as Cu, that bind to DOM through inner-

sphere complexes; it should not be influenced by the timing of addition of

DOM. A second mechanism is adsorption of DOM on specific solid phase

sorption sites not involved in the binding of HMs, followed by adsorption of

HMs on the DOM that is thus adsorbed. Again, this is expected to be most

pronounced for metals, like Cu, that bind through inner-sphere complexation.

However, in this case, timing would have an influence: addition of DOM

followed by later addition of HMs would be expected to have a larger

immobilizing influence on HMs than concurrent addition. In spite of a vast

body of literature dealing with the interactions of (heavy) metals with DOM,

surprisingly little attention has been paid so far to such possible timing effects

and to interactions between HMs, soil mineral constituents, SOM, and DOM.

Therefore, the main purpose of this study was to increase the understanding

about the combinations of Cu, Ni, and Zn with clay minerals, oxides, bulk

SOM, and DOM in the context of metal pollution of Egyptian soils. In

particular, we focus on the role of organic matter in regulating the mobility of

Cu, Ni and Zn in clay-rich soils from Egypt through (i) the influence of OM-

coating on soils, and (ii) the influence of the presence of DOM and the timing of

its application (before or concurrently with the HMs). In all processes,

competitive sorption phenomena between the three metals were also explicitly

considered.

As an overarching method, we employed the IM isotherm approach of Nodvin

et al. (1986) to describe adsorption processes. The IM isotherm model has been

widely used to describe DOM adsorption and desorption to mineral soils (e.g.,

Kaiser et al., 1996).

CHAPTER 3 | 41

3. Materials and methods

3.1. Sampling and study area

Twenty-eight soil samples (Pliocene clay deposits) were collected at East and

West Sohag governorate, Egypt, midway between Cairo and Aswan.

Specifically, samples were taken from the Al-Kwamel (KW), El-Kola (KO), Al-

Ahayua (AH) and Wadi Qasab (WQ) areas. The study area is represented by the

Nile basin stretch extending between 26 19′ 87′′ to 26 33′ 08′′ N lat and 31

39′ 04′′ to 32 03′ 62′′ E long. In addition, one sample was collected from the

Bahariya Oasis (BO) area, Egypt (27 47′ 84′′ N lat and 28 31′ 68′′ E long).

Five Egyptian soils (KW, KO, AH, WQ and BO) were selected from the larger

suite of 28 soils sampled in areas representative for irrigation with HM polluted

wastewater. The selection was based on obtaining a realistic range in

physicochemical characteristics. The soil samples from KW, KO, AH, WQ, and

BO were collected along the surface of vertical exposures (i.e. both artificial

and natural outcrops in the field that is characterized by irregular surfaces, such

as terraces), at heights of 4.0, 2.0, 3.5, 1.0, and 0.5 m, respectively. We were

looking for clay-rich soil material that could be potentially used for large-scale

processing of wastewater. The samples were transported from Egypt to The

Netherlands in sealed plastic bags and stored at 4C until analyzed.

3.2. Physico-chemical and mineralogical characteristics of the studied soils

For mineralogical identification, X-ray diffraction (XRD) analysis was

performed at Van der Waals-Zeeman Institute, University of Amsterdam, The

Netherlands, using a Philips (now PANalytical) PW 1830 instrument, with a

control unit Philips PW 3710 (Cu Kα radiation with wavelength 1.54056 Å

produced at 50 mA and 40 kV) to identify the clay minerals present in the clay

fraction (Brindley and Brown, 1980). Total carbon (TC) and total nitrogen (TN)

contents in the soils were determined with a C/N analyzer (Elementar Vario

EL). We assume that TC equals total organic carbon (TOC) since TC equals the

sum of organic and inorganic carbon, and no carbonates were found in the

selected soil samples as tested by addition of 2 M HCl. Total content of

pedogenic (hydr)oxides was estimated as dithionite-citrate-bicarbonate

extractable iron (Fed) (AAS, Perkin Elmer) using the method of Mehra and

42 | CHAPTER 3

Jackson (1960) and Holmgren (1967). Manganese-oxide and active (oxalate

extractable) Fe- and Al- (hydr)oxide (Feo and Alo) contents were measured

using the method of Searle and Daly (1977). Field water content was

determined by drying soil samples at 105◦C for 24 h. The soil pHH2O was also

measured (1:2.5 ratio). The CEC of soils was measured using the method of

Hendershot and Duquette (1986). Major cations (Ca2+

, Mg2+

, and K+, and Na

+ )

and major anions (Cl-, SO4

2-, and PO4

3-) were measured using inductively

coupled plasma optical emission spectrometry ICP-OES (Perkin Elmer-Optima

3000XL) and San2+

Automated Wet Chemistry Analyzer-Continuous Flow

Analyzer, respectively.

3.3. Dissolved organic matter (DOM) preparation

DOM was prepared by aqueous extraction from organic materials (soil with

natural manure) obtained from a commercial garden center “Intratuin bemeste

tuinaarde” in Amsterdam. We chose this product because it contains cattle

manure representative for wastewater application in an agricultural setting. In

addition, to be allowed to be sold as fertilizer in The Netherland in compliance

with Dutch Law, its contents of HM are restricted, as confirmed by the analyses

of major cations (Table 1.3). The extraction was performed by adding 100 g of

OM to 1 L of deionized H2O. The suspension was stirred for 24 h at 180 rpm.

The suspensions were centrifuged at 984xg for 40 min after which they were

centrifuged again at high speed (27,586 x g) for 40 min at 25◦C. The

supernatants were filtered over a 0.2-µm cellulose-acetate membrane filter.

Dissolved organic carbon (DOC) was determined by a TOC analyzer (TOC-

VCPH-Shimadzu-Kyoto, Japan). Total organic carbon and TN contents in solid

OM were determined with a C/N analyzer (Elementar Vario EL). The pH of the

DOC extraction was adjusted to pH 6 by adding appropriate amounts of 0.1 and

0.01 M NaOH. The main physicochemical properties of DOM and extracted

manure are presented in Table 1.3.

CHAPTER 3 | 43

Table 1.3: Characteristics of DOM and the extracted manure used in adsorption experiments

DOM

Cu Ni Zn Cr K Na Ca Mg Al Fe S P

__________________________________________ mgL-1________________________________________

0.02 0.00 0.03 1.04 169 49.7 17.4 5.68 0.10 0.19 13.5 6.01

DOM Manure

PO4 Cl SO4 pH (H2O) Ec25 TC TOC IC C N C/N S

_____ mgL-1_____ _ µS cm-1 _ _____ mgL-1 _____ ________ g kg-1 _______

23.3 115 37.0 7.55 926 68.2 65.4 2.83 214 13.1 16.4 2.01

3.4. Adsorption experiments

Mixed solution of nitrate salts of Cu, Ni, and Zn (50 mg L-1

) were used in the

adsorption experiments. The pH of the HM solution was adjusted to pH 6 by

adding appropriate amounts of 0.1 and 0.01 M NaOH. The initial pH of the

prepared HM solution was fixed to avoid precipitation of HMs. Batches of 1 g

of air-dried soil (< 2 mm) were combined with a range of volumes (10, 40, 70,

and 100 mL) of HM solution in polypropylene tubes. Temperature was held

constant at 20°C. A preliminary study showed that the sorption equilibrium of

HMs on the soils under study was reached within 2 h (results not shown).

Therefore, this time interval was chosen for subsequent experiments. All

experiments were performed in duplicate.

In all adsorption experiments, the supernatants were passed through 0.2-µm

cellulose-acetate membrane filters, the pH was recorded immediately, and then

metals in acidified solutions were measured by ICP-OES. The amount of

adsorbed HMs was calculated by subtracting the amount of added HMs from

the amount remaining in solutions using the mass-balance equation as follows:

𝑞𝑒 = 𝑉(𝐶0 − 𝐶𝑒)/𝑀 Eq. [1.3]

where qe is the adsorbed metal ion concentration (mg kg-1

), V is the solution

volume (L), C0 is the initial concentration of metal ions (mg L-1

), Ce is the

metal-ion concentration in a bulk solution at equilibrium (mg L-1

), and M is the

44 | CHAPTER 3

adsorbent mass (g). The following four sets of adsorption experiments (A, B, C,

and D) were conducted (see also Fig. 1.3).

3.4.1. Experiment A: Heavy metals adsorption in untreated soil samples (control experiment)

To evaluate the effect of soil constituents on the adsorption of Cu, Ni and Zn,

we added 10, 40, 70, and 100 mL of the 50 mg L-1

(pH 6) stock solutions

separately to 1 g of soil in 50, 100, 200, and 250 mL volume polyethylene

tubes. The tubes were shaken for 2 h on a horizontal reciprocating shaker (130

rpm) and then centrifuged at 2012 x g for 30 min.

3.4.2. Experiment B: Heavy metals adsorption in organic matter enriched soil samples (addition

of dissolved organic matter and subsequent re-drying before addition of heavy metals)

To examine the effect of prior adsorption of DOM on the solid phase of the

soils under study on subsequent adsorption of Cu, Ni and Zn. For this, 1 g of

soil was added to 20 mL of the DOM stock solution (pH 6) in polyethylene

tubes. The tubes were shaken and then centrifuged as in Exp. A. The

supernatant was collected, filtered, and analyzed for TOC and TC using the

TOC analyzer. The residue (i.e., soil material with adsorbed OM) was

subsequently freeze-dried to obtain the OM-enriched soil material. After that,

10, 40, 70, and 100 mL of the Cu, Ni and Zn stock solutions (pH 6) were added

separately to 1 g of OM-enriched soil in 50, 100, 200, and 250 mL volume

polyethylene tubes. The tubes were shaken and then centrifuged analogously to

Exp. A.

3.4.3. Experiment C: Heavy metal sorption of combined Heavy metal-Dissolved organic matter

solutions (Dissolved organic and heavy metal added simultaneously)

To examine the effect of simultaneous addition of DOM and HMs on the

sorption of the HMs onto the soils under study; 10, 40, 70, and 100 mL of the

Cu, Ni and Zn stock solution was added separately to 1 g of soil and 20 mL of

DOM in 50, 100, 200, and 250 mL volume polyethylene tubes. The tubes were

shaken and then centrifuged analogously to Exp. A and B.

CHAPTER 3 | 45

ICP-OES instrument; TOC-Analyzer

instrument

A Cu Zn

O H H

O

H

H O H

H Ni

Soil Shaking 2 h (Shak)

Centrifuge (Cent) & Filtration (Filt) + ICP

Shak + Cent + Filt

DOM

O H

H

O H

H O

H H

O

H H

Soil

Freeze-dried

Shak + Cent

B OM-

enriched

soil Filt + ICP

OM patches on dry soil surface

Cu Zn O

H H

O

H

H O H

H Ni

Soil C

Shak + Cent + Filt

ICP + TOC

Cu Zn O

H H

O

H

H O H

H Ni

DOM

O H

H

O H

H O

H H

O

H H

Soil Shak DOM

O H

H

O H

H O

H H

O

H H

D Shak + Cent + Filt

ICP + TOC

Cu Zn O

H H

O

H

H O H

H Ni

DOM

O H H

O H

H

O H

H

O

H H

DOM O H

H O H

H

O H H

O H H

DOM O H

H

O H H

O H

H

O

H H

Adsorbed OM on soil surface

3.4.4. Experiment D: Adsorption of heavy metal on soil surfaces with prior adsorbed dissolved

organic matter (first addition of Dissolved organic matter, then addition of heavy metal,

no drying between)

To examine the effect of sequential addition of DOM and HMs on sorption of

HMs onto the soils under study. For this, 1 g of soil was added to 20 mL of

DOM stock solution in 50, 100, 200, and 250 mL volume polyethylene tubes.

The tubes were shaken for 2 h and then 10, 40, 70, and 100 mL of the Cu, Ni

and Zn stock solutions was added separately to each tube. The tubes were

shaken again for 2 h and then centrifuged as previously described. In all

previous experiments the supernatants were collected, filtered, and analyzed for

TOC, Cu, Ni, and Zn.

Fig. 1.3: Schematic overview for all adsorption experiments (A, B, C, and D).

46 | CHAPTER 3

3.5. Modeling of sorption kinetics and statistical analysis

We applied IM isotherms to describe the adsorption of HMs and DOM because,

in contrast to traditional Langmuir or Freundlich based approaches, IM

isotherms are specifically designed to describe a net release of indigenous DOM

as well (Kaiser and Guggenberger, 2000). In the IM approach (Eq. [2.3]), the

quantity of a substance adsorbed or released, RE (µmol kg-1

) is plotted against

the initial quantity of the substance added, Xi (µmol L-1

):

RE = mXi-b Eq. [2.3]

The slope m of the linear regression isotherms is interpreted as a partitioning

coefficient Kd (Nodvin et al., 1986). This Kd is a measure of the affinity (ranging

between 0 and 1) of the sorbent. From this Kd value, the mobility and fate of

competing metals in the soil can be assessed (Gao et al., 1997; Cruz-Guzman et

al., 2006). The intercept of the regression line (b) indicates the amount of

substance (µmol kg-1

) released from soil when a solution without sorbent is

added. Therefore, the intercept may be defined as a desorption term (Ussiri and

Johnson, 2004). The estimated Kd values and the correlation coefficients (r2) for

a linear regression were determined with SigmaPlot for Windows 11.0. A one-

way ANOVA followed by a least significant difference (LSD) test were

employed to determine the significance of the differences between treatments.

For this, Origin (version 8 for Windows) was used.

4. Results

4.1. General properties of the tested soils

Clay contents in the studied soils ranged from 11 to 71%. Clay minerals

consisted of smectites and kaolinites, with the former being the dominant type

(Table 2.3). Only in soil-BO smectite and kaolinite contents were almost equal.

The studied soils had large CECs ranging from 42.4 to 65.4 cmolc kg-1

(Table

2.3). Crystalline Fe-oxide contents were small to moderate (3.6-17.5 g kg-1

)

with the highest contents in soil-KW and soil-WQ, while SOC contents were

low in all soils (0.3-1.7 g kg-1

, Table 2.3). The pH was always slightly basic

(Table 2.3). The dominant exchangeable cation was Na+ while Ca

+2, Mg

+2, and

K+ provided minor contributions (Table 2.3).

CHAPTER 3 | 47

Table 2.3: Selected physical and chemical properties of soil samples

# CBD ext. (Citrate Bicarbonate Dithionite extraction) ; Kao (Kaolinite); Sme (Smectite).

Soil

name

Soil

pH

CEC

Clay Minerals Particle size CBD

ext.

Oxalate

ext.

Major cations Major anions

Sme. Kao. Clay Silt Sand SOM Fe-oxide MnO2 Na Ca Mg K Cl- SO4

-

cmolc kg-1

__________ ~% ____________ _____________________ mg kg-1

_____________________ __ mg L-1

__

Soil-KW 7.5 64.2 95 4.0 18 78 2.7 300.0 14600 3500 5439 333 112 12 6504 451

Soil-KO 7.6 65.4 87 10 18 79 1.1 1700 9700.0 1300 2085 677 211 12 3562 173

Soil-WQ 7.5 58.8 84 15 11 85 3.8 1200 17500 1200 2409 970 77.8 23 4179 91.3

Soil-AH 7.7 55.4 76 22 19 75 2.8 900.0 10600 2100 4071 541 180 12 3818 48.0

Soil-BO 7.0 42.4 52 45 71 30 0.3 700.0 3600.0 30.00 2625 68.1 75.3 59 3892 548

48 | CHAPTER 3

4.2. Adsorption isotherms and adsorption coefficient (Kd)

In all experiments and for all HMs tested, the IM isotherm model was able to fit

the data very well (R2 = 0.94-0.99, Table 3.3). The trends in the Kd values for

the three metals and five soils in all four experiments are presented in Fig. 2.3.

In general, the great majority of the added HMs was adsorbed regardless of the

treatment processes. Specifically, for the lowest HM addition, 95 to100% of the

amount of HMs added was adsorbed (i.e., 10 ml HMs solution with 50 mg L- 1

).

For the highest HM addition (i.e., 100 mL HMs solution, 50 mg L-1

), the

percentage adsorbed still ranged between 55 and 81 % of metal added.

Table 3.3: Fitted initial mass isotherm parameters for Cu, Ni, and Zn in the four batch adsorption

experiments with five different soils (mean of two replicates).

Soil name Metal Exp. A Exp. B Exp. C Exp. D

Kd b r2 Kd b r2 Kd b r2 Kd b r2

Soil-KW Cu 0.77 4344 0.99 0.72 4832 0.98 0.78 3103 0.99 0.78 3211 0.99

Ni 0.73 5121 0.98 0.68 5935 0.97 0.75 4364 0.99 0.75 4496 0.99

Zn 0.71 5605 0.98 0.67 5895 0.97 0.73 4173 0.99 0.73 4270 0.98

Soil-KO Cu 0.69 5420 0.98 0.69 4746 0.98 0.69 4378 0.99 0.72 4217 0.99

Ni 0.52 6293 0.97 0.56 5750 0.98 0.61 5130 0.99 0.61 5163 0.98

Zn 0.51 6973 0.97 0.56 5840 0.97 0.61 4855 0.98 0.60 5013 0.98

Soil-WQ Cu 0.72 4821 0.99 0.68 4808 0.98 0.67 4844 0.98 0.71 4179 0.99

Ni 0.54 5318 0.98 0.53 6066 0.97 0.56 5723 0.98 0.59 5286 0.98

Zn 0.52 6117 0.98 0.52 6216 0.97 0.55 5467 0.98 0.57 5137 0.98

Soil-AH Cu 0.62 6101 0.97 0.64 5303 0.98 0.67 4989 0.98 0.70 4377 0.99

Ni 0.54 6480 0.97 0.54 6394 0.97 0.60 5785 0.98 0.61 5551 0.98

Zn 0.53 7004 0.96 0.52 6475 0.96 0.59 5474 0.97 0.60 5262 0.98

Soil-BO Cu 0.56 7004 0.96 0.57 6068 0.96 0.62 5290 0.98 0.64 5051 0.98

Ni 0.52 7441 0.95 0.51 7000 0.95 0.58 6442 0.97 0.59 6243 0.97

Zn 0.50 7975 0.94 0.50 6901 0.95 0.57 5955 0.97 0.57 5805 0.97

CHAPTER 3 | 49

4.2.1. Interaction of untreated soils with heavy metals in the absence of dissolved organic

matter (“Control” Experiment A)

In all studied soil samples the affinity of the three HMs for the solid soil phase

as indicated by the Kd values followed the sequence: Cu >> Ni ≈ Zn (Table 3.3).

The variation in the Kd values for the tested soils (Fig. 2.3 a) reflects the

influence of the various soil properties (clay minerals, oxides, and bulk OM) in

determining HM affinity to soils. The Kd values for Cu decreased as follows:

soil-KW > WQ > KO >> AH > BO (Fig. 2.3 a). The difference between the

tested soils was statistically significant for all studied soils (p < 0.05). The Kd

values for Ni and Zn were similar for both metals and similar for all soils but

one (Fig. 2.3 b,c). The exception is soil-KW, which had a much higher affinity

for Ni and Zn than all other soils (p < 0.001). The variation among the Kd values

of the other soils for Ni and Zn was not significant (p > 0.05) (Fig. 2.3 b,c).

The results of the correlative exploration of the relation between Kd values

(affinity of Cu, Ni and Zn) and relevant soil properties, such as clay minerals

content, CEC, bulk OM, cations, and oxides, revealed a significant effect (p <

0.05) of smectite on Cu affinity (R2 = 0.90) and MnO2 on Zn affinity (R

2 Zn =

0.77).

50 | CHAPTER 3

Fig. 2.3: Comparison between distribution coefficients (Kd) for individual heavy metals in all experiments.

4.2.2. Interaction of organic matter-enriched soils with heavy metals (Experiment B)

The OM enrichment of the soils that was achieved in the preparatory step of

Exp. B was always small, with a maximum of < 1 g C kg-1

soil adsorbed.

However, the proportions of adsorbed organic C in total SOC (sum of adsorbed

and original OC) were of similar magnitude: 71, 35, 39, and 60% for Soils KW,

WQ, AH, and BO, respectively. The only exception was Soil-KO, which had

the highest original OM content (Table 2.3) and did not adsorb measurable

amounts of OM.

Overall, the interaction of the studied HMs with the OM-enriched soils as

represented by their Kd values resulted in the same selectivity sequence (Cu >>

Ni ≈ Zn) as Exp. A. However, the absolute Kd values and their ordering differed

from those found in Exp. A (Fig. 3.3 a,b,c).

CHAPTER 3 | 51

a c

b

Fig. 3.3: Relative changes in Kd values of Cu, Ni, and Zn comparing to control experiment.

For Cu the affinity for the OM-enriched soil surfaces followed the sequence:

KW > KO > WQ > AH >> BO soil. This ordering correlates with the smectite

content and CEC value of the tested soils (Table 2.3). Compared to Exp. A, the

Kd values for Cu adsorption were significantly reduced in Soil-KW and Soil-

WQ (6.5 and 5.6%, respectively; p < 0.01). In the other soils, the differences

with the control experiment were much smaller (Fig. 3.3 a).

Analogous to Exp. A, the Kd values of Ni and Zn in Exp. B were similar for

both metals in all soils tested (Fig. 3.3 b,c). Also analogous to Exp. A, the

variation in Ni and Zn affinities for the studied soils was small, with the

exception of Soil-KW, which showed the highest Kd value of all studied soils

(Fig. 3.3 b,c). Compared to Exp. A, the Kd values for Ni and Zn adsorption were

52 | CHAPTER 3

significantly reduced in Soil-KW (7% for Ni and 6% for Zn; p < 0.01) and

increased in Soil-KO (7% for Ni and 8% for Zn; p < 0.01 ). The changes were

small but significant (p < 0.01) in the other soils (Fig. 3.3 b,c).

4.2.3. Interaction of untreated soils with simultaneously added heavy metals and dissolved

organic matter (Experiment C)

In these experiments where HMs and DOM were added simultaneously, a

positive correlation was found between the absolute amount of HMs added and

the absolute amount of DOM adsorbed. The amount of DOM adsorbed ranged

from 0.5 to 1.2 g C kg-1

soil (Fig. 4.3). A significant difference was observed in

the amount of adsorbed DOM between the 10 and 40 mL HM additions (P <

0.001) and the 40 and 70 mL HM additions (p < 0.05). The difference between

the amount of DOM adsorbed after the 70 mL HM addition and the 100 mL

HM addition was not significant (p > 0.05). No significant difference was

observed in the amount of DOM adsorbed and the type of HM added (p > 0.05).

As in Exp. A and B, the Kd values for the three metals followed the sequence:

Cu >> Ni ≈ Zn (Table 3.3). For Cu, the sequence in Kd values was similar to

that in Exp. B and followed the content of smectite and CEC values in tested

soils (Fig. 3.3 a; Table 2.3). However, the differences between soils varied

compared to Exp. B (Fig. 3.3). The highest relative increase in the Kd value for

Cu affinity compared to the control (Exp. A) was registered in Soil-AH (6%)

and Soil-BO (10%). However, underlying the Kd value was a change in the

shape of the affinity curve, with the amount of Cu immobilized at the two

lowest additions (10 and 40 mL) being higher in Exp. A, but the amount of Cu

immobilized at the two highest additions (70 and 100 mL) being higher in Exp.

C. As a result the difference in affinity for the metals in Soil-AH and Soil-BO

proved not statistically significant. In Soil-KW and Soil-KO (Fig. 3.3 a) the

affinity for Cu also increased. While the increase was smaller than in Soil-AH

and Soil-BO, it was statistically significant (p < 0.01). In Soil-WQ the affinity

for Cu significantly decreased (7%; p < 0.01).

The affinity sequence for Ni and Zn in Exp. C was similar as in the previous

experiments (Soil-KW >>> KO > AH > BO > WQ). In all cases the Kd values

CHAPTER 3 | 53

for Ni and Zn increased with respect to the control experiment (Exp. A). For Ni

the increase was significant (p < 0.05) for Soils KO, BO, AH, and WQ

(respectively 15.8, 12.0, 9.9, and 4.1%, respectively). No significant change was

found with respect to the Kd values for Ni in Soil-KW (p > 0.05) (Fig. 3.3 b).

For Zn the increase in the Kd values was significant (p < 0.05) for Soils KO,

BO, AH, WQ, and KW, and was 17.7, 13.2, 10.8, 5.9, and 2.9% , respectively,

(Fig. 3.3 c).

Fig. 4.3: Dissolved organic matter (DOM) adsorption by the studied soils in Exp. C and D based on heavy

metals addition.

4.2.4. Interaction of soils with dissolved organic matter and heavy metals added sequentially

(Experiment D)

As in Exp. C, increasing HM additions correlated with increasing adsorption of

DOM still present in solution (i.e. not yet adsorbed after 2 h of shaking; Fig.

4.3). However, the amount of DOM adsorbed was larger in all studied soils in

Exp. D (0.3-20%) than in Exp. C (Fig. 4.3).

54 | CHAPTER 3

Interactions of DOM and HMs did not change the general pattern of larger Kd

values for Cu than for Zn and Ni. Also, as in all previous experiments, Soil-KW

still had the highest affinity to the tested metals (Fig. 2.3).

The Kd values for Cu in Exp. D were larger than those in all other Experiments

(Exp. A, B, and C) with the exception of the soils that are rich in Fe-(hydr)oxide

(Soil-KW and Soil-WQ) (Fig. 3.3 a). In Soil-KW, Cu had nearly the same Kd

values in Exp. A, C, and D, but the differences between them were significant

(p < 0.01). Also in Soil-WQ the differences in affinity were small but

significant (p < 0.01; p < 0.05) (Fig. 3.3 a). The largest relative increase in the

Kd value compared to the control (Exp. A) was registered in Soil-AH (12.5%)

and Soil-BO (14.2%) (Fig. 3.3 a). However, the larger Kd value was derived

from the higher amounts of Cu immobilized at the two highest additions (70 and

100 mL) in Exp. D than in Exp. A. The amount of Cu immobilized at the two

lowest additions (10 and 40 mL) was lower in Exp. D than in Exp. A, but not

enough to result in an overall lower Kd value. As such, while appreciable, the

overall difference in Kd value between Exp. A and D were not statistically

significant.

The adsorption of Ni and Zn was enhanced in Exp. D compared to Exp. A and

B, and was similar to that in Exp. C (Fig. 2.3 b,c). Compared to Exp. A, the Kd

values for Ni were significantly enhanced (p < 0.05) in Soils KO, WQ, AH, and

BO by 15.4, 7.7, 11.7, and 13.8%, respectively (Fig. 3.3 b). No significant

change was found in Soil-KW (p > 0.05). The affinity of Zn was significantly

enhanced (p < 0.001) in Soils KW, KO, WQ, AH, and BO by 3, 16.5, 9.7%,

12.8, and 15.2%, respectively (Fig. 3.3 c).

4. Discussion

4.1. Overview of adsorption behavior of heavy metals in the soils tested

As expected, the adsorption behavior of the studied HMs was greatly influenced

by soil properties (clay minerals; iron oxides; SOM originally present). Na+ was

the main exchangeable cation (Table 2.3) present in the studied soils, and

according to the lyotropic series has a much lower affinity for clay surface than

the multicharged HMs under study (Bohn et al., 1985; Essington, 2004). This

CHAPTER 3 | 55

explains the large net absorption of all HMs studied under all treatments. In

addition, both the presence of DOM as well as the timing of its addition affected

the adsorption behavior of the HMs under study. An overview of the combined

effects of the various treatments on the affinity of Cu, Ni and Zn for the solid

phase in the soils tested is presented in Table 4.3. In general, there was a distinct

division in behavior between Cu on the one hand and Ni and Zn on the other

(Table 4.3).

Table 4.3: Summary of the results regarding the Kd values of the four different experiments.

Details of the four experiments are given in Fig. 1.3.

Experiment Cu affinity Ni and Zn affinity

Exp. A

Higher in Fe-oxide and smectite-rich

soils (Soil-KW and Soil-WQ).

Higher in smectite-rich soils (Soil-KW).

Exp. B Compared to Exp. A: reduced in Fe-

oxide-rich soils (Soil-KW and Soil-

WQ), slightly increased in Fe-oxide

poor soils (Soil-AH and Soil-BO), and

no change in the SOM-rich soil (Soil-

KO).

Compared to Exp. A: enhanced in the SOM-

rich soil (Soil-KO) and reduced in the other

soils (particularly Soil-KW, Fe-oxide and Mn-

oxide-rich soil).

Exp. C Compared to Exp. A and B: Enhanced

in Fe-oxide-poor soils (Soil-AH and

Soil-BO), slightly enhanced in the

SOM-rich soil (Soil-KO), and in the

soil with high smectite content (Soil-

KW). Reduced in the Fe-oxide-rich

soil (Soil-WQ).

Compared to Exp. A and B: enhanced in all

studied soils.

Exp. D Comparing to Exp. A, B and C:

enhanced in Fe-oxide-poor soils (Soil-

AH and Soil-BO) and the SOM-rich

soil (Soil-KO). In the smectite-rich

soil (Soil-KW) slightly enhanced

compared to Exp. A and C and

enhanced compared to Exp. B.

Compared to Exp. A, B, and C: enhanced in

soils with low smectite content and low CEC

values (Soil-WQ, Soil-AH, and Soil-BO).

Compared to Exp. A and B also strongly

enhanced in soils with high smectite content

and high CEC values (Soil-KW and Soil-KO),

but equal to Exp. C.

56 | CHAPTER 3

4.2. Adsorption of copper, Nickel, and Zinc in relation to soil composition and

(timing of) dissolved organic matter addition

4.2.1. The influence of inherent differences in soil composition on heavy metals adsorption

(Experiment A)

The preferential adsorption of Cu over Ni and Zn (Table 3.3) was most likely

caused by differences in binding type. It is well known that Cu adsorption

depends on covalent interactions with the mineral structure while Ni and Zn are

predominantly retained through electrostatic interactions with exchange sites in

soils (Gomes et al., 2001; Covelo et al., 2004a). Overall similarities in the

adsorption affinities of Ni and Zn are in line with previous work (Anderson and

Christensen, 1988; Gomes et al., 2001).

Also the observed significant relationship (p < 0.05) between Cu adsorption and

smectite content is in agreement with previous studies (e.g. Gomes et al., 2001).

A large cation content has been reported to result in rapid exchange with Cu on

the external clay minerals surfaces followed by a slow reaction in which Cu

ions diffuse into the inter-layer of smectite minerals (Al-Qunaibit et al., 2004).

In addition, Cu is known to form strong inner-sphere complexes with the

surfaces of Fe-(hydr)oxide, although this correlation was not significant in our

experiments. The presence of Fe-(hydr)oxide also increased the total surface

area of the soils (Feller et al., 1992; Peacock and Sherman, 2004). All this is

reflected in the large Kd values, and thus large affinity of Cu for Soil-KW and

Soil-WQ that had the highest smectite and Fe-oxide contents, respectively,

whereas the lowest calculated Kd value was found in Soil-BO due to its lowest

smectite and Fe-oxide contents of all soils tested (Fig. 2.3 a).

As expected smectite contents proved important for determining the affinity of

Zn and Ni to the solid phase as well, given the large permanent negative charge

of this family of clay minerals. In contrast to Cu, MnO2 contents also seem to

have played an important role in enhancing the affinities for both Ni and Zn

(e.g., Sheng et al., 2011). As a result, Soil-KW with the highest smectite and

MnO2 content showed the highest affinity to Ni and Zn (Tables 1.3 and 2.3).

CHAPTER 3 | 57

In general, the inherently present SOM (Exp. A) had a weak effect on the

adsorption of all HMs. Most likely, this was due to the absence of organic metal

complexes due to low initial OM contents (Table 2.3) (Gao et al., 1997).

4.2.2. The influence of prior organic matter enrichment of the soils (Experiment B)

The successful enrichment of all soils, except Soil-KO, with OM was related to

the presence of smectite (Wang and Xing, 2005), Al and Fe oxides, and

hydroxides (McKnight et al., 1992; Kaiser et al., 1996). Soil-KO did not adsorb

measurable amounts of DOM, in spite of it containing large amounts of Al and

Fe-(hydr)oxide, because this soil already had by far the largest inherent SOM

content (Table 2.3). As a result, available sorption sites on the Al and Fe-

(hydr)oxides were already saturated with OM (Kaiser et al., 1996; Kaiser and

Zech, 1998). This is further supported by the observed inverse relationships

between native SOM contents (KO > WQ > AH > BO > KW) and the

proportions of OM adsorbed during OM-enrichment (KW > BO > AH > WQ >

KO) (Table 2.3).

The decreased affinity of Cu for the solid phase of most soils after their

enrichment with OM as compared to the control experiment (Fig. 3.3) indicates

a dominant role of the mineral phase in Cu binding. This is in line with its

expected bonding through inner-sphere complexes, and with findings of Lair et

al. (2007), who showed that even in the presence of SOM, Cu is preferentially

bound by the mineral phase. Specifically, the decline in Cu affinity in the Fe-

(hydr)oxide and smectite rich soils (Soil-KW and Soil-WQ; Table 2.3) could be

explained by the blockage by OM coating of specific binding sites on the Fe-

(hydr)oxides (Feller et al., 1992; Kaiser and Guggenberger, 2000) and of the

interlamellar spaces and/or specific inner binding sites of smectite (Zhuang and

Yu, 2002) after enrichment with OM. Apparently, on its adsorption, the OM did

not adsorb enough Cu itself to counter the reduced affinity of the mineral phase.

The lack of an appreciable effect in the other soils might be explained by

smaller contents of Fe-(hydr)oxide in these soils also indicating an important

role of Fe-(hydr)oxide for adsorption of Cu.

Analogous to Cu, the largest reduction of Ni and Zn affinity to the solid phase

was observed for Soil-KW and can be attributed to blockage of active sites on

58 | CHAPTER 3

the mineral phase by OM coating (Fig. 3.3). This soil had the highest smectite

content (Table 2.3) indicating binding to smectite as a dominant adsorption

mechanism for Ni and Zn, albeit most likely through electrostatic interactions

instead of inner-sphere complexation. Because Fe-(hydr)oxides do not have

large permanent negative charge like smectite, electrostatic interactions will

have been much smaller, explaining the smaller reduction of Ni and Zn affinity

by OM compared to Cu in Soil-WQ (Fig. 3.3). The statistically significant

increase in Ni and Zn affinity in Soil-KO on OM enrichment was not expected

given that the SOM content of this soil did not increase significantly on OM

enrichment. A possible explanation is that, while the absolute amount of OM

did not increase, its molecular composition and/or steric configuration might

have resulted in an increase in specific binding sites on the SOM present (Liu

and Gonzales, 1999).

4.2.3. The effect of the timing of dissolved organic matter on heavy metals adsorption

(Experiments C and D)

The affinity of the solid phase for Cu in Soil-AH and Soil-BO was significantly

enhanced on concurrent addition with DOM (Exp. C), and even more on

sequential addition (Exp. D; Fig. 3.3). This indicates that Cu was adsorbed on

DOM that itself was previously immobilized on specific solid phase sites not

directly involved in the binding of HMs. In addition, Cu was probably

immobilized concurrently with DOM through cation bridging with the abundant

smectite in these soils. The fact that the affinity of Cu was not enhanced in the

oxides-rich soils (KW, WQ) suggests that the DOM in our experiments coated

the specific binding sites for Cu on the solid phase. The resulting blocking of

binding sites counteracted a potentially enhanced affinity through the previously

mentioned mechanisms. In addition, in this case binding sites on DOM for Cu

may have been occupied by Fe that also forms strong inner-sphere complexes

with DOM (Senesi et al., 1986), thereby reducing the occurrence of both

mechanisms.

For Ni and Zn, the increased affinity for binding on the solid phase on OM

addition was much smaller than for Cu and always increased regardless of the

timing (Fig. 3.3). This confirms that electrostatic binding mechanisms dominate

for these two metals, resulting in weaker association with DOM. The overall

CHAPTER 3 | 59

affinity to Ni and Zn remained the highest in Soil-KW, which correlates to large

surface area of its abundant smectite contents even in the face of adsorbed OM.

In fact, several authors suggest that interactions between smectite and OM may

promote Ni and Zn adsorption (Wattel-Koekkoek et al., 2003; Feng et al.,

2005).

5. Conclusion

Our experiments confirmed the influence of the amount and timing of DOM

addition on the affinity of Cu, Zn and Ni for the solid phase in the tested soils.

The results suggest that Cu was mostly bound through inner-sphere complexes

on smectite and Fe-(hydr)oxides. Further, we found that association in the case

of smectite was enhanced by inner-sphere complexation with DOM bound to

the solid phase directly and through cation bridges. As a result, in smectite- rich

soils, sequential addition of DOM and Cu resulted in a higher affinity for the

solid phase than concurrent addition. In Fe-(hydr)oxides rich soils, the enhanced

affinity by DOM addition was counteracted by coating of binding sites on the

Fe-(hydr)oxides by OM.

Nickel and Zn were found to bind predominantly through electrostatic

interactions. As a result, overall affinity for the solid phase was lower than for

Cu. Furthermore, the addition of DOM resulted in smaller increase in affinity

than for Cu, and the timing of the addition (concurrent with the metals or

sequential) had a much smaller effect. As such, our study points to interesting

differences in the influence of DOM addition on the retention of Cu, Ni and Zn

in clay-rich soils that warrant further investigation.

Nevertheless, regardless of the differences found, in all experiments, for all

metals and in all soils tested, the great majority of the HMs in the system were

adsorbed to the solid phase. This has important implications for the use of

natural soil material for wastewater treatment. It means that readily available

and abundant natural soil material can be used as a cheap and effective way of

removing a large percentage of Ni, Cu and Zn from wastewater, regardless of

whether the water is rich in DOM, and regardless of whether DOM had

previously been added.

60 |

CHAPTER 4 | 61

Chapter 4

Effects of clay minerals, hydroxides, and timing of dissolved

organic matter addition on the competitive sorption of Copper,

Nickel and Zinc: A column experiment

Abstract

Infiltration of heavy metal (HM) polluted wastewater can seriously compromise soil and

groundwater quality. Interactions between mineral soil components (e.g. clay minerals)

and dissolved organic matter (DOM) play a crucial role in determining HM mobility in

soils. In this study, the influence of the timing of addition of DOM, i.e. concurrent with

or prior to HMs, on HM mobility was explored in a set of continuous flow column

experiments using well defined natural soil samples amended with goethite, birnessite

and/or smectite. The soils were subjected to concurrent and sequential additions of

solutions of DOM, and Cu, Ni and Zn. The resulting breakthrough curves were fitted

with a modified dose-response model to obtain the adsorption capacity (q0). Addition of

DOM prior to HMs moderately enhanced q0 of Cu (8-25%) compared to a control

without DOM, except for the goethite amended soil that exhibited a 10% reduction due

to the blocking of binding sites. Meanwhile, for both Zn and Ni sequential addition of

DOM reduced q0 by 1-36% for all tested soils due to preferential binding of Zn and Ni

to mineral phases. In contrast, concurrent addition of DOM and HMs resulted in a

strong increase of q0 for all tested metals and all tested soil compositions compared to

the control: 141-299% for Cu, 29-102% for Zn and 32-144% for Ni. Our study shows

that when assessing the impact of soil pollution through HM containing wastewater it is

crucial to take into account the presence of DOM.

This chapter is based on: Refaey, Y., Jansen, B., Parsons, J., de Voogt, P., Bagnis, S., Markus, A., El-Shater,

A., El-Haddad, A., Kalbitz, K. 2016. Effects of clay minerals, hydroxides, and timing of dissolved organic

matter addition on the competitive sorption of Copper, Nickel and Zinc: A column experiment. (Revised version: Journal of Environmental Management).

62 | CHAPTER 4

1. Introduction

Heavy metals (HMs) are considered potentially highly toxic pollutants and may

pose a serious threat to environmental quality (e.g. Qin, 2006; Usman, 2008).

Soil contamination with HMs may occur due to irrigation with contaminated

water, the addition of fertilizers and metal based pesticides. Such problems are

especially acute in arid developing countries such as Egypt where wastewater

reuse could be a reasonable choice to mitigate the shortage and scarcity of fresh

water resources (Radwan and Salama, 2006; Alfarra et al., 2011).

The fate and transport of HMs in soils and subsequently in groundwater

aquifers are mainly controlled by the sorption capacity of soil constituents and

aquifer materials (Alloway, 1995; McBride et al., 1997, 1999). The adsorption

of HMs onto solid phases such as clay minerals (Al-Qunaibit et al., 2004;

Refaey et al., 2014; Colombani et al., 2015), Fe- and Mn-hydroxides (Cavallaro

and McBride, 1984; Elliott et al., 1986; Stahl and James, 1991) and organic

matter (Kalbitz and Wennrich, 1998) is the most important chemical process

regulating the mobility of HMs in the environment (Antoniadis et al., 2007a). In

addition, given their high adsorption capacity and specific surface area (SSA),

clay minerals are widely used as low-cost agent to remove HMs from

wastewaters (e.g., Ikhsan et al., 2005; Refaey et al., 2015). Nano-sized oxides of

Mn and Fe act as important scavengers for contaminants in soils and have been

successfully used for the removal of different HMs from wastewater owing to

their high reactivity and large SSA (Klaine et al., 2008; Hashim et al.,

2011; Tang and Lo, 2013).

Dissolved organic matter (DOM) is often present in considerable concentrations

either in the wastewater itself (e.g., industrial and agricultural effluents) or in

the soil (e.g. due to manuring). Such presence of DOM can exert a significant

influence on the fate and transport of HMs in soil. Sorption of DOM to mineral

surfaces is considered an important pathway for the retention and also the

stabilization of OM (e.g., Kaiser and Guggenberger, 2000; Kalbitz et al., 2005;

Mikutta et al., 2007). Therefore, the influence of DOM on HM mobility not

only concerns the interaction of DOM with HMs, but also processes altering the

mobility of the DOM itself in the soil. Understanding the mechanisms

controlling the interactions of HMs with both mineral surfaces and DOM is

CHAPTER 4 | 63

therefore essential to get insight into transport and fate of metals in soils

(Arshad et al., 2008; Cecchi et al., 2008). DOM was previously found to either

hinder or promote HM adsorption to mineral surfaces in soils depending on the

affinity of metal-ligand complexes for adsorbents (Kalbitz and Wennrich, 1998;

Shuman et al., 2002; Jansen et al., 2003; Refaey et al., 2014).

A previous study employing batch experiments demonstrated that the timing of

the addition of DOM to soils, i.e. concurrent with or sequential to HM addition,

may play a role in regulating the mobility of HMs in soils (Refaey et al., 2014).

However, this batch approach only provided a snapshot at a particular liquid to

solid ratio and is unsuitable for capturing the dynamics of a realistic soil system

where flow kinetics should be taken into account (Maszkowska et al., 2013). A

column approach enables time-dependent monitoring of contaminant leaching

from soil and waste materials; in addition, the flow-through pattern of such tests

resembles actual environmental conditions (Maszkowska et al., 2013).

The objectives of the present study were (1) to unravel the effect of the timing

of the addition of DOM on the competitive adsorption of Cu, Ni and Zn onto

different soil compositions in a kinetic system, (2) to quantify the fate and

transport of metals in different mineral surfaces as well as gain insights into

leaching behavior under actual environmental conditions. For this a column

approach was used in order to accommodate the dynamic characteristics of

metals interaction with DOM and soil minerals.

2. Materials and Methods

2.1. Sampling area and soil selection

Soils were sampled in Southern Limburg, The Netherlands at 50° 53' 37.61" N

lat; 5° 53' 34.56" E long. The samples were collected from the C horizons

according to their variations in soil composition (e.g. grain size, color, and

organic material) after removing the A horizons on the top (up to 45 cm depth).

The A horizon has a silty loam texture and is characterized by angular blocks,

more sticky, brown to dark grey, and clay-rich. The A horizon has a sharp

boundary to the B horizon. The B horizon (up to 75 cm depth) is yellowish

brown, soft, silty, burrows traces, less blocky, less permeable, and contains

64 | CHAPTER 4

brownish rich materials. The B horizon changes gradually downward to the C

horizon. The C horizon is rich in Fe-oxide, has a silty loam texture, and low

clay fraction content (28%).

The soil was selected based on a previous application in a different study

focusing on soil-water interactions. Selection criteria were that the C horizon is

of uniform grain size texture (silty texture), lacks native HMs, and is poor in

native OM content. Samples of the C horizon were therefore used in the current

column approach. The C horizon was extensively characterized before its

application in the present study (see 2.2).

2.2. Physico-chemical and mineralogical characteristics of the studied soil

Field water content was determined by drying soil samples at 105°C for 24 h.

The soil pHH2O was also measured (1:2.5 ratio). The pedotransfer functions

(PTFs) method was used to estimate the hydraulic conductivity (K) in tested

soils (Wösten, et al., 1999). The cation exchange capacity (CEC) of soils was

measured using the method of Hendershot and Duquette (1986). Major cations

(Ca2+

, Mg2+

, K+, and Na

+) and major anions (Cl

− and SO4

2−) were measured

using inductively coupled plasma optical emission spectrometry ICP-OES

(PerkinElmer-Optima 3000XL) and San++

Automated Wet Chemistry Analyzer-

Continuous Flow Analyzer (Skalar), respectively. Total carbon (TC) content

was determined with a C/N analyzer (Elementar Vario EL). Total content of

pedogenic (hydr)oxides was estimated as dithionite-citrate-bicarbonate

extractable iron (Fed) (AAS, PerkinElmer) using the method of Mehra and

Jackson (1960). Mn-oxide and active (oxalate extractable) Fe- and Al-

(hydr)oxide (Feo and Alo) contents were measured using the method of Searle

and Daly (1977). For mineralogical identification, X-ray diffraction analysis

was performed using a Philips (now PANalytical) PW 1830 instrument, with a

Philips PW 3710 control unit (Cu Ka radiation with wavelength 1.54056 Å

produced at 50 mA and 40 kV) to identify the clay minerals present in the clay

fraction (Brindley and Brown, 1980; Refaey et al., 2015).

CHAPTER 4 | 65

2.3. Column experiments

2.3.1. HMs solution, DOM solution, and soil preparation

Mixed solutions of chloride salts of Cu, Ni, and Zn ( 25 mg L−1

) were used.

DOM was prepared by aqueous extraction from soil with added natural manure

following the method described by Refaey et al. (2014). The same source of

DOM was used in our previous and current studies for reasons of comparison

(Refaey et al., 2014). The HM and DOM solutions were adjusted to pH 6 before

starting the experiments by adding appropriate amounts of 0.1 and 0.01 M

NaOH to avoid precipitation of HMs and DOM.

Previously, smectite, goethite and birnessite were found to play a prominent

role in regulating the binding affinity of Cu, Ni, and Zn to soil (Refaey et al.,

2014). Therefore, in our current study the original soil was amended with these

three minerals. Na-smectite (montmorillonite) of Wyoming (SWy-2) was

obtained from the Source Clays Repository of The Clay Minerals Society, West

Lafayette, USA; SWy-2 Na-rich Montmorillonite, Crook County, Wyoming,

USA. Goethite (α-FeOOH) was synthesized according to the method of

Schwertmann and Cornell (1991). Birnessite (δ-MnO2) was synthesized

according to the method of Händel et al. (2013).

The mineral amendments were used to create five different soil compositions as

described in Table 1.4. Each was mixed with 5.00 g of sand (50-70 mesh

particle size, SiO2, Sigma-Aldrich) to increase the hydrological conductivity of

the soil once packed in the column, so the final weight of each prepared soil was

10.0 g (Table 1.4).

Table 1.4: Composition and hydraulic conductivity (K) of the tested soils.

Soil compositions Original soil

(g)

SiO2

(g)

Smectite

(g)

Goethite

(g)

Birnessite

(g) K (m/s)

Soil-control 5.00 5.00 - - - 3.45x10-7

Soil-smectite 4.50 5.00 0.50 - - 2.84x10-7 Soil-goethite 4.92 5.00 - 0.08 - 3.28x10-7

Soil-birnessite 4.80 5.00 - - 0.20 3.29x10-7

Soil-smectite-oxides 4.22 5.00 0.50 0.08 0.20 2.99x10-7

66 | CHAPTER 4

2.3.2. Experimental set-up

Continuous flow sorption experiments were conducted in 12 cm high and 2.5

cm internal diameter glass columns (Glasinstrumentmakerij, FNWI, University

of Amsterdam). 10 g of each prepared soil was placed in the column to yield the

desired bed height (Fig. 1.4).

Soils were packed in the columns by a series of additions in thin layers.

Additionally, two sand layers of 35 g each were used to guarantee consistent

flux through the soil bed. The soil sample was retained in the column by means

of adaptors on the top and bottom of the column containing two paper filters. To

further stabilize the soil bed, a layer of glass wool of 3 g was placed on top of

the upper sandy filter (Fig. 1.4). To prevent preferential flow-paths and for

precise control of the flow rate, the HMs and DOM solutions were pumped

upwards against gravity by means of peristaltic pumps (Minipuls 3, Gilson).

The flow rate was set at 0.333 ml/min. To saturate the soil sample and eliminate

air bubbles, demineralized water was pumped for 12 h prior to the experiment.

Fig. 1.4: Schematic diagram of single column apparatus

CHAPTER 4 | 67

The experiments consisted of the following three scenarios (A, B and C);

conducted on each of the five soil compositions (see 2.3.1) and carried out in

quadruplicate:

(i) Scenario A (Control experiment): A total amount of ~ 9.41 L of HMs

solution without DOM was pumped through all soil columns (~ 0.47 L for

each soil column).

(ii) Scenario B (DOM and HMs added sequentially): The soil columns were

first enriched with DOM by continuous pumping of DOM solution (170 mg

C/l) overnight (~ 5.00 L; ~ 250 ml for each soil column), followed by ~

10.50 L of HM solution (~ 0.53 L for each column).

(iii) Scenario C (DOM and HMs added simultaneously): ~ 13.73 L of a solution

containing DOM and HMs solution was pumped through the soil columns

(~ 0.70 L for each column).

2.3.3. Analysis

Dissolved organic carbon (DOC) was determined by a TOC analyzer (TOC-

VCPH, Shimadzu, Kyoto, Japan) while TOC contents in solid OM were

determined with a C/N analyzer (Elementar Vario EL). Ultraviolet absorbance

(UVA) was measured at λ=254 in effluents with a UV-Vis spectrometer

(Spectroquant Pharo 300, Merck). Specific ultraviolet absorbance (SUVA)

values for each leached sample were obtained by dividing the UV absorbance

value by the DOC concentration (mg/l) in the leachate and reported in the units

of liter per milligram carbon per meter (L mg-1

m-1

). SUVA is related to the

average molecular weight of the DOM and provides a rough estimation of the

aromaticity per unit of carbon concentration (Weishaar et al., 2003; Piirso et al.,

2012). Effluent samples (35 ml) were collected from the exit of the column at

different intervals for a total time of 18 h and analyzed for Cu, Ni, Zn, Fe, Mn

and cations (Ca2+

, Mg2+

, K+, and Na

+) using ICP-OES (PerkinElmer-Optima

3000XL).

2.3.4. Modeling of adsorption of heavy metals

A variety of mathematical models have been used recently instead of

experimental determination for simulation of breakthrough curves (BTCs) data

68 | CHAPTER 4

and prediction of parameters such as the capacity of adsorbent (Meng et al.,

2012; Yi et al., 2012). The obtained data are presented in the form of BTCs

which in turn were analyzed using the modified dose-response model (Araneda

et al., 2011).

Yan et al. (2001) used the modified dose-response model to more adequately

describe the breakthrough data than the Bohart-Adams and Thomas models

(Araneda et al., 2011; Xu et al., 2013). In the current study linear regressions of

the modified dose-response model by Yan et al. (2001) were performed in order

to simulate the BTCs [Eq. (1.4)].

𝑙𝑛(𝐶𝑡/𝐶0 − 𝐶𝑡) = 𝑎 𝑙𝑛(𝑡) + 𝑎 𝑙𝑛 (𝐶0𝑄/𝑞0𝑋) Eq. 1.4

where Ct is the concentration of HM in the effluent, C0 is the concentration of

HM in the influent (mg/L), a is the modified dose-response model constant, t is

time (min), Q is the flow rate (L/min), q0 is the sorption capacity per unit mass

of adsorbent (mg/g) and X is the mass of adsorbent (g). The values of a and q0

were derived from the plot of ln[Ct/(C0-Ct)] against ln(t). The q0 of adsorbents

for toxic HMs is generally seen as an important indicator for the environmental

hazards HMs in the environment (Silveira et al., 2003).

3. Results

3.1. General properties of the tested soils

Table 2.4 summarizes the main physicochemical and mineralogical

characteristics of the studied soil. It had a moderate CEC value (23.4 cmolc kg-

1), low soil organic carbon (2.5 g kg

-1) and moderate crystalline Fe-oxide

content (15.9 g kg-1

) (Table 2.4). The particle-size distribution was dominated

by silt (67%) with a moderate contribution of clay (28%) and a minor

contribution of the sand fractions (5%) (Table 2.4). Illite, kaolinite, and smectite

were the most common clay minerals in the studied soil (Table 2.4). The clay

fraction was dominated by illite (65%) with a moderate amount of kaolinite

(32%) and minor amount of smectite (3%) (Table 2.4). The dominant

exchangeable cation was Ca2+

, while Na+, Mg

2+, and K

+ provided minor

contributions (Table 2.4).

CHAPTER 4 | 69

Table 2.4: Selected physico-chemical properties of the studied soil (C-horizon).

CEC

______

SOC

____

CBD

____

Oxalate

______

H2O

___

Clay minerals

_________________

Clay fractions

_________________

Major cations

________________________

Major anions

___________

Fe2O3 Mn2O Illi. Kao. Sme. Clay Silt Sand Ca+2 Mg+2 K+ Na+ Cl- SO42-

cmolc/kg ________ g/kg ________ ______________________ % __________________ _________ mg/l _________ ___ µmol/l__

23.40 2.50 15.91 0.77 1.28 65 32 3 28 67 5 5.63 1.02 0.51 0.20 409 127

# SOC-Soil organic carbon; CBD-Citrate bicarbonate dithionite extraction; Illi.-Illite; Kao.-Kaolinite; Sme.-Smectite.

70 | CHAPTER 4

3.2. Outflow concentrations of major cations and DOC

Figure 2.4 presents the concentrations of the major cations (Ca2+

, K+ and Na

+) in

the column outflows. In all scenarios the major cation concentrations varied

depending on the soil composition (Fig. 2.4). Soil-smectite-oxides and soil-

smectite released the largest amount of Na+ while the largest amounts of K

+ and

Ca2+

were released from soil-birnessite and soil-control (Fig. 2.4 a, b, c).

Figure 3 presents the DOC concentrations as well as the UV254 values in the

column outflows. Low DOC concentrations were recorded in the effluents in

scenario A compared to those in scenario B and C (Fig. 3.4 a). The UV254

values generally followed the trend, C >> B > A in all scenarios (Fig. 3.4 b).

The average SUVA254 values of DOM in the effluents were 6.46, 4.55, and 4.41

L mg-1

m-1

in scenario A, B, and C respectively (Fig. 3.4). The largest OC

leaching was from soil-goethite and soil-smectite in scenario B and C

respectively (Fig 3.4 a).

CHAPTER 4 | 71

a

d

b

c

e

e

72 | CHAPTER 4

Fig. 2.4: Effluent concentrations of Ca2+

, K+ and Na

+ (mg/l) at different soil compositions in experiments A, B and C.

g

h

i

CHAPTER 4 | 73

Fig. 3.4: Concentration (a) and UV (254 nm) absorbance (b) of the DOC leachate in experiments A, B and C.

3.3. Sorption capacities of metals

Figure 4.4 presents the BTCs determined in the various scenarios. The BTCs

show that Cu took the most time to reach breakthrough in all three scenarios

compared to Zn and Ni (Fig. 4.4). This indicates a higher removal capacity for

Cu than Zn and Ni (Fig. 4.4). The breakthrough point of both Zn and Ni was

much closer to each other (Fig. 4.4). The time intervals in BTCs between Cu on

b

a

C

B

A

74 | CHAPTER 4

one side and Zn and Ni on the other side was higher in scenario C compared to

scenario B and A (Fig. 4.4).

The q0 values (mg HM/g soil) for all experiments are presented in Table 3.4.

The adsorption capacity generally followed the trend, Cu > Zn > Ni (Table 3.4).

Both Zn and Ni cations displayed quite similar q0 values in the studied soils;

however, q0 for Zn was relatively higher than that of Ni. The differences in q0 of

tested soils were reported to indicate the goodness of fit of the model (Table

3.4).

Fig. 4.4: Selected breakthrough curves (BTCs) of Cu, Zn, and Ni in experiments A, B and C.

a b

c

f

d

e

CHAPTER 4 | 75

Table 3.4: Estimated parameters of modified dose-response model for the sorption of Cu, Zn and Ni.

Experiment Soil Cu Zn Ni

(Scenario) q0 (mg/g) a r2 q0 (mg/g) a r

2 q0 (mg/g) a r

2

Exp. A

Soil-control 1.13 3.06 0.85 1.00 2.51 0.86 0.80 3.32 0.82

Soil-smectite 0.82 2.44 0.82 0.85 2.54 0.79 0.66 2.94 0.80

Soil-goethite 1.12 2.47 0.75 1.07 2.15 0.72 0.78 2.44 0.71

Soil-birnessite 1.22 2.43 0.88 0.96 2.37 0.88 0.71 2.80 0.87

Soil-smectite-oxides 0.88 2.68 0.73 0.72 2.69 0.76 0.59 2.90 0.77

Exp. B

Soil-control 1.22 2.42 0.78 0.85 2.20 0.82 0.78 2.44 0.83

Soil-smectite 1.03 2.10 0.84 0.69 1.82 0.77 0.60 1.91 0.82

Soil-goethite 1.01 2.09 0.83 0.69 2.03 0.86 0.61 2.09 0.88

Soil-birnessite 1.36 2.32 0.74 0.81 2.40 0.83 0.70 2.52 0.84

Soil-smectite-oxides 1.08 2.32 0.74 0.64 2.09 0.71 0.60 2.13 0.76

Exp. C

Soil-control 3.34 1.14 0.83 1.68 1.61 0.80 1.41 1.67 0.76

Soil-smectite 2.17 1.32 0.90 1.00 1.93 0.91 0.87 2.00 0.91

Soil-geothite 3.01 1.36 0.84 2.17 1.52 0.78 1.89 1.64 0.76

Soil-birnessite 4.87 1.58 0.90 1.82 1.93 0.85 1.57 1.94 0.83

Soil-smectite-oxides 2.12 1.85 0.82 0.93 2.35 0.84 0.87 2.26 0.81

76 | CHAPTER 4

3.3.1. Scenario A (control experiment)

Figure 5.4 presents the different sorption capacities for the three metals in the

tested experimental situations, as well as the relative differences therein

between the different scenarios tested. The variation in the q0 values of the

tested soils reflected the influence of the various soil compositions in

determining the HM sorption to these soils (Fig. 5.4). Soil-birnessite showed the

highest q0 for Cu while soil-goethite and soil-control showed the highest q0

values for Zn and Ni (Fig. 5.4). Both soil-smectite and soil-smectite-oxides

showed the lowest q0 values for Cu, Zn, and Ni (Fig. 5.4 a). In general, the

differences in q0 between most of the tested soils were statistically significant (p

< 0.05). In addition, the difference in q0 between HMs in individual soil samples

was statistically significant (p < 0.05).

CHAPTER 4 | 77

Fig. 5.4: (a) Comparison between adsorption capacities (q0) for Cu, Zn and Ni in scenario A, B and C. (b)

Relative capacity changes (q0 %) in soils amended with Cu, Zn and Ni compared to control experiment.

a b

78 | CHAPTER 4

3.3.2. Scenario B (prior addition of DOM)

Differences in the sorption behavior of the studied metals were observed in

scenario B (exp. B) compared to the control experiment. The highest value of q0

observed for all the metals in all studied soils was for Cu (Table 3.4). Much

lower values of q0 were obtained for Zn and Ni (Fig. 5.4 a). Soil-birnessite had

the highest value of q0 for Cu, while soil-goethite showed the lowest one. For

both Zn and Ni, soil-control showed the highest q0 values while soil-smectite-

oxides showed the lowest values.

The differences in q0 of Cu for different soils were statistically significant only

between soil-goethite and soil-birnessite (p < 0.05). The relative increases in q0

for Cu in all tested soil samples augmented by prior addition of DOM amounted

to 8-25% except for soil-goethite which showed a 10% reduction in q0

compared to the control experiment (Fig. 5.4 b). On the other hand, the relative

changes in q0 for both Zn and Ni showed larger reductions for Zn (11-36%) than

for Ni (1-21%) (Fig. 5.4 b). Between individual soil samples, a statistically

significant difference was observed between Cu & Zn and Cu & Ni (p < 0.05)

but no statistically significant difference was observed between Zn & Ni (p >

0.05).

3.3.3. Scenario C (simultaneous addition of DOM and HMs)

Analogously to scenarios A and B, soil-birnessite had the highest q0 for Cu

while soil-goethite showed the highest q0 for Zn and Ni. On the other hand, soil-

smectite-oxides showed the lowest q0 for Cu, Zn, and Ni (Fig. 5.4 a).

Concurrent addition of DOM and HMs to soil columns resulted in a large

enhancement in q0 for Cu (141-299%), Zn (17-102%), and Ni (32-144%) (Fig.

5.4 b). In general, the differences in q0 for Cu, Zn, and Ni in the tested soils

were statistically significant (p < 0.05). Also, differences in q0 for Cu, Zn, and

Ni in all tested soils were statistically significant (p < 0.05) between scenarios C

& A and scenarios C & B.

CHAPTER 4 | 79

4. Discussion

4.1. Metal sorption and competition in the absence of DOM (scenario A)

The adsorption (both exchange and specific adsorption) capacity of a soil is

determined by the number and kind of binding sites available (Boulding, 1996).

Since sorption of Cu depends on covalent interactions (inner-sphere complex)

with the soil constituents, Cu was more strongly adsorbed than Zn and Ni,

which are predominantly retained through electrostatic interactions (outer-

sphere complex) (Anderson and Christensen, 1988; Gomes et al., 2001). This

finding confirms the sorption results for Cu, Zn, and Ni in our previous batch

study (Refaey et al., 2014). The adsorption of the metal that has a higher affinity

for sorbent sites is less affected by other metals with weaker affinities (Chen,

2012). Accordingly, Cu was found to be the most strongly sorbed and the

strongest competitor for soil constituents and OM in all scenarios (Fig. 4.4).

That Zn exhibited a higher q0 than Ni ions could be due to the fact that Zn

outcompetes Ni in occupying sites available for both metals (Trivedi et al.,

2001; Xu et al., 2006).

Compared to the other soil constituents, birnessite has a higher CEC (247 cmolc

kg-1

), higher SSA (76.5 m2/g) and holds a negative charge in a wider pH range

(Puppa et al., 2013). Consequently, Mn-oxide is a more effective sorbent for Cu

ions than the other soil constituents (Bradl, 2004; Fernandez et al., 2015). Soil-

birnessite showed the highest q0 for Cu of all soils, probably due to penetration

of metal cations into the birnessite layer structure, while soil-goethite showed

higher q0 for Zn and Ni (Fig. 5.4 a). This might be attributed to strong retention

of Cu by soil-birnessite (McKenzie, 1980) reducing the free binding sites for Zn

and Ni, while in soil-goethite the competition between Cu, Zn and Ni was

lower. However, this conclusion should be the subject of further research.

In addition, the presence of large competitive cations such as Ca2+

can affect

HM adsorption in soils. Ca2+

competes effectively with metals for adsorption

sites, and this competition is greater for Zn and Ni than for Cu because Zn and

Ni are predominantly retained in the soil by exchange reactions, while Cu forms

inner-sphere complexes with soil constituents (Pierangeli et al., 2003).

Furthermore, the presence of Ca2+

ions as the dominant cation in the tested soils

suppressed adsorption of metal on Fe-oxide and this also can explain the

80 | CHAPTER 4

superiority of soil-birnessite for adsorption of Cu compared to soil-goethite

(Cowan et al., 1991). The fact that soil-birnessite showed a higher q0 for Cu, Ni

and Zn than the soil-sm-oxides may be attributed to the following: in the weakly

acidic to neutral pH range, the surface of birnessite becomes more negatively

charged than that of goethite. Moreover, the Point-of-Zero Charge (PZC) of

birnessite is lower than that of goethite, and consequently there are more

hydroxy groups available for binding metal ions on birnessite (Xu et al., 2015).

The lower PZC of birnessite is less favorable for the protonation of its surface

(Xu et al., 2015), thereby enhancing the attraction forces between the sorbent

surface and the metal ions (Zhang et al., 2009). As a result, the removal ability

of HMs from their solution by the soil amended with birnessite was higher than

other amended soils in all pH conditions (Xu et al., 2015).

4.2. The effect of the (timing of) DOM addition on metal sorption (scenario B and C).

4.2.1. Effects of timing of DOM addition on Cu retention

Cu is more extensively complexed than Zn and Ni by DOM due to the

formation of strong and stable (inner-sphere) complexes with DOM (e.g.,

McBride et al., 1998; Refaey et al., 2014; Fernandez et al., 2015). The time

interval between breakthrough of Cu and that of both Zn and Ni in the BTCs

was longer in the presence of DOM, reflecting the high affinity of Cu for DOM

(Figueira et al., 2000) (Fig. 4.4).

The timing of DOM addition had a great influence on the mobility of the tested

HMs in current study. Following prior addition of DOM to soil constituents,

DOM ligands can form a bridge between the soil surface and the HMs (Bradl,

2004; Refaey et al., 2014) and a moderate 8-23% enhancement in q0 for Cu was

thus observed, except for soil-goethite which showed a 10% reduction in q0

(Fig. 5.4 b). Fe-oxides in soils form the most dominant reactive sites for DOM

complexation which can result in binding sites for Cu being blocked by OM

(Kothawala et al., 2009; Refaey et al., 2014). This blockage of binding sites can

counteract the potentially enhanced retention of Cu through the mechanisms

mentioned above (Feller et al., 1992; Kaiser and Guggenberger, 2000; Refaey et

al., 2014). The binding sites for Cu on DOM may also have been occupied by

Fe as this metal also forms strong inner-sphere complexes with DOM (Senesi et

CHAPTER 4 | 81

al., 1986; Zhang et al., 2016), thereby reducing the contributions of both

mechanisms (Refaey et al., 2014). Bradl (2004) stated that when the soil pH is

below 5.7 (slightly acidic) the Cu-DOM complex becomes unstable since Fe

displaces Cu and this is consistent with the observed reduction in q0 for Cu in

soil-goethite (pH in general below 5.7) compared to the control experiment

(Fig. 5.4 b).

Furthermore, the goethite surface preferentially removes high molecular weight,

aromatic compounds (Chorover and Amistadi, 2001; Chin et al., 1997; Hur et

al., 2006), which is consistent with the SUVA254 for the DOM used in the

current study (> 4 L mg-1

m-1

), indicating a highly hydrophobic and aromatic

character (Piirso et al., 2012). The q0 of soil-birnessite for Cu was not affected

by prior addition of DOM due lower amounts of DOM being adsorbed on the

birnessite surface. Both DOM and birnessite are negatively charged at pH 6

while goethite is net positively charged at the same pH (Chorover and Amistadi,

2001). As a result, the birnessite surface is less coated with adsorbed DOM than

goethite is due to repulsion of “like” charges.

In contrast, concurrent addition of DOM and HMs to all tested soils (scenario

C) showed remarkable enhancement in q0 for all metals, probably due to cation

bridging and precipitation (Bradl, 2004; Refaey et al., 2014). The q0 for Cu was

greatly enhanced (141-299%) by concurrent addition of DOM and HMs

compared to the control situation (Fig. 5.4 b). Soil-birnessite consistently

exhibited a higher q0 for Cu than other soils and this can be attributed to

birnessite and DOM being the most likely to bind Cu in a nonexchangeable

form. In addition, the presence of DOM increases the hydrolysis of Mn ions,

thereby increasing the likelihood of Mn precipitation, and the negative charge

on the exchange complex (Bradl, 2004).

4.2.2. Effects of timing of DOM addition on Zn and Ni retention

Both Zn and Ni were significantly affected by prior addition of DOM (scenario

B) to soil constituents. The largest reduction in q0 for Ni and Zn (1 to 36%) for

all tested soils compared to the control experiment can be attributed to blockage

of active sites on the soil constituents by OM (Refaey et al., 2014). In general,

the statistically significant (p < 0.05) reduction of q0 for Ni and Zn compared to

82 | CHAPTER 4

the control experiment confirms that electrostatic binding mechanisms and

mineral phases dominate for these two metals resulting in a weaker association

with OM (Refaey et al., 2014). Mineral phases such as clay minerals and

hydroxides predominantly controlled the q0 for Zn and Ni (Fujiyoshi et al.,

1994; Li et al., 2009).

In contrast, all tested soils showed a remarkable enhancement of q0 for both Zn

and Ni by concurrent addition of DOM and HMs (scenario C). The q0 for Ni was

higher (32-144%) than for Zn (17-102%), probably due to the low stability of

organic complexes with Zn (Kalbitz and Wennrich, 1998; Zhang et al., 2016)

and higher affinity of Ni than Zn for DOM (McBride, 1989). Soil-goethite

consistently exhibited a higher q0 for both Zn and Ni than soil-birnessite. This

can be explained by the pH of the effluents from soil-goethite being below 5.7

(slightly acidic) which resulted in Cu-DOM complexes becoming unstable as Fe

displaces Cu (Bradl, 2004) and forms strong inner-sphere complexes with DOM

(Senesi et al., 1986).

The enhanced leaching of OC induced by sequential addition of HMs in

scenario B (Fig. 3.4 a) was probably due to the competition of added HMs for

adsorption sites in the soil solids with the previously adsorbed OM (Weng et al.,

2009; Zhang and Zhang, 2010). In scenario C, concurrent addition of DOM and

HMs (Fig. 3.4 a) showed a remarkable increase of OC in the leachate. This is

probably due to cation bridging and precipitation as a result of metal-DOM

complexes formed in solution prior to adsorption as a result of their concurrent

addition (Seo et al., 2008; Bradl, 2004).

4.2.3. Effects of differences in mineralogical composition on HM retention

Clay minerals in soils play a minor role in the sorption of HMs to soil compared

to (oxyhydr)oxides and DOM (Fernandez et al., 2015). In our experiments, clay

minerals influenced the mobility of the tested metals in the presence of DOM to

some extent. The mineral kaolinite was detected in a considerable amount

(32%) in the soil-control and contributed to adsorption of a large amount of

DOM (Fig. 3.4 a); thereby enhancing q0 for Cu, Zn, and Ni (Stevenson and

Fitch, 1981). The observation that soil-smectite and the ternary complex soil

(soil-smectite-oxides) had a lower q0 for Cu, Zn and Ni in presence of DOM

CHAPTER 4 | 83

could be attributed to smectite reducing the DOM contribution to binding Cu

(Stevenson and Fitch, 1981). Smectite minerals tend to strongly bind

hydrophilic organic material from solution (Meleshyn and Tunega, 2011) but

bind the hydrophobic DOM used in the current study less strongly and thus

removes little HMs from the solutions. This conclusion is also supported by the

fact that a large amount of OC in scenario C was leached in the effluent from

soil-smectite (Fig. 3.4 a). In the ternary soil, DOM could fail to form stable

complexes with birnessite in presence of goethite and Ca2+

because both Fe and

Ca can substitute for Mn (Norvell and Lindsay, 1972).

4.3. Comparison between the batch and column experiments

The column experiments generally confirmed and substantiated the preliminary

results from our previous batch study (Refaey et al., 2014) with respect to the

timing of addition of DOM in retention of HMs. The adsorption of Cu, Zn and

Ni in both experiments in general showed stronger adsorption of Cu compared

to similar sorption of Zn and Ni. In addition, in both studies, Cu showed strong

sorption to DOM and mineral phases by forming strong complexes (inner-

sphere) whereas both Zn and Ni preferred mineral-phase by forming outer-

sphere complexes.

In theory, the batch approach would be expected to overestimate sorbed

concentrations because various kinetic reactions are studied under equilibrium

conditions, while under natural conditions they could be too slow to reach

equilibrium. This could lead to inappropriately optimistic predictions of metal

retention (Plassard et al., 2000; Antoniadis et al., 2007b). However, in our case

the column study showed a larger metal retention compared to our batch

experiments. This could be attributed to adsorption mechanism being

predominant in the batch experiments whereas in the column study other

additional retention mechanisms besides adsorption, such as surface

precipitation, may have been involved (Seo et al., 2008). It is worth mentioning

that in both batch and column studies, the timing of the DOM addition

(concurrent with the metals or sequential) had a large effect on metal retention.

Also, it seems that in both batch and column studies, previously added DOM

84 | CHAPTER 4

(scenario B) blocked the binding sites for Cu, Zn and Ni on soil-goethite and

hence reduced their adsorption capacity.

5. Conclusions

In our study, an important role in regulating the mobility of HMs in soils was

played by the timing of DOM addition (concurrent with or prior to HM

addition). All tested metals showed strong enhancement of adsorption with

concurrent addition of DOM (scenario C) compared to prior addition of DOM

(scenario B). Both Zn and Ni showed reduced retention to soil components

following prior adsorption of DOM, confirming our previous findings that

mineral-phases are preferential sorbents for these two metals. Conversely, Cu

exhibited higher sorption to both DOM and mineral phases by forming stable

inner-sphere complexes. Timing of DOM addition with respect to that of HM

therefore has to be taken into account when assessing the influence of HM

pollution of soils through polluted irrigation- or wastewater in a system where

DOM also enters the soil (e.g. agricultural irrigation in combination with

manuring). Similarly, both the presence of DOM and timing thereof should be

taken into account in design of strategies where soil constituents, e.g. clay

minerals, are used to clean-up HM polluted waste water.

| 85

86 |

CHAPTER 5 | 87

Chapter 5

The influence of organo-metal interactions on regeneration of

exhausted sorbent materials loaded with heavy metals

Abstract

Natural clay minerals can play an important role in crude remediation of wastewater

polluted with the heavy metals (HMs) Cu, Zn and Ni. We previously showed that the

presence and timing of addition of natural dissolved organic matter (DOM) plays a key

role in regulating HM removal by clay mineral sorbents. However, the influence of the

presence of DOM on the remediation of the used sorbents once saturated with HMs is

largely unknown. To resolve this, clay mineral rich column material of varying

composition previously loaded with Cu, Zn and Ni only; first with DOM followed by

Cu, Zn and Ni; or DOM, Cu, Zn and Ni simultaneously was used in a set of desorption

experiments. The columns were leached by 0.001 M CaCl2 dissolved in water as control

eluent and 0.001 M CaCl2 dissolved in DOM as treatment eluent. Our results show a

significant influence of the timing of DOM addition (sequential or concurrent with

HMs) during the preceding loading phase of the sorbent on the subsequent removal of

the HMs. In particular when the column was loaded with DOM and HMs

simultaneously, largely irreversible co-precipitation took place. Our results indicate that

regeneration potential of sorbents in wastewater treatment will be significantly reduced

when the treated water is rich in DOM. In contrast, for natural soil systems our results

suggest that when HMs enter together with DOM, e.g. in manured agricultural fields,

HM mobility will be lower than expected from interaction dynamics of HMs and clay

minerals alone.

This chapter is based on: Refaey, Y., Jansen, B., de Voogt, P., Parsons, J.B, El-Shater, A., El-Haddad, A., Kalbitz, K. 2016. The influence of organo-metal interactions on regeneration of exhausted sorbent materials

loaded with heavy metals (under review: Pedosphere Journal).

88 | CHAPTER 5

1. Introduction

The amount of wastewater contaminated with heavy metals (HMs) worldwide

increases continuously due to the expansion of industrial activities, and has

been subject of much research as result of the related public health hazards (Qin

et al., 2004; Fu and Wang, 2011). Adsorption technologies using natural clay

minerals are seen as an important remediation measure, particularly in

developing countries where more sophisticated techniques are usually not

widely available (Ikhsan et al., 2005; Gu et al., 2010; Neto et al., 2012; Refaey

et al., 2014). Clay minerals and/or hydroxides (Mn- and Fe-oxides) are

adsorbents that are both abundant and cheap (Moreno-Castilla and Rivera-

Utrilla, 2001; Al-Qunaibit et al. 2004; Colombani et al. 2015). Given their high

specific surface area (SSA) and cation exchange capacity (CEC), the mobility

and bioavailability of HMs can be substantially reduced by interactions with

clay minerals, hydroxides and (D)OM (Stahl and James, 1991; Kalbitz and

Wennrich, 1998; Al-Qunaibit et al., 2004; Antoniadis et al., 2007; Refaey et al.,

2014; Colombani et al., 2015; Refaey et al., 2015, 2016).

For reasons of cost-efficiency, ideally such wastewater treatment approaches

using clay minerals would use a continuous system in which sorbent materials

can be used in multiple cycles of metal sorption and desorption (Mehta and

Gaur, 2005; Kumar et al., 2012). Since desorption often controls the long term

environmental fate of most contaminants and treatment feasibility, insights in

the recovery potential of HMs from clay minerals used in wastewater treatment

is essential (Mustafa et al., 2004; Hu and Shipley, 2012). After the adsorbents

are exhausted, they are either to be discarded or, preferably, recovered for reuse.

Spent sorbents should be released into the environment only after removal of

the adsorbed HMs to avoid secondary pollution to soil and groundwater systems

(Karathanasis, 1999; Tzou et al., 2007; Lata et al., 2015). In spite of the

importance of understanding and optimizing regeneration adsorption materials

for wastewater treatment, this has received surprisingly little research attention

(Glover et al., 2002; Covelo et al., 2004b; Kandpal, 2005; Feng et al., 2012;

Lata et al., 2015). In particular, knowledge is lacking about the influence of

dissolved organic matter (DOM) in HM containing wastewater on the

subsequent removal potential of the HMs once adsorbed. In our previous work

we showed that not only the presence of DOM, but also the timing of its

CHAPTER 5 | 89

addition prior to or simultaneously with HM addition, has great influence on the

removal efficiency of Zn, Ni and Cu through adsorption on clay minerals

(Refaey et al., 2016).

Specifically, our previous work (Refaey et al., 2014, 2016) confirmed that Cu is

mostly retained to clay minerals and hydroxides through inner-sphere

complexes whereas both Zn and Ni were found to bound predominantly through

outer-sphere complexes (electrostatic interactions). Furthermore, the addition of

DOM and its timing of addition had a significant effect on the removal of HMs

from aqueous solution. The concurrent addition of DOM and HMs to the

sorbent materials resulted in a large enhancement of the affinity and adsorption

capacity for all tested HMs, and particularly of Cu because of its highly affinity

toward (D)OM (e.g., Lair et al., 2007). In contrast, sequential addition of DOM

to the sorbents (prior to HMs) resulted in decreased affinity and adsorption

capacity of all tested HMs due to coating or blocking the binding sites on the

clay minerals and hydroxides in sorbent materials (Refaey et al., 2014).

Therefore, both the presence of DOM and timing of its addition with respect to

that of HM should be taken into account in design of wastewater cleanup

strategies based on adsorption on clay minerals. However, while our previous

work showed significant effects of the timing of DOM addition on HM removal

from solution by clay minerals, it is unclear how it may influence desorption

behavior of Cu, Zn and Ni in column regeneration upon saturation with HMs.

Therefore, the objective of this study was to investigate the role of the presence

and timing of addition of DOM during loading of clay mineral based

wastewater treatment columns on the subsequent removal of the HMs from the

columns upon use. For this we used the same clay mineral columns of varying

clay mineralogical composition that were used in our previous study after their

saturation with HMs and under the various DOM addition scenarios previously

tested. Two desorption reagents were investigated based on the assumption that

such reagents should be cost effective, eco-friendly and must not damage the

sorbent materials (Das, 2010). Since the majority of metals are adsorbed via ion

exchange reactions and are in competition for adsorption sites with other cations

such as Ca2+

, Mg2+

, Na+ and K

+ (Harter, 1992). Therefore, using of the

competing cations for enhancing HMs desorption deserves attention. Many

authors have proposed the use of natural salts such as CaCl2 as an extraction

90 | CHAPTER 5

reagent because of its low cost, relatively low environmental impact, and

efficient for the regeneration of HMs without destroying the sorbent matrix

(e.g., Reed et al., 1996; Houba et al., 2000; Makino et al., 2006; Meers et al.,

2007). In addition, this extraction procedure (CaCl2) is used in the Dutch

legislation for the assessment of nutrients and HMs in soils (Pueyo et al., 2004).

Furthermore, Ca2+

is the most common divalent cation in soil and groundwater,

it is nontoxic at high concentrations, and there is no drinking water standard set

for this element (Wang et al., 1997). In addition, complexing agents such as

organic compounds have been also investigated recently for enhancing

desorption of HMs (Weber, 1988; Milczarek, 1994; Tan et al., 1994). As a

result a CaCl2 solution in water (control eluent) and a CaCl2 solution in DOM

(treatment eluent) were chosen as reagents for column regeneration.

2. Material and methods

2.1. Experimental procedure

In our preceding adsorption study (Refaey et al., 2016), soil samples from

Southern Limburg, The Netherlands and mixed solutions of chloride salts of Cu,

Ni, and Zn (~ 25 mg L−1

) were used in the column adsorption experiments. In

the adsorption step, the original soil was amended with smectite (10%), goethite

(1%) and birnessite (1%) because of their prominent role in regulating the

binding affinity of Cu, Ni, and Zn (Refaey et al., 2014, 2015). The mineral

amendments were used to create five different soil compositions as described in

previous study (Refaey et al. 2016; Table 1.4). Each prepared soil was mixed

with 5.00 g of sand (50-70 mesh particle size, SiO2, Sigma-Aldrich) to increase

the hydrological conductivity of the soil once packed in the column, so the final

weight of each prepared soil was 10.0 g (Table 1.5; Refaey et al. 2016). The

DOM used in both the current and our previous study was prepared by aqueous

extraction from OM (soil with natural manure) following the method described

by Refaey et al. (2014, 2016).

CHAPTER 5 | 91

2.1.1. Experimental set-up

The preparation of soil columns (12 x 2.5 cm) for the adsorption step was

described in detail in our previous study (Refaey et al., 2016; Table 1).

Afterwards, the same soil columns were used to study column regeneration in

the present study. To prevent preferential flow-paths and for precise control of

the flow rate, the regeneration eluent solutions were pumped upward against

gravity by means of peristaltic pumps (Minipuls 3, Gilson) with flow rate set at

0.333 ml/min. The desorption experiments were conducted on the previous

loaded sorbent with tested metals through three previously adsorption scenarios

(A, B and C) as described in Table 1.5.

Table 1.5: Experimental setup for adsorption-desorption column experiments

2.1.2. Regeneration eluent

Two CaCl2 eluents were prepared, the first eluent (control) was 0.001 M CaCl2

dissolved in deionized water and the second eluent (treatment) was 0.001M

CaCl2 dissolved in DOM. The eluent solutions were adjusted to pH 6 before

starting the experiments by adding appropriate amounts of 0.1 and 0.01 M

NaOH to avoid precipitation of DOM. The adsorption experiments in our

previous was carried out in quadruplicate (Refaey et al. 2016) for each tested

sorbent (soil), therefore in the present study for each sorbent, the first 2

duplicates were regenerated with the control eluent and the other 2 duplicates

Desorption

treatment Column pretreatment (for details see Yasser et al., 2016)

Eluent

Scenario A

(during adsorption

step)

Scenario B

(during adsorption step)

Scenario C

(during adsorption step)

I: Flush with

CaCl2 only

(control eluent)

I-A: sorbents were

loaded with HMs

only

I-B: sorbents were

loaded with DOM first,

then HMs

I-C: sorbents loaded with

DOM and HMs

simultaneously

II: Flush with

CaCl2 and DOM

(treatment eluent)

II-A: sorbents

were loaded with

HMs only

II-B: sorbents were

loaded with DOM first,

then HMs

II-C: sorbents loaded with

DOM and HMs

simultaneously

92 | CHAPTER 5

were regenerated with treatment eluent. Desorption of metal-loaded sorbents

was initiated by continuous flow of eluent solution at a flow rate of 0.333 mL

min-1

. A constant head of reagent solution was maintained in the column

throughout the desorption period (up to continuous 18 h). Eluted fractions were

collected at each 45 min from 60 columns (20 column for each scenario). The

concentration of HMs and DOM in each eluted fractions were determined by

ICP-OES (PerkinElmer-Optima 3000XL) and a TOC analyzer (TOC-VCPH,

Shimadzu, Kyoto, Japan), respectively.

2.2. Desorption parameters

To evaluate the regeneration process, the desorbed amounts of tested metals

were calculated by Eq. 1.5 (e.g., Voleski et al., 2003; Lodeiro et al., 2006) from

the desorption curve which is equivalent to the breakthrough curve in the

adsorption step.

𝑀𝑑 =𝐹

𝑀𝑠∫ 𝐶𝑑𝑑𝑡

𝑡=𝑒

𝑡=0 Eq. 1.5

where, Md is calculated from the numerical integration of the regeneration

curves from t=0 to t=e, the time (e) corresponds to the time required for total

elution of HMs in column. Cd is metal concentration (mg L-1

) after the elution

process at time t (min), the integrate part was calculated by the area below the

elution curve (Cd versus time) multiplied by the flow rate (F) and soil mass

(Ms). The computer program ORIGIN was used to calculate (by numerical

integration) the area under the curve.

The percentage of desorption was described by the desorption efficiency

percentage (E %) considering the amount of total removal (Md) as 100% of the

metal that could be eluted from the adsorbent (Voleski et al., 2003; Lodeiro et

al., 2006). This parameter was obtained from dividing the amount of metal

desorbed (Md) by the amount of metal bound to the sorbent in the previous

adsorption experiments (Ma) as follow:

𝐸 (%) =𝑀𝑑

𝑀𝑎100 Eq. 2.5

CHAPTER 5 | 93

3. Results

3.1. Regeneration of metal-loaded sorbents using control eluent

The removal efficiency percentage (E%) for the HMs followed the sequence Ni

> Zn > Cu in scenario A and Zn > Ni > Cu in scenario B and C (Table 2.5; see

Table 1.5 for a description of the scenarios).

The recovery of Cu from sorbents in scenario A and B was quite similar except

for the soil-smectite sorbent that showed the highest Cu removal in scenario A

compared to B (Fig. 1.5 a). Furthermore, a large variation in E% among the

tested sorbents was recorded for Cu in scenario A (19-64%) compared to B (26-

37%). For both Zn and Ni, the E% was higher in scenario A (37-81%, for Zn;

41-89%, for Ni) compared to B (39-57%, for Zn; 37-53%, for Ni) as shown in

Figure 2.5 a and 3.5 a. In scenario C, the E% for Cu (2-5%), Zn (11-18%) and

Ni (8-17%) was always much lower than A and B (Table 2.5; Figs. 1.5 a, 2.5 a,

3.5 a).

Table 2.5: Performance (desorbed metal mg/g) of control eluent (CaCl2) and efficiency of

removal percentage (E%) of metals from of loaded-sorbents.

Sorbent Metal Scenario A Scenario B Scenario C

mg/g E% mg/g E% mg/g E%

Soil-control Cu 1.201 35.4 0.366 29.1 0.121 3.5 Zn 0.970 49.1 0.382 47.0 0.255 14.4

Ni 0.770 50.2 0.319 40.8 0.177 11.6

Soil-smectite Cu 0.578 63.7 0.343 34.3 0.158 2.2 Zn 0.618 62.6 0.318 50.7 0.327 10.6

Ni 0.411 82.0 0.262 42.6 0.272 9.2

Soil-goethite Cu 0.326 31.7 0.384 34.0 0.207 4.5

Zn 0.762 70.6 0.375 52.3 0.361 18.3 Ni 0.641 89.1 0.311 45.8 0.278 16.7

Soil-birnessite Cu 0.432 19.3 0.348 25.6 0.111 1.8 Zn 0.494 37.4 0.336 39.1 0.293 12.6

Ni 0.405 40.8 0.284 36.7 0.217 10.9

Soil-smectite-oxides Cu 0.351 48.0 0.386 36.8 0.086 1.6

Zn 0.358 80.5 0.367 57.2 0.269 11.4

Ni 0.300 75.8 0.328 53.4 0.180 8.1

94 | CHAPTER 5

Fig. 1.5: Removal efficiency percentage (E%) for Cu in scenario A, B and C using control (a) and treatment

(b) eluents.

Fig. 2.5: Removal efficiency percentage (E%) for Zn in scenario A, B and C using control (a) and treatment (b) eluents.

a b

a b

CHAPTER 5 | 95

Fig. 3.5: Removal efficiency percentage (E%) for Ni in scenario A, B and C using control (a) and treatment

(b) eluents.

3.2. Regeneration of metal-loaded sorbents using treatment eluent

The E% for the HMs followed the order Cu > Ni >Zn in scenario A and B while

in scenario C the order was Cu > Zn > Ni (Table 3.5). By using DOM, the

desorption under scenario B was higher than A for all tested metals and for all

tested sorbents (Table 3.5). Furthermore, a large increase in desorption for only

Cu was achieved under scenario B (69-78%) compared to flushing with control

eluent (26-37%) (Fig. 1.5 b). Whereas for both Zn and Ni, mostly a reduction of

desorption under scenario A was found (47-53%, for Zn; 53-60%, for Ni)

sometimes combined with an enhancement of desorption under scenario B (54-

69%, for Zn; 68-74%, for Ni).

The recovering of the tested HMs under scenario C was always much lower

than under scenario A or B (Figs. 1.5 b, 2.5 b, 3.5 b). For Cu, while still

remaining smaller than for A and B, desorption under scenario C was

significantly enhanced by flushing with treatment eluent (15-25%) as compared

to control eluent (2-5%) (Fig. 1.5 b). Only small differences in desorption of

both Zn and Ni were found upon flushing with control (11-18%, for Zn; 8-17%,

for Ni) or treatment eluent (12-19%, for Zn; 12-15%, for Ni).

a b

96 | CHAPTER 5

Table 3.5: Performance (desorbed metal mg/g) of treatment eluent (DOM) and efficiency of

removal percentage (E%) of metals from of loaded-sorbents.

Sorbent Metal Scenario A Scenario B Scenario C

mg/g E% mg/g E% mg/g E%

Soil-control Cu 1.021 62.7 0.802 69.0 0.717 19.9

Zn 1.015 51.9 0.540 61.1 0.312 14.6

Ni 0.819 59.3 0.517 73.9 0.240 14.7

Soil-smectite Cu 0.946 73.9 0.713 77.6 0.650 25.3

Zn 0.960 53.3 0.440 59.6 0.328 17.6

Ni 0.766 59.9 0.388 68.5 0.253 14.4

Soil-goethite Cu 0.569 53.2 0.650 75.2 0.590 19.1

Zn 0.490 51.7 0.363 54.4 0.353 15.3

Ni 0.443 60.1 0.352 67.8 0.285 13.4

Soil-birnessite Cu 0.662 61.3 0.900 74.3 0.799 24.6

Zn 0.417 47.1 0.433 59.9 0.350 19.0

Ni 0.377 58.0 0.406 68.4 0.236 13.9

Soil-smectite-oxides Cu 0.701 70.8 0.875 78.2 0.702 15.2

Zn 0.421 49.2 0.456 68.9 0.325 12.1

Ni 0.378 53.7 0.400 68.4 0.245 11.7

3.3. The effect of sorbent composition on the performance of the reagent

Upon flushing with the control eluent, hydroxide-rich sorbents (soil-birnessite

and soil-goethite) showed a low release of Cu, while the obtained E% were

quite similar for scenario A and B (Table 2.5, 3.5). Whereas upon flushing with

DOM, the release of Cu was remarkably enhanced, in particular from soil-

goethite, as compared to the control eluent. For both Zn and Ni, their release

from soil-goethite upon flushing with the control eluent was low in scenario B

compared to A (Table 2.5; Fig. 2.5 a, 3.5 a). Also, the soil-smectite sorbent

showed a large release for Cu, Zn and Ni in scenario A compared to B (Table

2.5, 3.5).

CHAPTER 5 | 97

4. Discussion

4.1. Regeneration efficiency using the control eluent

The removal efficiency of HMs by CaCl2 may be considered as an evaluation

model for the strength of bonds and mobility of the HMs (Kowalkowski et al.,

2010). Since sorption of Cu depends mainly on covalent interactions (inner-

sphere complex) with the soil constituents (Kandpal et al., 2005; Kowalkowski

et al., 2010; Refaey et al., 2014, 2016), as expected Cu proved rather resistant to

exchange with Ca2+

in the present study. In contrast, both Zn and Ni as

predominantly retained by exchange reactions (outer-sphere complexes)

(Refaey et al., 2014; 2016), were effectively removed by the CaCl2 solution as a

result of cation exchange with the abundant Ca2+

(Figs. 2.5 a, 3.5 a; Table 2.5).

The absence of OM in the metal-loaded sorbents under scenario A led to much

higher recovering of HMs from smectite-amended sorbents (soil-smectite and

soil-sm-oxides) than the other soil compositions tested (Figs. 1.5 a, 2.5 a, 3.5 a).

This can be explained by the fact that smectite-rich soil is composed of

aluminosilicate minerals, which favored cation exchange of metal ions during

adsorption (Atanassova, 1995; Abat et al., 2012). Interestingly, soil-birnessite

showed the lowest release of Cu, Zn and Ni among the soils tested, indicating

that the previously observed large affinity of all tested metals for the birnessite-

rich sorbent (Refaey et al., 2016) contributed at the same time to the slowing

down of metal recovering, which is in line with other studies (Khan et al., 2005;

Wang et al., 2010; Fernandez et al., 2015). The overall removal sequence of

tested metals under scenario A was Ni > Zn > Cu which agrees with our

previous study where we found a high affinity of Zn and Cu compared to Ni for

adsorption on smectite and hydroxides (Refaey et al., 2016).

Owing to the stability of Cu, Zn and Ni complexes with previously adsorbed

OM, reduction in their recovery was recorded under scenario B compared to A,

in particular for Cu forming most stable complexes with OM compared to Zn

and Ni, thereby, the competitive exchange with Ca2+

was weak (Stevenson,

1994; Bradl, 2004). For that, the E% for Cu was higher than for both Zn and Ni

(Figs. 1.5 a, 2.5 a, 3.5 a). Moreover, the high electronegativity of Cu and Ni and

their tendency to form strong bounds with OM than Zn, led to decrease in their

E% from loaded sorbents (Tyler and McBride, 1982).

98 | CHAPTER 5

4.2. Regeneration efficiency using treatment eluent

Cu is much more susceptible to specific binding to OM than Zn and Ni owing to

its previously mentioned preference for inner-sphere complexation (Karlsson et

al., 2006; Lair et al., 2007). As a result, using the DOM containing treatment

eluent enhanced Cu desorption in all three scenarios as compared to the control

eluent that did not contain DOM. The enhancement in Cu recovery was

particularly strong in scenario B, most likely because in this scenario where

columns were previously loaded first by DOM after which Cu was added, the

Cu was to a large extent bound to DOM adsorbed on the mineral phase rather

than to the mineral phase itself. As a result removal of this fraction consisted of

simple partitioning between binding of Cu to DOM adsorbed on the mineral

phase and DOM present in the treatment eluent. In addition, the binding affinity

of Ca2+

for DOM is much weaker than that of Cu to DOM, which also explain

the removal efficiency of Cu over Zn and Ni (Ma et al., 1999).

In contrast, the removal efficiency of both Ni and Zn was reduced in scenario A

and remarkably enhanced in scenario B when the treatment eluent was used

instead of the control eluent (Figs. 2.5 b, 3.5 b). The most likely explanation is

that part of the Ca2+

in the treatment eluent was bound to the DOM in solution,

leaving less free Ca2+

to displace Zn2+

and Ni2+

adsorbed on the columns under

scenario A. The enhanced desorption of Zn and Ni in scenario B when using

treatment eluent suggest again partitioning between the DOM adsorbed on the

mineral phase and the DOM offered in the treatment eluent. Given the outer

sphere bonding character of Ni and Zn, such association with DOM on the

mineral phase was much likely through non-specific binding mechanisms

(Bradl, 2004).

4.3. Effect of previously timing of DOM addition on the elution of Cu, Zn and Ni

The present study confirms a clear effect of the timing of addition of DOM

during the loading of the columns on the desorption of Cu, Zn and Ni from the

various soils tested. The effects of previous sequential column loading first with

DOM and subsequently with HMs (scenario B) upon flushing with control or

treatment eluent was described in the previous paragraphs. Metal removal

behavior after column loading where HMs and DOM were added concurrently

CHAPTER 5 | 99

(scenario C) was markedly different. Removal percentages were much lower for

both the control and treatment eluents for all metals and from all soils tested

(Figs. 1.5, 2.5, 3.5).

Apparently, concurrent addition of HMs and DOM led to predominantly

irreversible immobilization of loaded metals. This suggests that immobilization

does not take place predominantly via adsorption of the metal cations

themselves to clay minerals or adsorbed DOM as in scenario B, but most

likely, through (co)precipitation of DOM-HM complexes in the column as also

hinted on in our previous study (Refaey et al., 2016). Organic ligands can form

chelate complexes with HMs, causing them to be tightly adsorbed to the soil

constituents (Alloway, 1995). With co-precipitation as dominant immobilization

mechanism, flushing with CaCl2 solutions in water or DOM is inefficient as

only little metal is available for desorption. These results once more underpin

the crucial, yet to our knowledge never previously considered role of the timing

of HM and DOM addition on HM mobility in soils. This has a significant

impact on risk assessment with respect to the mobility of HMs in soils and

connected groundwater systems, indicating that such mobility may be greatly

reduced in scenarios where HMs are introduced simultaneously with DOM, e.g.

through application of HM rich manure, but not in systems where they are

introduced sequentially.

4.4. Performance of examined eluents and potential reusability of sorbent material

The control eluent was highly effective in recovering both Zn and Ni from the

mineral phase in scenario A and the maximum removal was up to 80 and 90%,

respectively. Whereas, DOM was only effective in scenario B where the

maximum removal of Zn and Ni was 70 and 75%, respectively. No adequate

removal of Zn or Ni could be obtained under scenario C (concurrent saturation

of the column with DOM and HMs), the maximum being removed was, 19%

for Zn, and 17% for Ni. For Cu, the recovering process was effective only using

treatment eluent in scenario A and B, yielding a maximum removal of 75 and

80%, respectively. Again, removal efficiency under scenario C was inadequate

with a maximum of 25%. These results imply that regeneration of clay mineral

columns used in wastewater treatment of HM polluted waste water through

100 | CHAPTER 5

flushing with CaCl2 solutions with or without DOM may be a viable technique,

but only in cases where the wastewater itself did not simultaneously contain

HMs and DOM.

5. Conclusion

Results have shown that in absence of loaded DOM (scenario A), Ni and Zn

that are predominantly bound to the mineral phase through reversible outer-

sphere complexes showed substantial removal of up to 81% (Ni) and 89% (Zn)

through simple cation exchange with Ca2+

under using control eluent. However,

by using DOM in the eluent a reduction in desorption of Ni and Zn was

registered because of binding part of exchangeable Ca2+

to DOM. In contrast,

Cu was less readily recovered by Ca2+

(maximum 64%) due to its inner-sphere

complexation with sorbent, but its desorption remarkably enhanced to up to

74% when eluted with DOM because of its inner-sphere complexation with

functional groups on DOM. The previously loaded HMs together with DOM

showed substantial differences in desorption of loaded HMs depending on

whether HM loading had taken place concurrently with (scenario C) or

sequential to DOM loading (scenario B). When columns were first loaded with

DOM followed by HMs, the largest removal efficiency was achieved using

DOM in the eluent (up to 69% for Zn, 74% for Ni and 78% for Cu). This

indicates a partitioning of HMs bound to DOM adsorbed on the solid phase, to

DOM in solution. However, when columns were loaded by HMs and DOM

simultaneously prior to desorption the removal efficiencies were rather low for

all metals (2-25% for Cu; 11-19% for Zn; and 8-17% for Ni depending on clay

mineral composition) regardless of whether desorption treatment consisted of

CaCl2 solution in water or in DOM. This indicates that column loading by HMs

and DOM when added simultaneously takes place to a large extent through

irreversible co-precipitation rather than adsorption.

For the purpose of regeneration potential of clay minerals used in waste water

treatment, the present study indicates that such potential will be significantly

reduced when the water to be treated is rich in DOM. In contrast, for natural soil

systems our results suggest that when HMs enter a soil together with DOM, e.g.

as a result of the use of HM rich manure in agricultural fields, the mobility of

CHAPTER 5 | 101

the HMs will be lower than expected from interaction of HMs and clay minerals

alone. In general, Cu loaded soils are more susceptible to remobilization of Cu

when DOM rich water infiltrates, whereas Ni and Zn loaded soils are more

susceptible to remobilization when cation rich water infiltrates. But in

circumstances where precipitation plays a role (scenario C) additional measures

would be needed, for instance acidification to redissolve precipitates.

102 |

CHAPTER 6 | 103

Chapter 6

Synthesis

The synthesis chapter seeks to provide connections between the previous

chapters of this dissertation and derive overall conclusions and

recommendations. Specifically, this chapter will discuss the implication of the

insights gained in the previous chapters for the potential application of clay-rich

soils in wastewater treatment with a focus on removal of HMs.

1. Implication of using wastewater in irrigation in presence of Pliocene clay

The annual report (2009) issued by the Egyptian Environment Affairs Ministry

indicates that the contamination of drinking water has reached a critical stage as

farmlands are being irrigated by water polluted with sewage. This leads to the

spread of many diseases such as cancer (100,000 person/year), kidney failure

(15,000 person/year), cholera, typhoid, Schistosoma and hepatitis. In the new

reclamation areas in the Sohag governorate that are characterized by their sandy

and gravelly texture (chapter 2), Pliocene clay deposits are often added to the

soils to decrease the water drain. Based on our findings, irrigation of these areas

with wastewater can result in retention of a large amount of HMs in the soil in

particular when the amended soil is poor in OM. Such accumulation of Cu, Zn

and Ni in soil can pose a threat to human health and the human food chain

(chapter 2 and 4). On the other hand, when the soil amended with the Pliocene

104 | CHAPTER 6

clay is rich in OM for instance due to manuring (prior to irrigation with

wastewater), this can result in a reduction in retention of HMs due to coating of

the binding sites on the soil constituents by OM (chapters 2 and 4).

Consequently, in this case a large amount of HMs (included in wastewater) can

make their way down into the groundwater causing groundwater pollution

(chapter 4). Therefore, a pretreatment process for wastewater prior to irrigation

is recommended to control the drinking water pollution and protect the

surrounding environment. The Pliocene clays could be used for such

pretreatment that can provide a large amount of water in particular for areas

which suffer from water scarcity as in some of the desert outskirts adjacent to

the studied area in Egypt.

2. Implication of using Pliocene clay deposits in the clean-up of wastewater

In Egypt about 38 million people drink polluted water and the amount of

untreated or partially treated industrial effluents that enter the water supply is

about 4.5 million tons/year (Egyptian Organization for Human Rights, 2009).

Due to population growth and the low coverage of wastewater services in

villages and rural areas, the Mediterranean Sea, the Nile River and the Egyptian

desert area all receive large flows of mostly untreated domestic, pesticides and

chemical fertilizer residue from agricultural application, and industrial

wastewater (Abdel-Shafy and Aly, 2002). Using natural low-cost local materials

as sorbent is crucial in developing countries, such as Egypt, since more

sophisticated techniques are often not widely available. Purification of

wastewater could be a reasonable choice to mitigate the shortage and scarcity of

fresh water resources in Egypt (Radwan and Salama, 2006; Alfarra et al., 2011).

Furthermore, because it contains non-degradable HMs, there is a difficulty of

using sludge resulting from wastewater treatment plants in Egypt as a fertilizer

in agriculture or as an ingredient of compost (Gaber, 1994). By using Pliocene

clays for removal of the HMs from the wastewater in treatment plants (at the

effluent points), the problem of using the sludge as a fertilizer will be solved

and make the Pliocene clays potentially useful in high-value-added markets.

The information gained in the present study regarding the

adsorption/regeneration characteristics of the Pliocene clay deposits from the

Sohag area point to great promise for use in wastewater treatment technology.

CHAPTER 6 | 105

Our results clearly showed that large amounts of HMs can be removed by

Pliocene clays in particular when clay materials contain goethite or birnessite

minerals (chapter 3 and 4). The Al-Kwamel (West Sohag) clays and Wadi

Qasab (East Sohag) in addition contain hydroxides as part of their composition,

which can enhance the removal of HMs from wastewater (chapter 2). Since the

studied locations of Pliocene clay in the Sohag area vary in their proportional

composition and properties (i.e. clay minerals, SSA, CEC and hydroxides), a

mixture of materials from different locations is recommended to obtain an

optimal adsorption capacity. A next recommended step is to initiate a pilot to

apply and further test their application in the field.

3. Regeneration of spent Pliocene clay to be used in multiple treatment cycles

An important step in the application of adsorbents to remove contaminants is

the (im)possibility of regenerating the adsorbent after use. In particular in the

case of HMs, which cannot be degraded, this issue must be addressed. The

present study provides new data about the regeneration of sorbent materials

loaded with Cu, Zn and Ni and in particular the role of the presence of (D)OM

as well as the kinetics thereof. Again an important issue because wastewater

often contains significant amounts of DOM as well. Based on the findings of

this study, two scenarios can be sketched:

3.1. Sorbents filled with HMs in the absence of organic matter

In this case, with simple exchange cations, such as Ca2+

(0.001 M), a large

amount of Zn and Ni (~ 80-90%) can be recovered from the spent sorbent

materials (chapter 5) due to their weak binding to the sorbent (outer-sphere

complexes). Smectite amended sorbents showed the highest release of sorbed

HMs in the present study. Therefore, the Pliocene clays from the Sohag area

with their high smectite contents could easily be regenerated when loaded with

Ni and/or Zn. Because it was found to be sorbed through strong inner-spere

complexation, regeneration of Cu was only 64% using Ca2+

in water as a

removal agent. However by using DOM as solvent for CaCl2 salts instead of

water, the leached amount can be increased to 74% due to the formation of

106 | CHAPTER 6

inner-sphere complexes of Cu with DOM (Karlsson et al., 2006; Lair et al.,

2007). In the present study we examined only 0.001 M CaCl2 in water or DOM

solution as eluent. Studies using higher concentrations or other, stronger,

eluents, such as ethylenediaminetetraacetic acid (EDTA) with its known for its

high affinity for Cu (Schramel et al., 2000; Brun et al., 2001) are recommended

to further explore the regeneration of used sorbents. Also here cost efficiency

should be taken into account.

3.2. Sorbents filled with HMs in the presence of organic matter

When HMs were adsorbed on clay minerals in the presence of DOM, our results

show the regeneration will not be efficient regardless of whether desorption

treatment consists of flushing with a CaCl2 solution in water or in DOM. In all

cases and for all HMs the removal efficiency was very low (2-25% for Cu; 11-

19% for Zn; and 8-17% for Ni). This indicates that sorbent loading by HMs and

DOM when added simultaneously to a large extent takes place through

irreversible co-precipitation rather than adsorption. Further study should focus

on additional measures specifically targeted at redissolving precipitates, for

instance acidification, in this case.

4. Implications for soil pollution remediation

The influence of the presence of DOM is not limited to the regeneration

potential of clay minerals used in wastewater treatment. Remediation of soils

contaminated with HMs is important to reduce the associated risks for

groundwater and make the soil available for agricultural production. The

experimental situations in this study where DOM was added first to the sorbent

followed by HMs corresponds with the real life situation where a soil rich in

OM, e.g. due to manuring, is subsequently irrigated with wastewater polluted

with HMs. In such a case, based on our findings remediation through flushing

with a CaCl2 solution would be an option to remove the majority of Zn, Ni and

Cu adsorbed on clay minerals in the soil. However, in the case where HMs and

DOM are added simultaneously, such as when wastewater contains both HMs

and DOM, results from the corresponding experimental situations in our study

suggest alternative remediation routes would need to be pursued.

CHAPTER 6 | 107

5. Conclusions

Our findings provided new data about simple wastewater treatment technologies

which can be designed to provide clean water to meet the needs of a growing

economy and to protect the environment. The challenge in implementing this

strategy will be to persuade the Egyptian government and the private sector for

adoption the present findings to implement this technology on the large scale.

Implementation of this low cost treatment technology in Egypt should be the

next step towards reducing the impact of water scarcity and pollution by reuse

of wastewater. In addition, field approaches combined with laboratory

investigation for each industry are necessary for further improvements and to

fully understand and optimize wastewater clean-up applications using local clay

materials. A further study in the future concerning the activation (acidic or

thermal) of Pliocene clay and work on clay fraction content is necessary to

explore broadening of its application to the removal of organic and inorganic

pollutants. Moreover, different types of clay mineral that vary in their

mineralogical composition are distributed in different areas in Egypt as well as

other countries in the world facing similar problems. The fundamental insights

gained in the present study can guide local endeavors to explore local soil

materials in wastewater treatment there as well. This hold in particular for other

countries with large reserves of clay materials-rich in smectite and suffering

from either scarcity of water resources or environmental pollution.

108 |

REFERENCES | 109

References

Abat, M., Michael, J.M., Jason, K.K., Samuel, P.S., 2012. Adsorption and desorption of copper

and zinc in tropical peat soils of Sarawak, Malaysia. Geoderma 175-176, 58-63.

Abdel-Shafy, H.I., Aly, R.O., 2002. Water Issue in Egypt: Resources, Pollution and Protection

Endeavors. CEJOEM, 8 (1) 3-21.

Abollino, O., Aceto, M., Malandrino, M., Sarzanini, C., Mentasti, E., 2003. Adsorption of HMs

on Na-montmorillonite. Water Res., 37, 1619- 1627.

Afsin, B., Caglar, B., Tabak, A., Eren, E., 2009. Characterization of the cation-exchanged

bentonites by XRPD, ATR, DTA/TG analyses and BET measurement. Chemical

Engineering Journal 149, 242-248.

Agha, A., Ferrell, E., Hart, F., Abu El Ghar, S., Abdel-Motelib, A., 2013. Mineralogy of Egyptian

bentonitic clays II: Geologic Origin. Clays and clay minerals. 61 (6), 551-565.

Ahlberg, G., Gustafsson, O., Wedel, P., 2006. Leaching of metals from sewage sludge during

one year and their relationship to particle size. Environmental Pollution 144, 545-553.

Ahmed, A.M., Naser, L.E., 2004. Factors controlling the chemistry and mineralogy of selected

soil types of the Czech Republic and Egypt. Bulletin of Geosciences, 79 (1), 71-79.

Akpomie, K.G., Dawodu, F.A., 2015. Potential of a low-cost bentonite for heavy metal

abstraction from binary component system. Beni-suef university journal of basic and

applied sciences 4, 1-13.

Alfarra, A., Kemp-Benedict, E., Hötzl, H., Sader, N., Sonneveld, B., 2011. A Framework for

Wastewater Reuse in Jordan: Utilizing a Modified Wastewater Reuse Index. Water

Resour Manage, 25, 1153-1167.

Alloway, B.J., 1995. Heavy metals in soils. 2nd ed. John Wiley & Sons, New York, p. 3-10.

Al-Qunaibit, M.H., Mekhemer, W.K., Zaghloul, A.A., 2004. The adsorption of Cu(II) ions on

bentonite-a kinetic study. J. Colloid Interf. Sci. 283 (2), 316-321.

Alvim Ferraz, M., Lourenco, N., 2000. The influence of organic matter content of contaminated

soils on the leaching rate of heavy metals. Environ. Prog. 19, 53-58.

Anderson, R., Christensen, H., 1988. Distribution coefficients of Cd, Co, Ni and Zn in soils. J.

Soil Sci. 39, 15-22.

Antoniadis, A., McKinley, J.D., Zuhairi, W.W., 2007a. Single-element and competitive

metal mobility measured with column infiltration and batch tests. J. Environ. Qual. 36,

53-60.

Antoniadis, V., Tsadilas, C.D., Ashworth, D.J., 2007b. Monometal and competitive adsorption of

heavy metals by sewage sludge-amended soil. Chemosphere. 68 (3), 489-494.

Aparicio, P., Ferrell, R.E., Galán, E., 2010. Mg and K exchange cation effects on the XRD

analysis of soil clays. Philosophical Magazine, 90: 17, 2373-2385.

110 | REFERENCES

Araneda, C., Basualto, C., Sapag, J., Tapia, C., Cotoras, D., Valenzuela, F., 2011. Uptake of

copper (II) ions from acidic aqueous solutions using a continuous column packed with

microcapsules containing a β-hydroxyoximic compound. Chemical Engineering

Research & Design, 89 (12), 2761-2769.

Arshad, M., Silvestre, J., Pinelli, E., Kallerhoff, J., Kaemmerer, M., Shahid, M., Pradere, P.,

Dumat, C., 2008. A field study of lead phytoextraction by various scented Pelargonium

cultivars. Chemosphere, 71 (11), 2187-2192.

Arslan, M., Aslan, Z., 2006. Mineralogy, petrography and whole rock geochemistry of the

Tertiary granitic intrusions in the eastern Pontides, Turkey. Journal of Asian Earth

Sciences, 27, 177-193.

Atanassova, I.D., 1995. Adsorption and desorption of Cu at high equilibrium concentrations by

soil and clay samples from Bulgaria. Environmental pollution 87, 17-21.

Ayari, F., Srasra, E., Trabelsi-Ayadi, M., 2007. Effect of Exchangeable Cations on the

Physicochemical Properties of Smectite. Surface Engineering and Applied

Electrochemistry, 43 (5), 369-378.

Ayman, A., Mohamed H., 2011. Hydrochemical evolution and variation of groundwater and its

environmental impact at Sohag, Egypt. Arab J. Geosci., 4, 339-352.

Babel, S., Kurniawan, T.A., 2003. Low-cost adsorbents for heavy metals uptake from

contaminated water: a review. Journal of Hazardous materials B97, 219-243.

Benjamin, M., Leckie, O., 1982. Effects of complexation by Cl, SO4, and S2O3 on adsorption

behavior of Cd on oxide surfaces. Environ. Sci. Technol. 16:162-170.

Benson, C.H., Zhai, H., Wang, X., 1994. Estimating hydraulic conductivity of clay liners. J

Geotech Eng ASCE 120(2), 366-387.

Bhattacharyya, K.G., Gupta, S.S., 2008. Adsorption of a few heavy metals on natural and

modified kaolinite and montmorillonite. A review. Advances in Colloid and Interface

Science 140, 114-131.

Bhattacharyya, K., Naiya, T.K., Mandal, S.N., Da, S.K., 2008. Adsorption, kinetic and

equilibrium studies on removal of Cr(VI) from aqueous solution using different low-

cost adsorbents. Chem Eng J, 137, 529-541.

Biscaye, P.E., 1965. Mineralogy and sedimentation of recent deep-sea clays in the Atlantic Ocean

and adjacent seas and oceans. Geol. Soc. Am. Bull., 76, 803-832.

Bohn, H.L., B.L. McNeal, and G.A. O’Connor. 1985. Soil Chemistry. 2nd edition. Wiley

Interscience Publication, John Wiley and Sons, New York.

Bolland, M.D., Posner, A.M., Quirk, J.P., 1976. Surface charge on kaolinites in aqueous

suspension. Australian Journal of Soil Research, 14, 197-216.

Borisover, M., Gerstl, Z., Burshtein, F., Yariv, S., Mingelgrin, U., 2008. Organic sorbate

organoclay interactions in aqueous and hydrophobic environments: sorbate-water

competition. Environ. Sci. Technol. 42, 7201-7206.

Bouchet, A., Proust, D., Meunier, A., Beaufort, D., 1988. High-charge to low-charge smectite

reaction in hydrothermal alteration processes Clay Minerals, 23, 133-146.

Boulding, R., 1996. EPA, environmental assessment Sourcebook. Chelsea, MI: Ann Arbor Press.

Bradl, B., 2004. Adsorption of heavy metal ions on soils and soils constituents. Journal of colloid

and interface science 277, 1-18.

REFERENCES | 111

Brindley, G.W., Brown, G., 1980. Crystal structure of clay minerals and their X-ray

identification. Mineralogical Society, Monograph No. 5, London, 495 p.

Brun, L.A., Maillet, J., Hinsinger, P., Pépin, M., 2001. Evaluation of copper availability to plants

in copper-contaminated vineyard soils. Environ Pollut 111:293-302

Brusewitz, A.M., 1986. Chemical and physical properties of Paleozoic potassium bentonites from

Kinnekulle, Sweden: Clays & Clay Minerals 34, 442-454.

Carretero, M., Gomes, C., Tateo, F., 2006. Clays and human health. In: Bergaya, F., Theng,

B.K.G., Lagaly, G. (Eds.), Handbook of Clay Science. Elsevier, Amsterdam, 717-741.

Cavallaro, N., McBride, M.B., 1984. Zinc and copper sorption and fixation by an acid soil clay:

effect of selective dissolutions. Soil Sci. Soc. Am. J. 48, 1050-1054.

Cecchi, M., Dumat, C., Alric, A., Felix-Faure, B., Pradère, P., Guiresse, M., 2008. Multi-metal

contamination of a calcic cambisol by fallout from a lead-recycling plant. Geoderma.

144, 287-298.

Chairidchai, P., Ritchie, P., 1990. Zinc adsorption by a lateritic soil in the presence of organic

ligands. Soil Sci. SOC. Am. J. 54, 1242-1248.

Chamley, H., 1989. Clay sedimentology. Berlin, Springer-Verlag, 623 pp.

Chen, H., Zhao, Y., Wang, A., 2007. Removal of Cu(II) from aqueous solution by adsorption onto

acid-activated palygorskite. Journal of Hazardous Materials 149, 346-354.

Chen, J.P., 2012. Decontamination of Heavy Metals: Processes, Mechanisms, and Applications.

CRC Press. Boca Raton, pp. 291.

Chen, Z., Gu, J., Salem, A., 2012. Lagoons of the Nile delta, Egypt, heavy metal sink: With a

special reference to the Yangtze estuary of China. Estuarine, Coastal and Shelf Science

xxx (2012) 1-11.

Chin, Y.P., Aiken, G.R., Danielsen, K.M., 1997. Binding of pyrene to aquatic and commercial

humic substances: the role of molecular weight and aromaticity. Environ. Sci. Technol.

31 (6), 1630-1635.

Chorover, J., Amistadi, M., 2001. Reaction of forest floor organic matter at goethite, birnessite

and smectite surfaces. Geochimica et Cosmochimica Acta, 65, (1), 95-109.

Colombani, N., Mastrocicco, M., Di Giuseppe, D., Faccini, B., Coltorti, M., 2015. Batch and

column experiments on nutrient leaching in soils amended with Italian natural zeolitites.

CATENA. 127, 64-71.

Covelo, E.F., Andrade, M.L., Vega, F.A., 2004a. Heavy metal adsorption by humic umbrisols:

selectivity sequences and competitive sorption kinetics. Journal of Colloid and Interface

Science 280,1-8.

Covelo, E.F., Andrade, M.L., Vega, F.A., 2004b. Competitive adsorption and desorption of

cadmium, chromium, copper, nickel, lead and zinc by Humic Umbrisols. Comm. Soil

Sci. Plant Anal., 35, pp. 2709-2729.

Cowan, C.E., Zachara, J.M., Resch, C.T., 1991. Cadmium adsorption on iron oxides in the

presence of alkaline earth elements. Environ. Sci. Technol. 25, 437-446.

Cruz-Guzman, M., Celis, R., Hermosin, M.C., Koskinen, W.C., Nater, E.A., Cornejo, J., 2006.

Heavy metal adsorption by montmorillonites modified with natural organic cations. Soil

Sci. Soc. Am. J. 70, 215-221.

112 | REFERENCES

Dan’azumi, S., Bichi, M.H., 2010. Industrial pollution and heavy metals profile of Challawa river

in Kano, Nigeria. J. Appl. Sci. Environ. Sanit., 5, 56-62.

Daniel, D.E., 1998. Landfills for solid and liquid wastes. Environmental Geotechnics, P.S. Seco e

Pinto (Ed.), Balkema, Rotterdam, 4, 1231-1246.

Das, N., 2010. Recovery of precious metals through biosorption: a review. Hydrometallurgy

103:180-189.

Dawodu, F.A., Akpomie, G.K., Abuh, M.A., 2012. Batch sorption of Pb(II) from aqueous

solution by ekulu clay-equilibrium, kinetic and thermodynamic studies. Int J Multidisc

Sci Eng, 3 (10) 32-37.

Dixon, J.B., Weed, S.B., 1989. Minerals in Soil Environments, 2nd edn. Madison, WI, Soil

Science Society of America.

Donat, R., Akdogan, A., Erdem, E., Cetisli, H., 2005. Thermodynamics of Pb2+ and Ni2+

adsorption onto natural bentonite from aqueous solutions. Journal of Colloid and

Interface Science, 286, 43-52.

Doula, M., Ioannou, A., Dimirkou, A., 1999. Influence of ionic strength and pH on Cu2+

adsorption and on Mg2+, Ca2+, Mn2+, Zn2+ release by kaolinite. Agrochimica 43, 215-

222.

Dubinin, M.M., Radushkevich, V.L., 1947. Dokl. Akad. Nauk SSSR, Comm. USSR Acad. Sci.,

55, 33I (in Russian).

Eberl, D., Velde, B., 1989. Beyond the Kubler index. Clay Miner., 24 (4), 571-577.

Echeverría, J.C., Morera, M.T., Mazkiarán, C., Garrido, J., 1999. Characterization of the porous

structure of soils: adsorption of nitrogen (77 K) and carbon dioxide (273 K), and

mercury porosimetry. Eur. J. Soil Sci. 50, 497-503.

Ehrmann, W., Schmiedl, G., Hamann, Y., Kuhnt, T., Hemleben, C., Siebel, W., 2007. Clay

minerals in late glacial and Holocene sediments of the northern and southern Aegean

Sea. Palaeogeogr. Palaeoclimatol. Palaeoecol. 249, 36-57.

Eisenhour, D., Brown, K., 2009. Bentonite and its impact on modern life. Elements 5, 83-88.

El-Badry, A.A., 2016. Distribution of Heavy Metals in Contaminated Water and Bottom Deposits

of Manzala Lake, Egypt. J Environ Anal Toxicol, 6:1.

Elliott, A., Liberati, R., Huang, P., 1986. Effects of iron oxide removal on heavy metal sorption

by acid subsoils. Water Air Soil Pollut. 27, 379-389.

El-Shahat, A., Ayyad, N., Abdalla, A., 1997. Pliocene facies and fossil contents of Qaret El-

Muluk formation at Wadi El-Natrun depression, Western Desert, Egypt. Jornal of

Facies. 37 (1), 211-224.

Erdem, E., Karapinar, N., Dona, R., 2004. The removal of heavy metal cations by natural zeolites.

Journal of Colloid and Interface Science 280, 309-314.

Essington, J., 2004. Soil and Water Chemistry: An integrative approach. CRC Press, New York.

Eusterhues, K., Rumpel, C., Kögel-Knabner, I., 2005. Organo-mineral associations in sandy acid

forest soils: importance of specific surface area, iron oxides and micropores. Eur. J. Soil

Sci. 56, 753-763.

Feller, C., Schouller, E., Thomas, F., Rouiller, J., Herbillon, J., 1992. N2-BET specific surface

areas and their relation- ships with secondary constituents and organic matter contents.

Soil Sci. 153 (4), 293-299.

REFERENCES | 113

Feng, Q., Zhang, Z., Ma, Y., He, X., Zhao, Y., Chai, Z., 2012. Adsorption and desorption

characteristics of arsenic onto ceria nanoparticles. Nanoscale Res Lett, 7, p. 84.

Feng, J., Simpson, A.J., Simpson, M.J., 2005. Chemical and mineralogical controls on humic acid

sorption to clay mineral surfaces. Organic Geochemistry 36, 1553-1566.

Fernandez, M.A., Soulages, O.E., Acebal, S.G., Rueda, E.H., Sanchez, R.T., 2015. Sorption of

Zn(II) and Cu(II) by four Argentinean soils as affected by pH, oxides, organic matter

and clay content. Environ Earth Sci. 74, 4201-4214.

Figueira, M., Volesky, B., Azarian, K., Ciminelli, V., 2000. Biosorption column performance

with a metal mixture. Environ. Sci. Technol. 34, 4320-4326.

Fu, F., Wang, Q., 2011. Removal of heavy metal ions from wastewaters: A review. J. Environ.

Manage., 92, 407-418.

Fujiyoshi, R., Okamoto, T., Katayama, M., 1994. Behavior of radionuclides in the environment-

II. Application of sequential extraction to Zn(II) sorption studies. Appl. Radiat. Isotopes

45 (2), 165-170.

Gaber, A., 1994. Wastewater Innovative and Alternative Technology Overview. Chemonics

Egypt Consultants.

Gao, S., Walker, J., Dahlgren, A., Bold, J., 1997. Simultaneous sorption of Cd, Cu, Ni, Zn, Pb,

and Cr on soils treated with sewage sludge supernatant. Water, Air, and Soil pollution,

93, 331-345.

Ghandour, M., Abd El-Hameed, T., Faris, M., Marzouk, A., Maejima, W., 2004. Textural,

Mineralogical and Microfacies Characteristics of the Lower Paleogene Succession at

the Nile Valley and Kharga Oasis Regions, Central Egypt. Journal of Geosciences, 47

(4), 39-53.

Glover, J., Mathew, H., Eick, J., Brady, V., 2002. Desorption kinetics of cadmium and lead from

goethite: Influence of time and organic acids’, Soil Sci. Soc. Am. J. 66, 797-804.

Gomes, C., Fontes, F., Da Silva, G., Mendonça, S., Netto, R., 2001. Selectivity sequence and

competitive adsorption of heavy metals by Brazilian soils. Soil Science Society of

America Journal, 65, 1115-1121.

Gomes, C., Silva, J., 2007. Mineral and clay minerals in medical geology. Applied Clay Science

36, 4-21.

Gray, B., Nickelsen, P., 1989. Pedogenic slickensides, indicators of strain and deformation

processes in red bed sequences of the Appalachian foreland. Geology, 17, 72-75.

Griffin, L., 2002. Aridity and humidity: two aspects of the late Miocene climate of North Africa

and the Mediterranean, Palaeoogeogr. Palaeoclimatol. Palaeoecol., 182, 65-91.

Griffin, J., Windom, H., Goldberg, D., 1968. The distribution of clay minerals in the World

Ocean. Deep-Sea Res., 15: 433-459.

Grim, R.E., 1968. Clay mineralogy. 2nd. ed. McGraw-Hill, New York.

Gu, X., Evans, J., Barabash, J., 2010. Modeling the adsorption of Cd, Cu, Ni, Pb and Zn onto

montmorillonite. Geochem. Cosmochim. Acta 74, 5718-5728.

Gupta, K., Carrott, M., Ribeiro Carrott, L., Suhas, L., 2009. Low-cost adsorbents: growing

approach to wastewater treatment-a review. Crit. Rev. Environ. Sci. Technol., 39 (10),

783-842.

114 | REFERENCES

Händel, M., Rennert, T., Totsche, U., 2013. A simple method to synthesize birnessite at ambient

pressure and temperature. Geoderma, 193-194, 117-121.

Hashim, A., Mukhopadhyay, S., Sahu, N., Sengupta, B., 2011. Remediation technologies for

heavy metal contaminated groundwater. J. Environ. Manag., 92, 2355-2388.

Hassan, M., El-Haddad, A., Omer, A., Ibrahim, M., 2005. Geochemical characteristics of the

surficial Nile basin sediments and their environmental relevance, Sohag area, Egypt.

M.Sc. thesis, South Valley University (Sohag), Egypt.

Harter, D., 1992. Competitive sorption of cobalt, copper, and nickel ions by a calcium saturated

soil. Soil Sci. Soc. Am. J., 56: 444-449.

Harter, D., Naidu, R., 1995. Role of metal-organic complexation metal sorption by soils.

Advances in Agronomy, vol. 55- Academic Press, Inc.

Heister, K., 2014. The measurement of the specific surface area of soils by gas and polar liquid

adsorption methods-Limitations and potentials. Geoderma. 216, 75-87.

Hendershot, H., Duquette, M., 1986. A simple barium chloride method for determining cation

exchange capacity and exchangeable cations. Soil Sci. Soc. Am. J. 50, 605-608.

Hendriks, F., 1985. Upper Cretaceous to lower Tertiary sedimentary environmentsand clay

mineral associations in the Kharga Oasis area, Egypt. N. jb. Geo. Paliiont. Mh., 10, 579-

591.

Henry, P., 1997. Relationship between porosity, electrical conductivity and cation exchange

capacity in Barbados wedge sediments, Proc. Ocean Drill. Program Sci. Results, 156,

137-149.

Hoffman, J., Hower, J., 1979. Clay mineral assemlages as low grade metamorphic

geothermometers: application to the thrust faulted disturbed belt of Montana, USA. In:

Aspects of Diagenesis (Scholle, P.A. and Schluger, P.S., Eds.), Special PubIs., Soc.

Econom. Paleont.Miner., 26, 55-79.

Holmgren, S., 1967. A rapid citrate-dithionite-extractable iron procedure. Soil Sci. Soc. Am. Proc.

31:210-211.

Hong, H., Churchman, J., Gu, Y., Yin, K., Wang, C., 2012. Kaolinite-smectite mixed-layer clays

in the Jiujiang red soils and their climate significance. Geoderma 173-174, 75-83.

Houba, M., Temninghoff, M., Garikhorst, A., Van Vark, W., 2000. Soil analysis procedures using

0.01 M calcium chloride as extraction reagent. Communications in Soil Science and

plant analysis 31: 1299-1396.

Hower, J., Mowatt, C., 1966. The mineralogy ofillites and mixed-layer illite/montmorillonites:

Amer. Mineral 51, 825-854.

Hu, J., Shipley, J., 2012. Evaluation of desorption of Pb(II), Cu(II) and Zn(II) from titanium

dioxide nanoparticles. Sci Total Environ. 431:209-220.

Hu, J., Shipley, J., 2013. Regeneration of spent TiO2 nanoparticles for Pb (II), Cu (II), and Zn (II)

removal. Environ Sci Pollut Res Int.; 20(8), 5125-37.

Hur, J., Williams, M., Schlautman, M., 2006. Evaluation spectroscopic and chromatographic

techniques to resolve dissolved organic matter via end member mixing analysis.

Chemosphere 63, 387-402.

REFERENCES | 115

Ikhsan, J., Wells, J.D., Johnson, B.B., Angove, M.J., 2005. Surface complexation modeling of the

sorption of Zn(II) by montmorillonite. Colloids and Surfaces A: Physicochemical and

Engineering Aspects. 252, (1), 33-41.

Inglezakis, V.J., Loizidou, M.D., Grigoropoulou, H.P., 2003. Ion exchange of Pb2+, Cu2+, Fe3+,

and Cr3+ on natural clinoptilolite: selectivity determination and influence of acidity on

metal uptake. J. Colloid Interface Sci., 261, p. 49.

Inoue, A., Utada, M., 1983. Further investigations of a conversion series of dioctahedral

mica/smectites in the Shinzan hydrotherrnal alteration area northeast Japan: Clays &

Clay Minerals, 31, 401-412.

Ismaiel, H., 2013. Cement Kiln Dust Chemical Stabilization of Expansive Soil. Exposed at El-

Kawther Quarter, Sohag Region, Egypt. International Journal of Geosciences, 4, 1416-

1424.

Issawi, B., Hassan, W., Osman, R., 1978. Geological studies in the area of Kom Ombo, Eastern

Desert, Egypt. Ann. Geol. Survey, Egypt, V.VIII, 187-235.

Jamil, T.S., Ibrahim, H.S., Abd El-Maksoud, I.H., El-Wakeel, S.T., 2010. Application of zeolite

prepared from Egyptian kaolin for removal of heavy metals: I. Optimum conditions.

Desalination. 258 (1-3), 34-40.

Jeon, C., Park, J.Y., Yoo, Y.J., 2001. Removal of heavy metals in plating wastewater using

carboxylated alginic acid. Korean J. Chem. Eng., 18, p. 955.

Jackson, M.L., 1969. Soil Chemical Analysis-Advanced Course. Soil Science Department,

Wisconsin University, Madison. Published by author.

Jacobs, B.F., Kingston, J.D., Jacobs, L.L., 1999. The origin of grass-dominated ecosystems.

Annals of the Missouri Botanical Garden 86, 590-643.

Jansen, B., Nierop, K.J., Verstraten, J.M., 2003. Mobility of Fe(II), Fe(III) and Al in acidic forest

soils mediated by dissolved organic matter: Influence of solution pH and metal/organic

carbon ratios. Geoderma, 113(3-4), 323-340.

Jansen, B., Nierop, J., Verstraten, M., 2004. Mobilization of dissolved organic matter, aluminium

and iron in podzol eluvial horizons as affected by formation of metal-organic

complexes and interactions with solid soil material. European Journal of Soil Science,

55, 287-297.

Jansen, B., Nierop, J., Verstraten, M., 2005. Mechanisms controlling the mobility of dissolved

organic matter, aluminium and iron in podzol B horizons. European Journal of Soil

Science, 56, 537-550.

Järup, L., 2003. Hazards of heavy metal contamination. British Medical Bulletin 2003; 68: 167-

182.

Johri, N., Jacquillet, G., Unwin, R., 2010. Heavy metal poisoning: the effects of cadmium on the

kidney. Biometals. 23:783-92.

Jones, A., Breuning-Madsen, H., Brossard, M., Dampha, A., Deckers, J., Dewitte, O., Gallali, T.,

Hallett, S., Jones, R., Kilasara, M., Le Roux, P., Micheli, E., Montanarella, L.,

Spaargaren, O., Thiombiano, L., Van Ranst, E., Yemefack, M. , Zougmoré R., 2013.

Soil Atlas of Africa. European Commission, Publications Office of the European

Union, Luxembourg. 176 pp.

116 | REFERENCES

Jones, K.P.N., McCave, I.N., Patel, D., 1988. A computer-interfaced SediGraph for modal size

analysis of fine grained sediment. Sedimentology, 35(1), 163-172.

Kadirvelu, K., Thamaraiselvi, K., Namasivayam,C., 2001. Removal of heavy metals from

industrial wastewaters by adsorption onto activated carbon prepared from an

agricultural solid waste. Bioresource Technology 76, 63-65.

Kaiser, K., Guggenberger, G., 2000. The role of DOM sorption to mineral surfaces in the

preservation of organic matter in soils. Org. Geochem. 31, 711-725.

Kaiser, K., Guggenberger, G., 2003. Mineral surfaces and soil organic matter. Eur. J. Soil Sci. 54

(2), 219-236.

Kaiser, K., Guggenberger, G., Zech, W., 1996. Sorption of DOM and DOM fractions to forest

soils. Geoderma 74, 281-303.

Kaiser, K., Zech, W., 1998. Soil DOM sorption as infuenced by organic and sesquioxide coatings

and sorbed sulfate. Soil Science Society of America Journal, 62, 129-136.

Kalbitz, K., Schwesig, D., Rethemeyer, J., Matzner, E., 2005. Stabilization of dissolved organic

matter by sorption to the mineral soil. Soil Biol. Biochem. 37, 1319-1331.

Kalbitz, K., Wennrich, R., 1998. Mobilization of heavy metals and arsenic in polluted wetland

soils and its dependence on dissolved organic matter. Science of the Total Environment,

209 (1): 27- 39.

Kandpal, G., Srivastava, C., Ram, B., 2005. Kinetics of desorption of heavy metals from polluted

soils: Influence of soil type and metal source. Water Air Soil Pollut. 161, 353.

Karakaya, C., Karakaya, N., Bakır, S., 2011. Some properties and potential applications of the

Na- and Ca-bentonites of ordu (N.E. Turkey). Applied Clay Science 54, 159-165.

Karathanasis, D., 1999. Subsurface migration of copper and zinc mediated by soil colloids. Soil

Sci. Soc. Am. J., 63, pp. 830-838.

Karlsson, T., Persson, P., Skyllberg, U., 2006. Complexation of copper(ll) in organic soils and in

dissolved organic matter--EXAFS evidence for chelate ring structures. Environ Sci

Technol. 2006 Apr 15;40(8):2623-2628.

Kaufhold, S., Dohrmann, R., Klinkenberg, M., Siegesmund, S., Ufer, K., 2010. N2-BET specific

surface area of bentonites. J. Colloid Interface Sci. 349, 275-282.

Kayabali, K., 1997. Engineering aspects of a novel landfill liner material: bentonite amended

natural zeolite. Eng Geol 46:105-114.

Khan, R., Bolan, S., MacKay, D., 2005. Adsorption and Desorption of Copper in Pasture Soils.

Communications in Soil Science and Plant Analysis, 36: 2461-2487, 2005.

Klaine, J., Alvarez, J., Batley, E., Fernandes, F., Handy, D., Lyon, Y., Mahendra, S., McLaughlin,

M.J., Lead, J.R., 2008. Nanomaterials in the environment: behaviour, fate,

bioavailability and effects. Environ. Toxicol. Chem., 27, 1825-1851.

Kothawala, D.M., Moore, T.R., Hendershot, W.H., 2009. Soil properties controlling the

adsorption of dissolved organic carbon to mineral soils. Soil Sci Soc Am J., 73, 1831-

1842.

Kowalkowskia, T., Tutu, H., Cozmuta, L., Sprynskyy, M., Cukrowska, E., Buszewski, B., 2010.

Assessment of mobility of heavy metals in two soil types by use of column leaching

experiments and chemometric evaluation of elution curves. Intern. J. Environ. Anal.

Chem. 90 (10), 797-811.

REFERENCES | 117

Kumar, D., Alpana, S., Gaur, P., 2012. Continuous metal removal from solution and industrial

effluents using Spirogyra biomass-packed column reactor. Water Research 46, 779-788.

Kwon, S., Pignatello, J., 2005. Effect of natural organic substances on the surface and adsorptive

properties of environmental black carbon (char): pseudo pore blockage by model lipid

components and its implications for N2-probed surface properties of natural sorbents.

Environ. Sci. Technol. 39, 7932-7939.

Lair, J., Gerzabek, H., Haberhauer, G., 2007. Sorption of heavy metals on organic and inorganic

soil constituents. Environ. Chem. Lett. 5, 23-27.

Lata, S., Singh, K., Samadder, R., 2015. Regeneration of adsorbents and recovery of heavy

metals: a review. Int. J. Environ. Sci. Technol., 12, 1461-1478.

Lewis, A., Berry, L., 1988. African Environments and Resources. Boston: Unwin Hyman, p.41.

Li, Y., Wang, L., Huang, H., Zhang,Y., Guo, H., 2009. Adsorption of Cu and Zn onto Mn/Fe

oxides and organic materials in the extractable fractions of River surficial sediments.

Soil & Sediment Contamination, 18, 87-101.

Lin, H., Lee, H., Jian, H., Wong, M., Shieh, J., Wang, Y., 2002. A study of purified

montmorillonite intercalated with 5-fluorouracil as drug carrier. Biomaterials, 23, 1981-

1987.

Lin, H., Juang, S., 2002. Heavy metal removal from water by sorption using surfactant-modified

montmorillonite, J. Hazard. Mater. B 92, 315-326.

Liu, A., Gonzales, D., 1999. Adsorption/desorption in a system consisting of humic acid, heavy

metals, and clay minerals. Journal of colloid and interface science, 218, 225-232.

Lodeiro, P., Herrero, R., Sastre de Vicente, E., 2006. Batch desorption studies and multiple

sorption-regeneration cycles in a fixed-bed column for Cd(II) elimination by

protonated Sargassum muticum. Journal of Hazardous Materials. 137(3), 1649-1655.

Lu, S.G., Xu, Q.F., 2009. Competitive adsorption of Cd, Cu, Pb and Zn by different soils of

Eastern China. Environmental Geology, 57 (3), 685-693.

Ma, H., Kim, D., Cha, K., Allen, E., 1999. Effect of kinetics of complexation by humic acidon the

toxicity of copper to Ceriodaphnia dubia. Environ Toxicol Chem., 18, 828-37.

Macklin, M.G., Brewer, P.A., Hudson-Edwards, K.A., Bird, G., Coulthard, T.J., Dennis, I.A.,

Lechler, P.J., Miller, J.R., Turner, J.N., 2006. A geomorphological approach to the

management of rivers contaminated by metal mining. Geomorphology, 79 (3-4), 423-

447.

Makino, T., Sugahara, K., Sakurai, Y., Takano, H., Kamiya, T., Sasaki, K., Itou, T., Sekiya, N.,

2006. Restoration of cadmium contamination in paddy soils by washing with chemicals:

selection of washing chemicals. Environ. Pollut., 144 (1), 2-10.

Martin, C.W., 2000. Heavy metal trends in floodplain sediments and valley fill, River Lahn,

Germany. Catena, 39, 53-68.

Maszkowska, J., Kołodziejska, M., Białk-Bielińska, A., Mrozik, W., Kumirska, J., Stepnowski,

P., Palavinskas, R., Krüger, O., Kalbe, U., 2013. Column and batch tests of sulfonamide

leaching from different types of soil. Journal of Hazardous Materials. 260, 468-474.

McKenzie, R.M., 1980. The adsorption of lead and other heavy metals on oxides of manganese

and iron. Aust J Soil Res, 18(1), 61-73.

118 | REFERENCES

McKnight, D., Bencala, K.E., Zellweger, G.W., Aiken, G.R., Feder, G.L., Thorn, K.A., 1992.

Sorption of dissolved organic carbon by hydrous aluminum and iron oxides occurring at

the confluence of Deer Creek with the Snake River, Summit County, Colorado.

Environmental Science and Technology 26, 1388-1396.

Miller, J.R. , Turner, J.N., 2006. A geomorphological approach to the management of rivers

contaminated by metal mining. Geomorphology, 79, 423-447.

Momodu, M.A., Anyakora, C.A., 2010. Heavy metal contamination of ground water: the Surulere

case study research. J. Environ. Earth Sci., 2 (1), 39-43.

Marzouk, A.M., 1985. Sedimentological and stratigraphical studies on the Upper Cretaceous-

Lower Tertiary succession near Qena, Egypt. Unpublished M.Sc. Thesis, Tanta

University, Egypt, 170 pp.

McBride, M.B, 1989. Reactions controlling heavy metal solubility in soils. Advances in Soil

Science, 10, 1-57.

McBride, M.B, 1998. Soluble trace metals in alkaline stabilized sludge products. J. Environ. qual.

27, 578-584.

McBride, M.B., 1994. Environmental chemistry of soils. Oxford University Press, Oxford, United

Kingdom, 406p.

McBride, M.B., Richards, B.K., Steenhuis, T.S., Russo, J.J., Sebastien, S., 1997. Mobility and

solubility of toxic metals and nutrients in soil fifteen years after sludge application. Soil

Sci. 162, 487-500.

McBride, M.B., Richards, B.K., Steenhuis, T.S., Spiers, G., 1999. Long-term leaching of trace

elements in a heavily sludgeamended silty clay loam soil. Soil Sci. 164, 612-613.

Meers, E., Samson, R., Tack, G., Ruttens, A., Vandegehuchte, M. 2007. Phytoavailability

assessment of heavy metals in soils by single extractions and accumulation by

Phaseolus vulgaris, Environmental and Experimental Botany 60, 385-396.

Mehra, O., Jackson, M., 1960. Iron oxide removal from soils and clays by a dithionite citrate

system buffered with sodium bicarbonate. Clays Clay Miner. 7, 313-317.

Mehta, S.K., Gaur, J.P., 2005. Use of algae for removing heavy metal ions from wastewater:

progress and prospects. Critical Reviews in Biotechnology 25, 113e152.

Meleshyn, A., Tunega, D., 2011. Adsorption of phenanthrene on Na-montmorillonite: A model

study. Geoderma, 169, 41-46.

Mellah, A., Chegrouche, S., 1997. The removal of zinc from aqueous solutions by natural

bentonite. Water Research, 31, 621-629.

Meng, M., Wang, Z., Ma, L., Zhang, M., Wang, J., Dai, X., Yan, Y., 2012. Selective adsorption

of methylparaben by submicrosized molecularly imprinted polymer: batch and dynamic

flow mode studies. Industrial & Engineering Chemistry Research, 51 (45), 14915-

14924.

Meunier, A.,Velde, B., 2004. Illite: Origins, Evolution and Metamorphism. Springer, Berlin.

Mikutta, R., Mikutta, C., Kalbitz, K., Scheel, T., Kaiser, K., Jahn, R., 2007. Biodegradation of

forest floor organic matter bound to minerals via different binding mechanisms.

Geochimica et Cosmochimica Acta, 71, 2569-2590.

REFERENCES | 119

Milczarek, M.A., 1994. Sorption and Transport Characteristics of Cadmium and Dissolved

Natural Organic Matter with a Coarse Loamy Soil. MSc Thesis, University Arizona,

Tucson, AZ.

Mitchell, J.K., 1976. Fundamentals of Soil Behavior. John Wiley and Sons, Inc., New York.

Moore, D.M., Reynolds, R.C., 1997. X-ray diffraction and the identification and analysis of clay

minerals, 2nd ed.: Oxford University Press, Oxford, 332 pp.

Moreno-Castilla, C., Rivera-Utrilla, J., 2001. Carbon materials as adsorbents for the removal of

pollutants from the aqueous phase, Mater. Res. Soc. Bull. 26, 890-894.

Motsi, T., Rowson, A., Simmons, H., 2009. Adsorption of heavy metals from acid mine drainage

by natural zeolite. Int. J. Miner. Process. 92, 42-48.

Murray, H., 1999. Applied clay mineralogy today and Tomorrow. Clay Minerals, 34, 39-49.

Murray, H., 2000. Traditional and new applications for kaolin, smectite, and palygorskite: a

general overview. Applied Clay Science 17, 207-221.

Mustafa, G., Singh, B., Kookana, R.S., 2004. Cadmium adsorption and desorption behaviour on

goethite at low equilibrium concentrations: effects of pH and index cations.

Chemosphere, 57, 1325-1333.

Neto, F., Vieira, A., Silva, C., 2012. Cu(II) adsorption on modified bentonitic clays: different

isotherm behaviors in static and dynamic systems. Mater. Res., 15, 114-124.

Nodvin, S.C., Driscoll, C.T., Likens, G.E., 1986. Simple partitioning of anions and dissolved

organic carbon in a forest soil. Soil Science 142, 27-35.

Norvell, W.A., Lindsay, W.L., 1972. Reactions of DTPA chelates of iron, zinc, copper and

manganese with soils. Soil Science Society of America Journal, 36, 778-783.

Nouri, J., Mahvi, A.H., Jahed, G.R., Babaei, A.A., 2008. Regional distribution pattern of

groundwater heavy metals resulting from agricultural activities. Environmental

Geology, 55, 1337-1343.

Odoma, A.N., Obaje, N.G., Omada, J.I., Idakwo, S.O., Erbacher, J., 2013. Paleoclimate

Reconstruction during Mamu Formation (Cretaceous) Based on Clay Mineral

Distributions. IOSR Journal of Applied Geology and Geophysics (IOSR-JAGG). 1 (5),

40-46.

Omer, A., 1996. Geological, mineralogical and geochemical studies on the Neogene and

Quaternary Nile basin deposits, Qena-Assiut stretch, Egypt. Ph.D. Thesis, Geology

Dept., Fac. Sci. Sohag, South Valley University. 320 p.

Omer, A., Issawi, B., 1998. Lithostratigraphical, mineralogical andgeochemical studies on the

Neogene and Quaternary Nile basin deposits, Qena-Assiut stretch, Egypt. The 4th

International Conference on Geology of the Arab World, Cairo (Abstract).

Panday, K., Parsed, G., Singh, N., 1985. Copper(II) removal from aqueous solutions by fly ash.

Water Res., 19, p. 869.

Parfitta, R., Giltrapa, D., Whittona, J., 1995. Contribution of organic matter and clay minerals to

the cation exchange capacity of soils. pages 1343-1355. 26 (9-10).

Park, J.H., Cho, S., Ok, .S., Kim, H., Kang, W., Choi, W., Heo, S., DeLaune, D., Seo, C., 2015.

Competitive adsorption and selectivity sequence of heavy metals by chicken bone-

derived biochar: Batch and column experiment. Journal of Environmental Science and

Health, Part A 50, 1194-1204.

120 | REFERENCES

Peacock, C.L., Sherman, D.M., 2004. Copper (II) sorption onto goethite, hematite and

lepidocrocite: a surface complexation model based on ab initio molecular geometries

and EXAFS spectroscopy. Geochimica et Cosmochimica Acta 68, 2623-2637.

Pedro, G., 1994. Clay mineral in weathered rock materials and soils. In: Paquat, H., Clauer, N.

(Eds.), Soil and Sediments: Mineralogy and Geochemistry. Elsevier-Verlag, Berlin, pp.

1-20.

Petersen, L.W., Moldrup, P., Jacobsen, O.H., Rolston, D.E., 1996. Relations between specific

surface area and soil physical and chemical properties. Soil Sci. 161, 9-21.

Pierangeli, P., Guilherme, G., Oliveira, R., Curi, N., Silva, N., 2003. Efeito da força iônica da

solução de equilíbrio na adsorção de cádmio em Latossolos brasileiros. Pesquisa

Agropecuária Brasileira (Effect of ionic strength of the solution balance in cadmium

adsorption in Brazilian Oxisols). Brazilian Agricultural Research, 38, 737-745.

Piirsoo, K., Viik, M., Kõiv, T., Käiro, K., Laas, A., Nõges, T., Pall, P., Selberg, A., Toomsalu, L.,

Vilbaste, S., 2012. Characteristics of dissolved organic matter in the inflows and in the

outflow of Lake Võrtsjärv, Estonia. Journal Hydrology, 475, 306-313.

Plassard, F., Winiarski, T., Petit-Ramel, M., 2000. Retention and distribution of three heavy

metals in a carbonated soil: comparison between batch and unsaturated column studies.

J. Contam. Hydrol., 42, 99-111.

Pueyo, M., Lopez-Sanchez, F., Rauret, G., 2004. Assessment of CaCl2, NaNO3 and NH4NO3

extraction procedures for the study of Cd, Cu, Pb and Zn extractability in contaminated

soils. Analytica Chimica Acta 504, 217-226.

Puppa, D., Komárek, M., Bordas, F., Bollinger, C., Joussein, E., 2013. Adsorption of copper,

cadmium, lead and zinc onto a synthetic manganese oxide. Journal of Colloid and

Interface Science, Elsevier, 399, 99-106.

Qin, F., Shan, X., Wei, B., 2004. Effects of low-molecular-weight organic acids and residence

time on desorption of Cu, Cd, and Pb from soils, Chemosphere 57, 253-263.

Qin, F., Wen, B., Shan, X., Xie, Y., Liu, T., Zhang, S., Khan, S., 2006. Mechanisms of

competitive adsorption of Pb, Cu, and Cd on peat. Environ. Pollut., 144, 669-680.

Radwan, M.A., Salama, A.K., 2006. Market basket survey for some heavy metals in Egyptian

fruits and vegetables. Food and Chemical Toxicology. 44 (8), 1273-1278.

Raigemborn, S., Gómez-Perala, E., Kraused, M., Matheosa, D., 2014. Controls on clay minerals

assemblages in an early paleogene nonmarine succession: Implications for the volcanic

and paleoclimatic record of extra-andean patagonia, Argentina. Journal of South

American Earth Sciences. 52, 1-23.

Rashed, M.N., Soltan, M.E., 2002. Removal of nutrients and heavy metals from urban wastewater

using aeration, alum and kaolin ore. Proceedings of International Symposium on

Environmental Pollution Control and Waste Management, pp.621-627.

Rateev, M.A., Gorbunova, Z.N., Lisitzyn, A.P., Nosov, G.L., 1969. The distribution of clay

minerals in the oceans. Sedimentology, 13:21-43.

Raucsik, B., Merenyi, L., 2000. Origin and environmental significance of clay minerals in the

Lower Jurassic formations of the Mecsek Mts., Hungary. Acta Geol. Hungarica, 43,

405-429.

REFERENCES | 121

Reed, B.E., Carriere, P.C., Moore, R., 1996. Flushing of a Pb(II) contaminated soil using HCl,

EDTA, and CaCl2. J. Environ. Eng. 122, 48-50.

Refaey, Y., El-Shater, A., El-Haddad, A., Abu Seif, E., 2008. Mineralogical and geotechnical

studies on the weathered zones of the basement rocks of Aswan area, Egypt.

Unpublished M. Sc. Thesis, Sohag University, Egypt, chapter 6, p. 81-129.

Refaey, Y., Jansen, B., El-Shater, A., El-Haddad, A., Kalbitz, K., 2014. The role of dissolved

organic matter in adsorbing heavy metals in clay-rich soils. VZJ. Vol. 13 No. 7.

Refaey, Y., Jansen, B., El-Shater, A., El-Haddad, A., Kalbitz, K., 2015. Clay minerals of Pliocene

deposits and their potential use for the purification of polluted wastewater in the Sohag

area, Egypt. Geoderma Regional. 5, 215-225.

Refaey, Y., Jansen, B., Parsons, J.R., de Voogt, P., Bagnis, S., Markus, A., El-Shater, A., El-

Haddad, A., Kalbitz, K. 2016. Effects of clay minerals, hydroxides, and timing of

dissolved organic matter addition on the competitive sorption of Copper, Nickel and

Zinc: A column experiment (Revised version, Journal of Environmental Management).

Reza, R., Singh, G., 2010. Heavy metal contamination and its indexing approach for river water.

International Journal of Environmental Science and Technology, 7, 785-792.

Rice, J., T.J., Weed, S.B., Buol, S.W., 1985. Soil-saprolite profiles derived from mafic rocks in

the North Carolina piedmont: I. Chemical, morphological, and mineralogical

characteristics and transformations. Soil Sci. Soc. Am. J., 49, 171-178.

Rowe, K., Quigley, M., Booker, R., 1995. Clayey barrier systems for waste disposal facilities. E

& FN, Spon Lond, 390 pp.

Rutherford, W., Chiou, T., Eberl, D., 1997. Effects of exchanged cation on the microporosity of

montmorillonite, Clays Clay Miner, 45: 534-543.

Said, R., 1971. Explanatory notes to accompany the geological map of Egypt. 1:2,000000.

Geol.Surv. Egypt. pp. 56, 123.

Said, R., 1981. The geological evolution of the River Nile. Springer-Verlag, NewYork. 151p.

Said, R., 1990. The geology of Egypt. Balkema, Rotterdam, 731p.

Salman, A., Melegy, A., Shaban, M., Hassaan, M., 2013. Hydrogeochemical Characteristics and

classification of Groundwater in Sohag Governorate, Egypt. Journal of Applied

Sciences Research, 9(1): 758-769.

Salem, H.M., Eweida A.E., Farag, A., 2000. Heavy metals in drinking water and the

environmental impact on human health. Egypt: ICEHM, Cairo University, 542-56.

Sanfeliu, T., Gomez, E., Alvarez, C., Hernandez, D., Martin, J., Ovejero, M., Jordan, M., 2002. A

valuation of the particulate atmospheric aerosol in the urban area of castellon, Spain. In:

E. Galan, F. Zezza (Eds.), Protection and Conservation of the Cultural Heritage of the

Mediterranean Cities, Balkem, Publishers, pp. 61-65.

Schaetzl, R.J., Anderson, S.N., 2005. Soils Genesis and Geomorphology. Cambridge, UK:

Cambridge University Press.

Schramel, O., Michalke, B., Kettrup, A., 2000. Study of the copper distribution in contaminated

soils of hop fields by single and sequential extraction procedures. Sci Total Environ

263:11-22.

Schwertmann, U., Cornell, R.M., 1991. Iron oxides in the laboratory. preparation and

Characterization. VCH Verlagsgesellschaft Weinheim. pp. 5-13.

122 | REFERENCES

Searle, P.L., Daly, B.K., 1977. The determination of aluminium, iron, manganese and silicon in

acid oxalate soil extracts by flame emission and atomic absorption spectrometry,

Geoderma, 19, 1-10.

Senesi, N., Sposito, G., Martin, J.P., 1986. Copper (II) and iron (II) complexation by soil humic

acids: An IR and ESR study. The Science of the Total Environment, 55, 351-362.

Searle, P.L., Daly, B.K., 1977. The determination of aluminium, iron, manganese and silicon in

acid oxalate soil extracts by flame emission and atomic absorption spectrometry,

Geoderma, 19, 1-10.

Senesi, N., Sposito, G., Martin, J.P., 1986. Copper (II) and iron (II) complexation by soil humic

acids: An IR and ESR study. The Science of the Total Environment, 55, 351-362.

Seo, D.C., Yu, K., De Laune, R.D., 2008. Comparison of monometal and multimetal adsorption

in Mississippi River alluvial wetland sediment: Batch and column experiments.

Chemosphere 73, 1757-1764.

Sheng, G., Sheng, J., Yang, S., Hu, J., Wang, X., 2011. Behavior and mechanism of Ni (II) uptake

on MnO2 by a combination of macroscopic and EXAFS investigation. J. Radioanal.

Nucl. Chem., 289, 129-135.

Shuman, M., Dudka, S., Das, K., 2002. Cadmium forms and plant availability in compost-

amended soil. Communications in Soil Science and Plant Analysis, 33, 737-748.

Silveira, A., Alleoni, F., Guilherme, G., 2003. Biosolids and heavy metals in soils. Sci. agric., 60

(4), 793-806.

Singer, A., 1984. The paleoclimatic interpretation of clay minerals in sediments: a review. Earth-

Sci. Rev. 21, 251-293.

Spark, K.M., Wells, J.D., Johnson, B.B., 1995. Characterizing trace metal adsorption on kaolinite.

Eur. J. Soil Sci. 46, 633-640.

Sposito, G., 1989. The Chemistry of Soils. New York, Oxford University Press, Inc.

Sprynskyy, M., Kowalkowski, T., Tutu, H., Cozmuta, L.M., Cukrowska, E.M., Buszewski, B.,

2011. The Adsorption Properties of Agricultural and Forest Soils Towards Heavy Metal

Ions (Ni, Cu, Zn, and Cd), Soil and Sediment Contamination: An International Journal,

20 (1) 12-29.

Srivastava, P., Singh, B., Angove, M., 2005. Competitive adsorption behavior of heavy metals on

kaolinite. J. Colloid Interf. Sci. 290, 28-38.

Srivastava, S.K., Tyagi, R., Pant, N., Pal, N., 1989. Studies on the removal of some toxic metal

ions. Part (II): Removal of lead and cadmium by montmorillonite and kaolinite.

Environ. Technol. Lett., 10, 275–282.

Stahl, R.S., James, B.R., 1991. Zinc sorption by manganese-oxide-coated sand as a function of

pH. Soil Sci. Soc. Am. J., 55, 1291-1294.

Stein, R., 1985. Rapid grain-size analysis of clay and silt fraction by SediGraph 500 D:

comparison with Coulter and Atterberg methods. Journal of sedimentary petrology vol.

55 (4): 590-593. Tulsa. USA.

Stevenson, J., Fitch, A., 1981. Reactions with organic matter. JF. Loneragan, AD Robson, RD

Graham (Eds.), Copper in Soils and Plants, Academic Press, Sydney, Australia, 69-95.

Stevenson, J., 1982. Humus Chemistry. Genesis, composition, reaction. New York: Wiely.

REFERENCES | 123

Stevenson, J., 1994. Humus Chemistry: Genesis, Composition, Reactions, 2nd ed., John Wiley

& Sons, New York.

Stumm, W., 1992. Chemistry of the solid-water interface. New York: John Wiley and Sons.428p.

Swezey, C.S., 2009. Cenozoic stratigraphy of the Sahara, Northern Africa. J. Afr. Earth Sci., 53,

pp. 89-121.

Taha , R., Kabir, H., 2005. Tropical residual soil as compacted soil liners. Environmental

Geology, 47:375-381.

Talaat, A., El Defrawy, M., Abulnour, G., Hani, A. Tawfik, A., 2011. Evaluation of heavy metals

removal using some Egyptian clays. International Proceedings of Chemical, Biological

and Environmental Engineering, 6, 37-42.

Talbot, R., Williams, J., 2009. Cenozoic evolution of the Nile basin. In:Dumont, H.J. (Ed.), The

Nile: Origin, Environments, Limnology and Human Use.

Tan, H., Champion, J.T., Artiola, J.F., Brusseau, M.L. and Miller, R.M., 1994. Complexation of

cadmium by a Rhamnolipid Biosurfactant. Environ. Sci. Technol., 28: 2402-2406.

Tang, C., Lo, M., 2013. Magnetic nanoparticles: essential factors for sustainable environmental

applications. Water Research, 47 (8), 2613-2632.

Tantawy, A., Keller, G., Adatte, T., Stinnesbeck, W., Kassab, A., Schulte, P., 2001. Maastrichtian

to Paleocene depositional environment of the Dakhla Formation, Western Desert,

Egypt: sedimentology, mineralogy and integrated micro- and macrofossil

biostratigraphies. eret. Res., 22, 795-827.

TEGPC, CONOCO, 1987. Geological Map of Egypt (Scale 1: 500000), sheet: NG 36 NW Assiut.

The Egyptian Organization for Human Rights, 2009. Report: Water pollution time

bomb threatening the life of the Egyptians. http://ar.eohr.org/wp -

content/uploads/2009/12/d8aad982d8b1d98ad8b1d985d98ad8a7.pdf.

Theng, B.K., Ristori, G.G., Santi, C.A., Percival, H.J., 1999. An improved method for

determining the specific surface areas of topsoils with varied organic matter content,

texture and clay mineral composition. Eur. J. Soil Sci. 50, 309-316.

Thiry, M., 2000. Paleoclimatic interpretation of clay minerals in marin deposits: an outlook from

the continental origin. Earth-science Reviewer 49, 201-221.

Trivedi, P., Axe, L., Dyer, J., 2001. Adsorption of metal ions onto goethite: single-adsorbate and

competitive systems. Colloids and Surfaces A: Physicochemical and Engineering

Aspects, 191 (1-2), 107-121.

Troeh, Frederick R., Louis M., 2005. Soils and Soil Fertility. 6th ed. Ames, Iowa: Blackwell Pub.,

2005.

Tsao, T., Chen, Y., Sheu, H., Tzou Y., Chou, Y., Wang, M., 2013. Separation and identification

of soil nanoparticles by conventional and synchrotron X-ray diffraction. Applied Clay

Science 85, 1-7.

Tyler, L.D., McBride, M.B., 1982. Mobility and Extractability of Cadmium, Copper, Nickel, and

Zinc in Organic and Mineral Soil Columns. Soil science. 134( 3).

Tzou, Y., Wang, S., Hsu, L., Chang, R., Lin, C., 2007. Deintercalation of Li/Al LDH and its

application to recover adsorbed chromate from used adsorbent. Applied Clay Science,

37 (1-2), 107-114.

124 | REFERENCES

Udom, B.E., Mbagwu, J.S.C., Adesodun, J.K., Agbim, N.N., 2004. Distributions of zinc, copper,

cadmium and lead in tropical ultisol after long-term disposal of sewage sludge. Environ.

Int. 30, 467-470.

Unisa, S., Jagannath, P., Dhir, V., 2011. Populationbased study to estimate prevalence and

determine risk factors of gallbladder diseases in the rural Gangetic basin of North India.

HPB, 13, 117-25.

Usman, A., 2008. The relative adsorption selectivies of Pb, Cu, Zn, Cd, and Ni by soils developed

on shale in New Valley, Egypt. Geoderma, 144, 334-343.

Ussiri, N., Johnson, E., 2004. Sorption of organic carbon fractions by Spodosol mineral horizons.

Soil Science Society of America Journal, 68, 253-262.

Vaghetti, P., Lima, C., Royer, B., Brasil, L., Da Cunha, M., Simon, M., et al., 2008. Application

of Brazilian pine fruit coat as a biosorbent to removal of Cr(IV) from aqueous solution-

kinetic and equilibrium study. Biochem Eng J, 42, pp. 67-76.

Velde, B., 1986. Compositional variation in component layers in natural illite/smectite. Clays and

Clay Minerals, 34, (6), 651-657.

Velde, B., Meunier, M., 2008. The Origin of Clay Minerals in Soils and Weathered Rocks.

Springer-Verlag, Berlin, Heidelberg, 406 pp.

Voleski, B., Weber, J., Park, M., 2003. Continuous-flow metal biosorption in a regenerable

Sargassum column. Water Res., 37, 297-306.

Volzone, C., Ortiga, J., 2004. Influence of exchangeable cations of montmorillonite on gas

adsorptions. Journal IchemE B. Process Safety and Environmental Protection; Lugar:

USA; Año, 82, 172-176.

Volzone, C., 2007. Retention of pollutant gases: Comparison between clay minerals and their

modified products. Applied Clay Science, 36, 191-196.

Wang, K.J., Xing, B.S., 2005. Structural and sorption characteristics of adsorbed humic acid on

clay minerals. J Environ Qual., 34(1), 342-9.

Wang, M., Wang, Y., Tan, W., Liu, F., Feng, X., Koopal, K., 2010. Effect of 1-1 electrolyte

concentration on the adsorption/ desorption of copper ion on synthetic birnessite. J Soils

ediments, 10, 879-885.

Wang, W., Brusseau, M.L, Artiola, J.F., 1997. The use of calcium to facilitate desorption and

removal of cadmium and nickel in subsurface soils. Journal of Contaminant Hydrology,

25, 325-336.

Wattel-Koekkoek, W., Buurman, P., van der Plicht, J., Wattel, E., van Breemen, N., 2003. Mean

residence time of soil organic matter associated with kaolinite and smectite. European

Journal of Soil Science, 54, 269-278.

Weaver, C.E., 1956. The distribution and identification of mixed-layer clays in sedimentary

rocks: Am. Mineral., 41, 202-221.

Weishaar, J.L., Aiken, G.R., Bergamaschi, B.A., Fram, M.S., Fujii, R., Mopper, K., 2003.

Evaluation of Specific Ultraviolet Absorbance as an Indicator of the Chemical

Composition and Reactivity of Dissolved Organic Carbon. Environ. Sci. Technol., 37,

4702-4708.

REFERENCES | 125

Weng, L., Van Riemsdijk, W.H., Hiemstra, T., 2009. Effects of fulvic and humic acids on

arsenate adsorption to goethite: Experiments and modelling. Environmental Science and

Technology, 43,7198-7204.

Weber, J.H., 1988. Binding and transport of metals by humic materials. IN: Humic Substances

and Their Role in the Environment. Wiley, New York, 165-178.

Whittig, L.D., 1965. X-ray diffraction techniques for mineral identification and mineralogical

composition. Agronomy No. 9 Part 1, 671-698.

Wösten, M., Lilly, A., Nemes, A., Le Bas, C., 1999. Development and use of a database of

hydraulic properties of European soils. Geoderma 90, 169-185.

Xu, W., Lan, H., Wang, H., Liu, H., Qu, J., 2015. Comparing the adsorption behaviors of Cd, Cu

and Pb from water onto Fe-Mn binary oxide, MnO2 and FeOOH. Frontiers of

Environmental Science & Engineering , 9 (3), 385-393.

Xu, Y., Axe, L., Yee, N., Dyer, J., 2006. Bidentate complexation modeling of heavy metal

adsorption and competition on goethite. Environ. Sci. Technol., 40 (7), 2213-2218.

Xu, Z., Cai, J., Pan, B., 2013. Mathematically modeling fixed-bed adsorption in aqueous systems.

Journal of Zhejiang, Science A (Appl Phys & Eng) 14(3), 155-176.

Yan, G., Viraraghavan, T., Chen, M., 2001. A new model for heavy metal removal in a

biosorption column. Adsorption Science & Technology 19 (1).

Yi, H., Deng, H., Tang, L., Yu, F., Zhou, X., Liu, Y., 2012. Adsorption equilibrium and kinetics

for SO2, NO, CO2 on zeolites FAU and LTA. Journal of Hazardous Materials, 203-204,

111-117.

Youssef, A.M., 2008. Mapping the Pliocene Clay Deposits Using Remote Sensing and its Impact

on the Urbanization Developments in Egypt: Case Study, East Sohag Area. Geotech

Geol Eng., 26:579-591.

Yukselen-Aksoy, Y., Kaya, A., 2010. Predicting soil swelling behaviour from specific surface

area. Geotech. Eng. 163, 229-238.

Zaki, R., Ismail, E. A., Mohamed, W. S., Ali, A., 2015. Impact of Surface Water and

Groundwater Pollutions on Irrigated Soil, El Minia Province, Northern Upper Egypt.

Journal of Water Resource and Protection, 7, 1467-1472. Zeid, I.M., Ghazi, S.M.,

Nabawy, D.M., 2013. Alleviation of heavy metals toxicity in waste water used for plant irrigation.

International journal of Agronomy and Plant Production, 4 (5), 976-983.

Zhang, G.S., Liu, H.J., Liu, R.P., Qu, J.H., 2009. Removal of phosphate from water by a Fe-Mn

binary oxide adsorbent. Journal of Colloid and Interface Science, 335(2), 168-174.

Zhang, J., Hua, P., Krebs, P., 2016. The influences of dissolved organic matter and surfactant on

the desorption of Cu and Zn from road-deposited sediment. Chemosphere 150, 63-70.

Zhang, M., Zhang, H., 2010. Co-transport of dissolved organic matter and heavy metals in soils

induced by excessive phosphorus applications. Journal of Environmental Sciences, 22

(4), 598-606.

Zhuang, J., Yu, G., 2002. Effects of surface coatings on electrochemical properties and

contaminant sorption of clay minerals. Chemosphere, 49, 619-628.

126 |

SUMMARY IN ENGLISH | 127

Summary in English

There is an urgent and increasing global need for purification of drinking water

and wastewater, given the human and environmental health concerns caused by

contaminated by heavy metals (HMs) and organic pollutants. In developing

countries, such as Egypt, sophisticated techniques are often not widely

available, using natural low-cost local materials as sorbents is therefore an

important alternative approach. Pliocene clays from Egypt have a unique

physcio-chemical properties and these materials may be alternative scavengers

for toxic HMs, although their potential application in wastewater treatment

technology has not yet assessed because of a lack of information regarding their

adsorption/regeneration characteristics.

The aim of the present study was firstly to assess the potential for using

Pliocene clay deposits from Egypt in inexpensive purification of industrial

wastewater and irrigation water polluted with HMs. Secondly, this thesis also

addresses the remediation of the sorbents contaminated with HMs, as this is a

crucial step in the regeneration of the sorbents for reuse in multiple cycles of

metal adsorption/desorption and/or their clean-up prior to disposal.

The main objectives of the present study are:

1) To identify and characterize the different clay mineral types in the context

of their application in local wastewater treatment

2) To shed light on the paleoclimatic conditions that prevailed during

formation of the sediments and their influence on the sediment’s

composition.

3) To assess the influence of OM-coating on sorbent materials on their

adsorption of Cu, Ni and Zn.

4) To unravel the effect of the timing of the addition of DOM on the

competitive adsorption of Cu, Ni and Zn onto different sorbent

compositions in static and kinetic systems.

128 | SUMMARY IN ENGLISH

5) To quantify the fate and transport of metals in different mineral sorbents as

well as gain insights into leaching behavior under actual environmental

conditions.

Large reserves of Pliocene clay deposits are distributed along River Nile banks

in the Sohag area, Egypt. The suitability of clay materials for potential water

clean-up is normally governed by their physico-chemical properties, such as

cation exchange capacity (CEC), specific surface area (SSA), micropore

volume, and the clay mineral compositions. To this end the Pliocene clay

samples were sampled in the Egyptian Sohag region from four different areas

(Al-Kwamel, Al-Kola, Al-Ahaywa and Wadi Qasab). The Pliocene clay were

characterized by SediGraph analysis, X-ray diffraction analysis through

different treatments, ICP-OES analysis and CO2 gas adsorption (Chapter 2).

The tested sorbent materials was dominated by fine particles (i,e., mainly silt

and clay, 85-98%) and consisted almost exclusively of smectite (59-94%) and

kaolinite (4-38%) minerals. In addition, the mineral assemblages in Pliocene

deposits suggest an origin from chemical weathering conditions under warm

and semi-arid conditions (Chapter 2). Furthermore, the Pliocene clay showed

high values of CEC, SSA and micropore volume. It seems that these physico-

chemical properties of theseclay materials as well as the type and amount of

smectites might be potentially useful in high added-value markets, e.g., as

environmentally friendly and inexpensive raw material for waste water

treatment. However, to further examine such applications, additional

investigations should focus on unraveling the sorption mechanisms between

HMs and sorbent components. In addition, as natural dissolved organic matter

(NDOM) is often present either in the wastewater itself (e.g., industrial and

agricultural effluents) or in the soil (e.g. due to manuring), the presence of

DOM can have a significant influence on the removal of HMs from the

wastewater. Batch adsorption experiments were performed to determine

equilibrium partitioning between the HM ions and the soil adsorption sites

(Chapter 3). The batch sorption experiments were designed to determine the

adsorption of HMs to the materials over a range of HM concentrations. Our

results showed that the strong adsorption of the HMs and the sorption isotherms

were well described by the initial mass (IM) isotherm model. The large

retention of Cu over both Zn and Ni in this research revealed that Cu was

SUMMARY IN ENGLISH | 129

mostly bound through inner-sphere complexes on mineral-phase whereas Ni

and Zn were found to bind predominantly through electrostatic interactions.

This results underline the importance of presence of DOM and its timing of

addition on the affinity of Cu, Zn and Ni for the tested sorbent materials. The

sequential addition of DOM to sorbent material resulted in reduction in the

affinity of tested HMs due to the blocking of binding sites on the surface of

sorbents, in particular when hydroxides are part of sorbents. In contrast,

concurrent addition resulted in enhancement of all metal adsorption affinity.

This chapter concluded that readily available and abundant natural clay

materials has important implications for removing a large percentage of Cu, Zn

and Ni from wastewater.

Although batch experiments are less time consuming and cheaper than

continuous experiments, they do not simulate actual environmental conditions

or allow time-dependent monitoring of contaminants leaching from sorbents

and waste materials. Therefore, column experiments were performed (Chapter

4) which focused on unraveling the mechanisms of such interactions,

specifically in quasi-realistic operational settings. Using the column approach,

information on the kinetics of adsorption of HMs were determined by

quantifying the adsorption capacity for the HMs. Three different scenarios were

employed in this research: columns were loaded with Cu, Zn and Ni only

(control), first loaded with DOM followed by Cu, Zn and Ni, and DOM, Cu, Zn

and Ni simultaneously. The HM mobility was explored in a set of continuous

flow column experiments using a well-defined natural sorbent amended with

goethite, birnessite and/or smectite. The resulting breakthrough curves were

fitted to a modified dose-response model to obtain the adsorption capacity (q0).

Our results revealed moderately enhanced q0 of Cu (8-25%) compared to the

control without DOM, except for the goethite-amended sorbent that exhibited a

10% reduction due to the blocking of binding sites. Meanwhile, for both Zn and

Ni sequential addition of DOM reduced q0 by 1-36% for all tested soils due to

preferential binding of Zn and Ni to mineral phases, in a line with our previous

findings in chapter 3. In contrast, concurrent addition of DOM and HMs

resulted in a strong increase of q0 for Cu, Zn and Ni and all tested sorbents

compared to the control: by 141-299% for Cu, 29-102% for Zn and 32-144%

for Ni. Timing of DOM addition with respect to that of HM therefore has to be

130 | SUMMARY IN ENGLISH

taken into account when assessing the impact of HM pollution on soils through

polluted irrigation- or wastewater in a system where DOM also enters the soil

(e.g. agricultural irrigation in combination with manuring). Similarly, both the

presence of DOM and timing thereof should be taken into account in design of

strategies where soil constituents, e.g. clay minerals, are used to clean-up HM

polluted waste water. The maximum metal adsorption capacity for Cu, Ni and

Zn in the column experiments (Chapter 4) was higher than that in the batch

experiment (Chapter 3) indicating that other metal retention mechanisms such

as precipitation could be involved in addition to adsorption (Chapter 4).

Therefore, both column and batch approach are needed for assess the adsorption

capacities and removal efficiencies of HMs.

The remediation of the sorbents contaminated with HMs is a crucial step in

regeneration of the sorbents and/or their clean-up prior to disposal. Furthermore,

it is important to assess the influence of the presence of natural DOM on the

regeneration process. To this end, clay mineral-rich column material of varying

composition was previously loaded with Cu, Zn and Ni only; first with DOM

followed by Cu, Zn and Ni; or DOM, Cu, Zn and Ni simultaneously (Chapter

4) and these were used as basis for a set of column desorption experiments

(Chapter 5). The columns were leached with 0.001 M CaCl2 in water as a

control eluent and 0.001 M CaCl2 in DOM-containing water as a treatment

eluent. The removal efficiency (E) of the HMs was calculated from the

numerical integration of the regeneration curves. Our results revealed that Ni

and Zn that are predominantly bound through outer sphere complexes showed

substantial removal of up to 81% (Ni) and 89% (Zn) through simple cation

exchange with Ca2+

. As a result, the removal efficiency of Ni and Zn was

reduced when eluted with CaCl2 dissolved in DOM instead because of binding

of exchangeable Ca2+

to the DOM. In contrast, removal of predominantly inner-

sphere complexed Cu2+

was increased by up to 74% when eluted with CaCl2

dissolved in DOM because of inner sphere complexation of Cu with the DOM.

When columns were first loaded with DOM followed by HMs , the highest

removal efficiency was achieved using CaCl2 dissolved in DOM (up to 69% for

Zn, 74% for Ni and 78% for Cu). This indicates a partitioning of HMs bound to

DOM adsorbed on the solid phase to DOM in solution. However, when

columns were loaded by HMs and DOM simultaneously prior to desorption,

SUMMARY IN ENGLISH | 131

removal efficiencies were low for all metals (2-25% for Cu; 11-19% for Zn; and

8-17% for Ni depending on clay mineral composition) regardless of whether

desorption treatment consisted of CaCl2 solution in water or in DOM. This

indicates that column loading by HMs and DOM when added simultaneously

takes place to a large extent through irreversible co-precipitation rather than

adsorption. Upon flushing with the control eluent, hydroxide-rich sorbents (soil-

birnessite and soil-goethite) showed a low release of tested HMs (Chapter 5)

and this due to the previously observed large affinity of all tested metals for the

birnessite-rich sorbent. In contrast, the soil-smectite sorbent showed a large

release for Cu, Zn and Ni which can be explained by the fact that smectite-rich

soil is composed of aluminosilicate minerals, which favored cation exchange of

metal ions during adsorption. These results have important consequences for the

regeneration potential of clay minerals used in wastewater treatment aimed at

removal of HMs, as they indicate that such potential will be significantly

reduced when the water to be treated is rich in DOM. In contrast, for natural soil

systems our results suggest that when HMs enter a soil together with DOM, e.g.

as a result of the use of HM-rich manure in agricultural fields, the mobility of

the HMs will be lower than expected from interaction dynamics of HMs and

clay minerals alone. This confirms that Cu-loaded soils are more susceptible to

remobilization of Cu when DOM rich water infiltrates, whereas Ni and Zn-

loaded soils are more susceptible to remobilization when cation rich water

infiltrates. In circumstances where precipitation plays a role (scenario C)

additional measures would be needed, for instance acidification to redissolve

precipitates.

The synthesis of the findings of this thesis (Chapter 6) discusses the potential

application of local Pliocene clay in wastewater clean-up and impact of the

insights gained on science and society. Our findings provided new data about

low-cost and simple wastewater treatment technologies which can be designed

to provide clean water to meet the needs of a growing economy and to protect

the environment. Application of Pliocene clays in wastewater treatment can

reduce the load of HMs that can pose threat to ground water and soil. As result,

large amount of clean water can be provided for irrigation of the new

reclamation area in Egypt. Further, this research present a new insights about

presence of DOM and its timing of addition in regulating the mobility of HM

132 | SUMMARY IN ENGLISH

as this can has an important impact either in treatment technology or protecting

groundwater reservoirs from metals pollution.

In general, the data presented in our present study forms a foundation for the

potential removal of HM from wastewater using the Pliocene clay material in

question. This research work fills some of the existing knowledge gaps on the

adsorption mechanisms of Cu, Zn and Ni by clay materials in the presence of

DOM and their implications for wastewater treatment technology. The insight

in the obtained present work provides new data about the impact of timing of

addition of DOM on the wastewater treatment. Nevertheless, further work on

the regeneration of exhausted sorbent materials and enhancement of the

adsorption capacity of Pliocene clay has to be done to further develop this

method.

SAMENVATTING IN HET NEDERLANDS | 133

Samenvatting in het Nederlands

Water dat is vervuild met zware metalen (ZMs) en organische verontreinigingen

heeft mogelijk schadelijke gevolgen voor mens en milieu. Er is daarom een

urgente, wereldwijde vraag naar zuiveringsmethodes om dergelijke

contaminanten te verwijderen uit drink- en afvalwater. In ontwikkelingslanden

als Egypte zijn geavanceerde waterzuiveringstechnieken slechts mondjesmaat

voorhanden. Het gebruik van goedkope natuurlijke materialen als

adsorptiemateriaal is daarom een belangrijke alternatieve methode. Pliocene

klei uit Egypte heeft unieke fysisch-chemische eigenschappen waardoor dit

materiaal mogelijk ZMs uit vervuild water kan verwijderen. De mogelijke

geschiktheid van dergelijke klei om ZMs uit vervuild water te halen is echter

nog onvoldoende duidelijk, vanwege een gebrek aan inzicht in de adsorptie- en

regeneratie-eigenschappen ervan.

Het doel van deze studie was om de potentie te onderzoeken van Egyptische

Pliocene kleiafzettingen om ingezet te worden als goedkoop materiaal voor

zuivering van industrieel afvalwater en irrigatiewater. Daarbij adresseert deze

thesis ook de regeneratie van het adsorptiemateriaal nadat het vervuild is

geraakt met ZMs, aangezien dit een cruciale stap is voor het hergebruik van dit

materiaal in meerdere cycli van adsorptie en desorptie voordat het van de hand

wordt gedaan.

De belangrijkste onderzoeksdoelen zijn :

1) Het identificeren en karakteriseren van verschillende kleimineraaltypes en

hun mogelijk gebruik voor lokale waterzuivering

2) Het ontrafelen van de paleo-klimatologische condities gedurende de

formatie van de sedimenten, en hun invloed op de samenstelling van het

sediment

3) Het onderzoeken van de invloed van coating van de klei met organisch

materiaal op de adsorptie van Cu, Ni en Zn

4) Het onderzoeken van het effect van het moment van toevoegen van

opgeloste organische stof op de competitieve adsorptie van Cu, Ni en Zn

134 | SAMENVATTING IN HET NEDERLANDS

aan verschillende adsorptiemateriaalsamenstellingen in statische en

kinetische systemen

5) Het vaststellen van het lot en het transport van metalen in verschillende

minerale adsorptiematerialen gecombineerd met het verkrijgen van inzicht

in het uitspoelingsgedrag onder normale milieucondities

Enorme Pliocene klei-afzettingen zijn te vinden in de Sohag regio langs de Nijl

in Egypte. De geschiktheid van deze klei voor waterzuivering wordt bepaald

door fysisch-chemische eigenschappen, zoals de kation-uitwisselings-capaciteit

(KUC), het specifieke oppervlak (SO), het volume van de microporiën en de

samenstelling van de kleimineralen. Om deze eigenschappen te onderzoeken is

de Pliocene klei in vier verschillende gebieden in de Egyptische Sohag regio

opgegraven (Al-Kwamel, Al-Kola, Al-Ahaywa en Wadi Qasab). De Pliocene

klei is gekarakteriseerd door middel van SediGraph analyse, X-ray diffraction

analyse met verschillende behandelingen, ICP-OES analyse en CO2 gas

adsorptie (Hoofdstuk 2). De geteste adsorptiematerialen worden gedomineerd

door fijn materiaal (voornamelijk silt en klei, 85-98%) en bestaan voornamelijk

uit smectiet (59-94%) en kaoliniet (4-38%). De minerale samenstelling

suggereert dat de materialen zijn ontstaan als gevolg van chemische verwering

onder warme en semi-aride condities (Hoofdstuk 2). Verder kent de Pliocene

klei hoge KUC-, SO- en microporiënvolume waardes. Deze fysisch-chemische

eigenschappen alsmede het gehalte aan smectiet zorgen ervoor dat de klei

potentieel waardevol is als milieuvriendelijk en goedkoop materiaal voor

waterzuivering. Om hier zeker van te zijn is verder onderzoek naar de

sorptiemechanismen tussen ZMs en het adsorptiemateriaal nodig. Verder kan

natuurlijke opgeloste organische stof (DOM), wanneer het aanwezig is in

afvalwater (industrieel of uit de landbouw) of in de bodem (bijvoorbeeld

vanwege bemesting), een sterke invloed hebben op de verwijdering van ZMs uit

afvalwater. Kortlopende adsorptie-experimenten zijn uitgevoerd om het

verdelingsevenwicht tussen ZM-ionen en bodemadsorptieplekken te

onderzoeken (Hoofdstuk 3). De losstaande adsorptie-experimenten zijn

dusdanig opgezet dat de adsorptie van ZMs aan het adsorptiemateriaal over een

reeks van ZM concentraties onderzocht kon worden. Het onderzoek liet zien dat

de sterke adsorptie van ZMs en de sorptie-isothermen goed beschreven kunnen

worden door het ‘initiële massa’ isothermmodel. De sterkere retentie van Cu in

SAMENVATTING IN HET NEDERLANDS | 135

vergelijking met Zn en Ni duidde erop dat Cu adsorbeert door middel van

complexatie met de minerale fase, terwijl Ni en Zn adsorberen door middel van

elektrostatische interacties. Dit resultaat belicht de invloed van de aanwezigheid

van DOM en het juiste moment van toevoegen hiervan aan het

adsorptiemateriaal, op de affiniteit van Cu, Zn en Ni voor de geteste

adsorptiematerialen. Het stap voor stap toevoegen van opgeloste organische stof

aan het adsorptiemateriaal resulteert in een verlaagde affiniteit voor ZMs als

gevolg van het blokkeren van adsorptielocaties op het oppervlak van het

materiaal. Dit effect was nog sterker wanneer hydroxides deel uitmaken van het

adsorptiemateriaal. In tegenstelling tot deze observatie leidt gelijktijdige

toevoeging (ZMs+DOM) tot een sterkere adsorptie van alle ZMs. Op grond van

de resultaten beschreven in dit hoofdstuk wordt geconcludeerd dat gemakkelijk

verkrijgbare klei-materialen een belangrijke rol kunnen vervullen in de

verwijdering van Cu, Zn en Ni uit afvalwater.

Batchexperimenten nemen minder tijd in beslag en zijn goedkoper dan

kolomexperimenten. Daarentegen simuleren batchexperimenten de werkelijke

milieucondities niet goed en is het niet mogelijk om de tijdsafhankelijke

uitspoeling van de te adsorberen stoffen en het afvalmateriaal te monitoren. Om

dit te ondervangen zijn kolomexperimenten uitgevoerd waarbij de interactie

tussen de stoffen is onderzocht onder semi-realistische omstandigheden

(Hoofdstuk 4). Door middel van het kolomexperiment is er informatie

verkregen over de kinetiek van de adsorptie van ZMs door het kwantificeren

van de adsorptiecapaciteit voor ZMs. In dit onderzoek zijn drie verschillende

scenario’s gebruikt: kolommen voorzien van Cu, Zn en Ni (controle), initiële

toevoeging van opgeloste organische stof gevolgd door Cu, Zn en Ni, en

simultane toevoeging van opgeloste organische stof, Cu, Zn en Ni. De

mobiliteit van de ZMs is onderzocht in een continue doorstroom-

kolomexperiment gebruikmakend van een goed beschreven

standaardadsorptiemateriaal waaraan goethiet, birnessiet en/of smectiet was

toegevoegd. De verkregen doorbraakcurves zijn vergeleken met een modified

dose-response model voor het berekenen van de adsorptie capaciteit (q0). Het

experiment liet een licht verhoogde q0 zien voor Cu (8-25%) vergeleken met de

controlekolom, uitgezonderd voor het met goethiet verrijkte adsorptiemateriaal

dat een 10% reductie liet zien als gevolg van het blokkeren van de

136 | SAMENVATTING IN HET NEDERLANDS

adsorptieplaatsen. Voor de ZMs Ni en Zn resulteerde de sequentiële toevoeging

van eerst opgeloste organische stof en vervolgens ZMs in een verlaging van de

q0 van 1-36% voor alle geteste bodems als gevolg van preferente binding van

Zn en Ni aan de minerale fase. Dit komt overeen met de resultaten in hoofdstuk

3. Daarentegen resulteerde de gelijktijdige toevoeging van opgeloste organische

stof en ZMs in een sterke verhoging van de q0 voor Cu, Zn en Ni in alle geteste

adsorptiematerialen in vergelijking met de controlegroep: 141-299% voor Cu,

29-102% voor Zn en 32-144% voor Ni. Het moment van het toevoegen van

opgeloste organische stof ten opzichte van de toevoeging van ZMs moet daarom

in acht worden genomen wanneer de impact van met ZMs vervuild irrigatie- of

afvalwater wordt onderzocht in een bodemsysteem waar ook opgeloste

organische stof aanwezig is of wordt toegediend (bijvoorbeeld irrigatie met

water vervuild met ZMs in combinatie met bemesting). Evenzo moet rekening

worden gehouden met de aanwezigheid van opgeloste organische stof bij

strategieën waarbij bodemdeeltjes, bijvoorbeeld kleimineralen, worden gebruikt

voor de verwijdering van ZMs uit vervuild afvalwater. De maximale adsorptie-

capaciteit voor Cu, Ni en Zn in de kolomexperimenten (Hoofdstuk 4) was

hoger dan in de batchexperimenten (Hoofdstuk 3). Dit geeft aan dat naast

adsorptie mogelijk ook andere mechanismen zoals neerslaan betrokken kunnen

zijn (Hoofdstuk 4). Deze resultaten laten zien dat zowel een kortlopende batch-

als continue kolomaanpak nodig is voor het onderzoeken van

adsorptiecapaciteit en verwijderingsefficiënties van ZMs.

Het regenereren van adsorptiematerialen vervuild met ZMs is een cruciale stap

bij het eventuele hergebruik van deze adsorptiematerialen of bij het zuiveren

van deze materialen voordat ze worden afgedankt. Verder is het belangrijk om

de rol van natuurlijk aanwezig DOM op het saneringsproces te bestuderen. Om

de rol van natuurlijk aanwezig DOM te bestuderen zijn kolommen met kleirijk

materiaal op drie manieren behandeld: verzadigd met alleen Cu, Zn of Ni;

verzadigd met DOM en daarna Cu, Zn of Ni; en tegelijkertijd verzadigd met

DOM, Cu, Zn en Ni (zie Hoofdstuk 4); deze kolommen werden vervolgens

gebruikt voor een set kolomdesorptieëxperimenten (Hoofdstuk 5). De

kolommen zijn geëlueerd met een 0.001 M CaCl2 oplossing als controle en met

een 0.001 M CaCl2 + DOM bevattende oplossing als behandeling. De

verwijderingsefficiëntie (E) van de ZMs is berekend uit de numerieke integratie

SAMENVATTING IN HET NEDERLANDS | 137

van de regressiediagrammen. De resultaten geven aan dat aanzienlijke

percentages Ni (81%) en Zn (89%), hoofdzakelijk gebonden in de vorm van

outer sphere complexen, verwijderd werden door kationuitwisseling met Ca2+

.

De uitwisseling met Ca2+

vermindert de verwijderingsefficiëntie van Ni en Zn

wanneer een CaCl2 oplossing met daarin DOM wordt toegepast door de binding

van uitwisselbaar Ca2+

met het DOM. Daarentegen nam de verwijdering van

Cu2+

, dat voornamelijk inner sphere bindingen aangaat, toe met 74% wanneer

de kolommen werden gespoeld met een DOM bevattende CaCl2 oplossing door

de inner sphere complexatie van Cu met DOM. Wanneer de kolommen eerst

verzadigd worden met DOM, gevolgd door belading met ZMs, is de hoogste

verwijderingsefficiëntie te zien bij het gebruik van een DOM bevattende CaCl2

oplossing (tot 69% voor Zn, 74% voor Ni en 78% voor Cu). Dit impliceert een

partitie van ZMs tussen het in de vaste fase (kolommateriaal) aanwezige DOM

en het DOM in de oplossing. Echter, wanneer de kolommen voorafgaand aan

het desorptie-experiment tegelijkertijd beladen werden met HMs en DOM, dan

zijn de verwijderingsefficiënties laag voor alle metalen (2-25% voor Cu, 11-

19% voor Zn en 8-17% voor Ni) onafhankelijk van behandeling met een CaCl2-

oplossing alleen of een CaCl2-oplossing met DOM. Dit betekent dat wanneer

ZMs en DOM gelijktijdig worden toegevoegd aan de kolom er kennelijk een

onomkeerbare co-precipitatie plaatsvindt in plaats van reversibele adsorptie. Bij

doorspoeling met de controle-oplossing toonden hydroxide-rijke

adsorptiematerialen (bodem+birnessiet en bodem+goethiet) een lage desorptie

van de geteste ZMs (Hoofdstuk 5) vanwege de eerder geobserveerde hoge

affiniteit van alle geteste metalen voor birnessiet-rijke adsorptiematerialen.

Bodemmateriaal dat smectiet bevat toont daarentegen een hoge desorptie van

Cu, Zn en Ni, hetgeen kan worden verklaard doordat smectiet-rijke bodems

hoofdzakelijk aluminiumsilicaten bevatten; aluminiumsilicaten prefereren

kationuitwisseling van metaalionen tijdens adsorptie. Deze resultaten hebben

belangrijke implicaties voor het regeneratiepotentieel van kleimineralen die

gebruikt worden in waterzuivering gericht op de verwijdering van ZMs. De

resultaten betekenen dat wanneer het water dat gezuiverd moet worden een

hoog gehalte aan DOM bevat, de geschiktheid van klei als zuiveringsmateriaal

sterk af neemt. Echter, voor natuurlijke systemen impliceren deze resultaten dat

wanneer ZMs een bodem infiltreren tegelijkertijd met DOM, bijvoorbeeld als

resultaat van het gebruik van ZM-rijke meststoffen in de landbouw, de

138 | SAMENVATTING IN HET NEDERLANDS

mobiliteit van HMs lager zal zijn dan verwacht wordt als er alleen rekening

wordt gehouden met de interacties tussen HMs en kleimineralen. Uit de

resultaten volgt dat Cu-rijke bodems vatbaarder zijn voor de hermobilisering

van Cu wanneer DOM-rijk water infiltreert, terwijl Ni- en Zn-rijke bodems

vatbaarder zijn voor hermobilisering wanneer kationrijk water infiltreert.

Omstandigheden waarin precipitatie een rol speelt (Scenario C) zouden verder

bestudeerd moeten worden, bijvoorbeeld als verzuring leidt tot oplossing van de

neergeslagen complexen.

De synthese van de bevindingen van deze thesis (Hoofdstuk 6) behandelt de

potentiële toepassing van lokaal gewonnen Pliocene kleien in waterzuivering en

de impact van de verkregen inzichten voor de wetenschap en de maatschappij.

De bevindingen verschaffen nieuwe gegevens over goedkope en simpele

waterzuiveringstechnieken die kunnen worden toegepast om schoon water te

produceren en die het milieu kunnen beschermen in ontwikkelingslanden.

Toepassing van Pliocene kleien in waterzuivering kan het ZM-gehalte

verminderen van afvalwater dat anders een mogelijke bedreiging vormt voor

grondwater en bodem wanneer het water voor bevloeiingsdoeleinden wordt

gebruikt. Op deze manier kunnen reclamatiegebieden in Egypte met grote

hoeveelheden schoon water worden geïrrigeerd en zo geschikt gemaakt worden

voor bijvoorbeeld landbouw. Verder biedt het onderzoek een nieuw inzicht in

de invloed van de aanwezigheid van DOM en vooral ook in het moment van

toevoegen van DOM, zowel aan het sorptiemateriaal als in de

regeneratieoplossing, bij de regulatie van de mobiliteit van ZMs, dat op zijn

beurt een grote invloed kan hebben op zuiveringsinstallaties of het beschermen

van grondwaterreservoirs tegen vervuiling met HMs.

De gegevens verkregen in deze thesis vormen een basis voor een technologie

om ZMs uit vervuild water te verwijderen met behulp van het gebruik van

Pliocene kleien. Het onderzoek vult een aantal van de bestaande kennishiaten op

met betrekking tot de adsorptiemechanismen van Cu, Zn en Ni door klei in de

aanwezigheid van DOM en de implicaties hiervan voor waterzuivering.

Desondanks is er verder onderzoek vereist naar de regeneratie van gebruikte

adsorptiematerialen en naar manieren om de adsorptiecapaciteiten van Pliocene

kleien verder te verhogen teneinde deze technologie verder te ontwikkelen.

139 | الملخص باللغة العربية

الملخص باللغة العربية

املعادن الثقيةل واملواد ىف الاونة احلديثة ظهرت احلاجة املاسه لتنقية مياه الرشب واملياه امللوثه نتيجه لالهامتم ابلصحة العامة وجتنب اخملاطر الناجتة من

ن مواد طبيعيه قليةل التلكفة وصديقة للبيئه للتخلص من هذة امللواثت ىف املياه خاصة ىف املياه املس تخدمة. ذلكل اكن من املهم البحث ع املتواجدة العضوية

ىف ادلول الناميه مثل مرص حيث ان بعض معليات التنقية احلديثة واملعقدة غري متوفرة ىف كثري من الاماكن.

ئية وكيميائية فريدة حيث ان هذة اصخصائص جتعل من هذة املواد يوجد ىف مرص خامات طبيعية مثل رواسب البليوسني الطينية والىت تمتزي خبصائص فزياي

قص املواد السامة من املياه امللوثة عىل الرمغ من ان اماكنية اس تخدام هذة املواد ىف تكنولوجيا تنقية املياه مل تمت حىت الان نتيجة لن الزاةلاصخام مادة بديةل

ملياه امللوثة وايضا اعادة اس تخدام هذة املواد مرة اخرى بعد التش بع ابملواد السامة.املعلومات حول معلية امتصاص املواد السامة من ا

:الهدف الرئيىس لدلراسة احلالية هوذلكل اكن

ناجتة من رصف تقيمي اماكنية اس تخدام رواسب البليوسني ىف مرص مكواد قليةل التلكفة لتنقية املواد امللوثة وتقليل نس بة املعادن الثقيةل السامة ال اوال:

املصانع واحلقول الزراعية والىت غالبا ما تكون محمةل ابملعادن الثقيةل السامه.

املياه امللوثة حيث تعترب اثنيا: الهدف الثاىن لهذة ادلراسة هو كيفية معاجلة رواسب البليوسني املس تخدة ىف التنقية بعد تش بعها ابملعادن الثقيهل اثناء تنقية

قبل ان يمت التخلص من املواد املتس تخدمة او اعادة اس تخداهما بشلك متكرر ىف معليات التنقيه. هذة خطوه هامة جدا

الغرض من هذة ادلراسة يتكون من:

عىل خصائص املعادن الطينية اخملتلفة املوجودة ىف املواد املس تخدمة ىف س ياق اس تخداهما ىف معلية تنقية املياه امللوثة التعرف (1

لظروف املناخية القدميه الىت اكنت سائدة اثناء تكون رواسب البليوسني واتثري ذكل عىل مكوانت الرواسب.القاء الضوء عىل ا (2

العضويه عىل جحب الاسطح اصخارجية ملكوانت املواد املس تخدمه عىل معلية امتصاص امللواثت من املياه امللوثه. املواد تقيمي مدى اتثري (3

لعضويه عىل امتصاص عنارص النحاس والزنك والنيلك من املياه امللوثة ىف حاهل الاوساط املتوازنه او كشف مدى اتثري توقيت اضافة املواد ا (4

احلركيه وايضا املنافسة فامي بني هذة العنارص عىل مناطق الامتصاص عىل اسطح املكوانت اخملتلفة ىف رواسب البليوسني.

المكية املتبقيه من املياه امللوثة ىف املكوانت اخملتلفة ىف رواسب البليوسني وايضا اعطاء التقدير المكى لمكية امللواثت الىت مت اس تخالصها وايضا (5

البيئه احمليطه. رؤيه واحضة عن كيفية ترسب هذة املواد السامة من الرواسب حتت الظروف الطبيعية ىف

تقيمي رواسب البليوسن ان وزع عىل طول ضفىت هنر النيل.مرص تت احتياطات كبرية جدا من خام رواسب البليوسني الطينية ىف منطقة سوهاج توجد

يائية مثل سعه تبادل الطينيه مكواد خام واماكنية اس تخداهما ىف تكنوجليا معاجلة املياه امللوثه يتحمك فهيا بشلك رئيىس خواص هذة املواد الفزيايئيه والكيم

، املسام ىف جحم امليكرو، ومكوانهتا من املعادن الطينيه. ذلكل مت مجع عينات من مساحة السطح ملكوانت املادة اصخام العنارص داخل هذة الرواسب،

وادى قصب. ولتوصيف ودراسه هذة الرواسب الكوال، الاحايوه، رواسب البليوسني ىف حمافظة سوهاج من اربع مناطق رئيس يه ىه مناطق الكوامل،

اهجزة س يدجيراف وحيود الاشعه الس ينيه من خالل عده معاالجات قبل معلية القياس ابس تخدام الرواسب ابلتفصيل متت التحاليل والقياسات لهذة

العينات الىت مت مجعها واختبارها اكن .لفصل الثاىن(( وامتصاص غاز اثىن اكس يد الكربون )اICP-OESوايضا مت اس تخدام حتليل هجاز البالزما الطيفى )

الملخص باللغة العربية | 140

%( 94-59%( وحتتوى بشلك رئيىس عىل معادن السمكتيت )98-85با ما تكون ىف جحم الغرين والطني يغلب عىل مكوانهتا احلجم ادلقيق حلبيباهتا )غال

ان هذة الرواسب تكونت نتيجة لعوامل جتويه كيميائيه ىف %(. ابالضافة ذلكل فان هذة الصحبة من املعادن الطينيه تشري اىل38-4ومعادن الكولينيت )

ء والش به جاف )الفصل الثاىن(. والاك ر من ذكل ان رواسب البليوسني الطينيه اظهرت سعه كبريه لتبادل ظروف مناخية يغلب علهيا الطقس ادلاىف

كيميائيه لرواسب البليوسني وايضا نوع -مساحة كبرية لسطح حبيباهتا ومكية املسام ادلقيقة هبا. ومن الواحض من خالل هذة اصخصائص الفزييو العنارص وايضا

الس تخدام هذة املواد اصخام مكواد طبيعية صديقة للبيئة ورخيصة المثن ملعاجله املياه امللوثة ابملعادن الثقيةل هناك اماكنيه كبريه هبا ان ومكيه معادن السمكتيت

فهناك دراسات فان االتكنولوجي من النوع هذا واذلى بدوره يكس هبا قيمة اقتصاديه كبرية ىف املس تقبل. ورمغ ذكل فانة الختبار هذة املواد بشلك تفصيىل ىف

ياه امللوثة ومكوانت اك ر تفصيال جيب الرتكزي علهيا مثل مياكنيكية وطبيعة معلية امتصاص امللواثت والتفاعالت الىت تمت بني العنارص املراد ازالهتا من امل

رواسب البليوسني.

لوثة نفسها الناجتة من املصانع او الرصف الزراعى او تكون موجودة ىف الرتبة اضافة الا ذكل فان وجود املواد العضويه الىت ميكن ان توجد سواء ىف املياه امل

ثقيةل السامة من عىل الزراعية نتيجة التسميد الطبيعى فان هذة املواد العضوية من املمكن ان ينتج عهنا اتثري كبري وملحوظ عىل امتصاص او ازاةل املعادن ال

دلراسة batch adsorptionوملعرفة لك هذة التفاصيل مت اجراء جتارب سوف يؤثر ابلطبع عىل البيئه احمليطه. اسطح املواد املس تخدة الزالهتا وايضا هذا

هذا النوع من التجارب يس تخدم بصفة عامة الجياد العالقة املتبادل بني املعادن .ظروف التوازن بني املواد املس تخدة واملعادن السامة )الفصل الثالث(

هذة لسطح الىت متتص علية )خام البلوسني( وذكل بتغيري مكية العنارص السامة ىف احملاليل الىت يمت حتضريها والىت حتاىك املياه امللوثة ىفالثقيةل السامة وا

من هجة وبني لكاحلاةل. ىف هذة ادلراسة تبني امهيه وتأ ثري وجود املواد العضويه وتوقيت اضافهتا عىل مدى قوة الرتابط بني عنارص النحاس والزنك والني

. من خالل Initial Mass (IM) isothermوتفسريه ابس تخدام منوزج املواد اصخام املس تخدمة من انحية اخرى. امتصاص الايزوسريم قد مت وصفة

البلوسني ترجع اىل ان النتاجئ قد تبني ان مكيه ايوانت النحاس الىت مت امتصاصها والىت اكنت تفووق بكثري مكيه ايوانت الزنك والنيلك ابس تخدام خام

( بيامن يرتبط الك من الزنك والنيلك بشلك inner-sphere complexesالنحاس يرتبط ابسطح املعادن املكونه صخام البليوسني بروابط تسامهية قويه )

دامئ بروابط ايونية ضعيفة )كهروس تاتيكيه ضعيفة(.

صاص العنارص السامة فانة وجد انه عندما يمت اضافة املواد العضويه اوال اىل خام البليوسني قبل ان فامي يتعلق بتاثري توقيت اضافة املواد العضوية عىل امتو

اص املعادن السامة هبا يمت اضافة املياه امللوثة ابملعادن الثقيةل فان هذا يؤدى اىل ان املواد العضوية املضافة تعمل عىل جحب )جحز( الاماكن الىت يمت امتص

املكونة صخام البليوسني خاصة ىف وجود معادن الهيدروكس يد ىف املواد املس تخدمة مثل اكس يد احلديد واملنجنزي. ونتيجة ذلكل فان مكية عىل سطح املعادن

نارص وية والع قليةل جدا من النحاس والزنك والنيلك سوف يمت امتصاصها واذالهتا من املياه امللوثة. وعىل العكس من ذكل فانة عندما تضاف املواد العض

لبليوسني. وبناءا السامة ىف نفس الوقت اىل خام البليوسني فانة وجد ان مكية كبريه جدا من النحاس والزنك والنيلك متتص عىل سطح املعادن املكونه صخام ا

تنقية املياه امللوثة و امتصاص عىل هذا قد وحضت النتاجئ ان اصخصائص الفريدة واملمزيه صخام البليوسني من املمكن ان جتعل منة خام ميكن اس تخدامة ىف

مكية كبريه من املعادن السامة.

من التجارب الىت تس تغرق زمنا ليس كبريا وارخص تلكفة مقارنة بتجارب اخرى ولكهنا غري مناس بة حملااكة ما batch adsorptionقد تكون جتارب

ت الصلبة حيث انه ىف هذة احلاةل دراسة احلاةل املياكنيكية للملواثت هامة جدا وهذا حيدث ىف الطبيعة اثناء ترسب املعادن السامة من املواد او من اخمللفا

)الفصل الرابع( والىت من خاللها ميكن دراسة مياكنيكية column experimentsذلكل مقنا ابجراء جتارب .batch adsorptionال يتوفر ىف جتارب

عادن املكونة صخام البليوسني ىف بيئة شييه ملا حيدث ىف الطبيعة. ابس تخدام هذة التجارب ميكننا التفاعالت الىت حتدث بني املعادن الثقيهل وسطح امل

ن لهذة املواد ان اس تخالص معلومات غاية ىف الامهية مثل احلركة ادليناميكية للملواثت ومكيهتا عىل اسطح املواد املس تخدمة لتنقيهتا وايضا اقىص سعة ميك

ت من املياه. لقد مقنا ىف هذ الفصل )الفصل الرابع( بتصممي ثالثة سيناريوهات خمتلفة وىه لكتاىل:متتلكها الزاةل امللواث

اكن بضخ حملول من املعادن الثقيهل ) النحاس والزنك والنيلك( خالل الامعدة الىت حتتوى عىل املواد املس تخدمة لغرض التنقية حيث : الس ناريو الاول

ملرجع للرجوع الية ملقارنة نتاجئة ابلنتاجئ ىف الس ناريوهات الباقية.يعترب هذا الس ناريو مبثابة ا

141 | الملخص باللغة العربية

غرض التنقية ملدة ليةل السيناريو الثاىن: ىف هذا الس ناريو مقنا اوال بضخ حملول من املواد العضوية اذلائبة اىل الامعدة الىت حتتوى عىل املواد املس تخدمة ل

ول اذلى حيتوى عىل املعادن الثقيهل.اكمةل وىف صباح اليوم الثاىن مت بدأ خض احملل

لىت حتتوى عىل السيناريو الثالث: ىف هذا السيناريو قد مت خض خليط من املواد العضوية اذلائبة وحملول املعادن الثقيةل ىف نفس الوقت خالل الامعدة ا

املواد املس تخدمة لغرض التنقية.

دن الثقيةل اكن بشلك مس متر ابس تخدام مضخة كهرابئية واكنت الامعدة اململؤة ابملواد املس تخدمة معلية خض احملاليل سواء اكنت العضوية او حملول املعا

من لك هذة لغرض التنقية حتتوى عىل نوع من الرتبة املعدةل ابضافة معدن السمكتيت، جيوثيت )اكس يد احلديد(، برينزييت )اكس يد املنجنزي(، او خليط

حيتوى فقط عىل تربة فقط بدون اى اضافات او تعديل ىف مكوانهتا الس تخداهما مكرجع لالمعدة احملتوية عىل مواد معدةل. اعداد معود املعادن ابالضافة اىل

breakthrough curve)( لسعة املواد املس تخدمة ىف التنقية ىف اس تخالص املعادن الثقيةل مت برمس منحىن الاخرتاق )q0حلساب احلد الاقىص )

(. ولقد اظهرت النتاجئ ان هناك حتسن ىف modified dose response modelوزج معدل يطلق علية الاس تجابة للجرعة )وتفسرية ابس تخدام من

% ىف السيناريو الثاىن )الامعدة مش بعة ابملواد العضويه( 25-8معلية امتصاص النحاس ىف لك الامعدة املش بعة مس بقا ابملواد العضوية مبقدار يرتواح من

ناريو الاول )غياب املواد العضويه( خبالف الالعمدة املعدةل ابضافة معدن اجليوثيت والىت اظهرت اخنفاض ىف سعة الامتصاص تصل اىل مقارنة ابلسي

املواد العضوية % مقارنة ابلسيناريو الاول وذكل نتيجة ان اسطح املعادن ىف الامعدة مت جحب الاماكن الىت يمت علهيا امتصاص العنارص من جانب10

% ىف لك الامعدة بدون 36-1ملضافة مس بقا. وعىل اجلانب الاخر فان سعة امتصاص الك من الزنك والنيلك قد اخنفضت ايضا مبقدار يرتاوح من ا

خ اس تثناء وهذا يرجع اىل ان هذين العنرصيني يمت امتصاصصهم عىل اسطح املعادن حمل ادلراسة عن طريق روابط ايونية سطحية ضعيفة ونتيجة لض

ل املواد العضوية مس بقا لسطح تكل املعادن فان هذا ادى اىل تقلليل فرص ارتباطهم بسطح هذة املعادن. عىل العكس من ذكل فان ىف حاةل حملو

ة ا ىف سعالسيناريو الثالث حيث انة مت خض الك من حملول املواد العضوية وحملول املعادن الثقيةل ىف نفس الوقت فقد ادى ذكل اىل زايدة كبرية جد

141واح من امتصاص املعادن الثقيةل الثالثة ىف لك الامعدة مقارنة ابلسيناريو الاول والثاىن حيث اكنت الزايدة ىف سعة الامتصاص ملعدن النحاس ترت

%.144اىل 32% اما النيلك فاكنت نس بة التحسن ترتاوح من 102-29% و ملعدن الزنك 299اىل

املواد العضوية ىف وجود عنارص ثقيةل سامة ىف املياه جيب ان يأ خذ ىف احلس بان عند تقمي ودراسة وتأ ثري املعادن وعىل ذكل يتضح جليا ان توقيت اضافة

املواد العضوية ممكن ان الثقيةل اخملتلفة وتلوهثا للرتبة عىل سييل املثال واذلى يكون انجت عن رى هذة الرتبة ابملياه امللوثة من الرصف الصناعى او الزراعى.

وثة فان وجود او اجد مصاحبة للمعادن الثقيةل ىف املياه امللوثة او تكون موجودة ىف الرتبة نتيجة التسميد وعىل ذكل فانة عند رى اى تربة ابملياه امللتتو

ما تكون الرتبة غنية ابملواد غياب املواد العضوية سوف يكون ةل ابلغ الاثر ىف تلوث الرتبة او املياه اجلوفية الضحةل القريبة من السطح حيث انة ىف حاةل

فقرية ىف املواد العضوية العضوية فان رى هذة الرتبة ابملياه امللوثة سوف يؤدى اىل ترسب مكية كبريه من املعادن الثقيةل للمياه اجلوفية اما اذا اكنت الرتبة

ية معا فان ذكل سوف يؤدى اىل ترامك العنارص ىف الرتبة ومهنا اىل الغذاء مث الانسان مما يشلك خطورة كربة واملياه ملوثة ابملعادن الثقيةل واملواد العضو

احلس بان وجود او عدم للحياه وللبيئة احمليطة. ايضا عند وضع اسرتاتيجية لتنقية املياه امللوثة بتكل املعادن الثقيةل ابس تخدام خام البليوسني جيب الاخذ ىف

هذة اد العضوية وعىل اى شلك تكون موجودة والىت حامت سوف تؤثر عىل معلية التنقية وكفاءهتا. من النتاجئ الاخرى الهامة الىت ظهرت ايضا ىفوجود املو

الفصل الرابع( عهنا )ادلراسة ان نس بة املعادن الثقيةل الىت مت ازالهتا والتخلص مهنا ابس تخدام اصخام املس تخدم اكنت عالية ىف حال اس تخدام جتارب الامعدة

)الفصل الثالث( وهذا يرجع اىل احامتلية وجود مياكنيكية اخرى خمتلفة حدثت ىف حاهل جتارب الامعدة مثل معلية batchىف حاةل اس تخدام جتاب

ةل دراسة امتصاص املعادن من املهم اختبارمه سواي ىف حا batchالرتسيب جبانب معلية الامتصاص. وعىل ذكل فان الك من جتارب الامعدة وجتارب ال

الثقيةل وتنقية املياه امللوثة.

اد او اعادة من العمليات املهمة ايضا ىه معلية معاجلة املواد اصخام املس تخدمة ىف تنقية املياه حيث هذا يعترب امر رضورى قبل التخلص من هذة املو

ن عدهما وتوقيت اضافهتا للمواد املراد اعادة اس تخداهما ايضا نقطة حيوبة وجيب الرتكزي اس تخداهما مرة اخرى ىف معلية التنقية. تأ ثري وجود املواد العضوية م

اس تخدمت ىف علهيا خاصة عند دراسة اجلدوى الاقتصادية الس تخدام مادة معينة ىف معليات تنقية املياة امللوثة. ودلراسة هذة النقطة فان الامعدة الىت

الملخص باللغة العربية | 142

سعة الامتصاص للمواد حمل ادلراسة سوف يمت معل جتارب علهيا دلراسة كيفية ازاةل ما تعلق بتكل املواد من معادن املرة السابقة )الفصل الرابع( دلراسة

مةل ابملعادن الثقيةل ثقيةل وكيفية معاجلهتا الس تخدام تكل املواد ىف معليات التنقية بشلك متكرر )الفصل اصخامس(. ىف هذا الشأ ن مت معل معاجلة املواد احمل

( وايضا بضخ حملول من لكوريد الاكلس يوم برتكزي control eluentمول/لرت مذاب ىف ماء مقطر ) 0.001لول من لكوريد الاكلس يوم برتكزي بضخ حم

. ىف هذة ادلراسة مت دراسة معاجلة الامعدة بطريقتني (treatment eluent)مول/لرت ولكن هذة املرة مذاب ىف حملول مواد عضوية مذابة 0.001

اثناء بضخ لكوريد اكلس يوم مذاب ىف املاء والطريقة الثانية بضخ لكوريد اكلس يوم مذاب ىف املواد العضوية )لك معود مت حتضري اربعة نسخ منهالاوىل

معودين معلية اختبار سعة الامتصاص ىف الفصل الرابع وبذكل هنا ىف هذة ادلراسة سوف يمت اختبار لك حملول من حماليل لكوريد الاكلس يوم عىل لك

معاجلة وازاةل املعادن الثقيةل من املواد حمل ادلراسة حبساب التاكمل العددى للمنحنيات الناجتة ابس تخدام معادالت (E)للك مادة(. مت حساب كفائة

من الزنك والنيلك واذلى الفصل الرابع( ان الك-معينة )الفصل اصخامس(. قد اظهرت النتاجئ من الامعدة الغري مش بعة مبادة عضوية )السيناريو الاول

نس بة تصل اىل يرتبط الك مهنم بروبط ضعيفة عىل سطح املواد املس تخمة قد متت ازالهتم بنس بة كبرية جدا ابس تخدام لكوريد الاكلس يوم املذاب ىف املاء ب

% للزنك نتيجة للتبادل بيهنم وبني عنارص الاكلس يوم )الفصل اصخامس(.89% للنيلك و 81

% للزنك و 69ام حملولو لكوريد الاكلس يوم املذاب ىف مادة عضوية فان نس بة ازاةل الزنك والنيلك تقلصت بشلك ملحوظ لتصل اىل أ ما عند اس تخد

ضوية املضافة % للنيلك. ويرجع هذا اىل ان بدال من ان حتل عنارص الاكلس يوم حمل الزنك والنيلك الزالهتا من سطح املواد فاهنا تتفاعل مع املادة الع74

% ىف حاةل اس تخدام لكوريد اكلس يوم مذاب ىف املاء اىل 64أ ما فامي خيص النحاس فان المكية املزاةل ارتفعت من تقل فرصة ازاةل الزنك والنيلك.بذكلو

دن مت ازالهتا من % عند اس تخدام لكوريد اكلس يوم مذاب ىف مادة عضوية نتيجة لقوة ارتباط النحاس ابملواد العضوية. ايضا نس بة كبرية من لك املعا74

69الفصل الرابع( حيث اكنت نس بة ازاةل العنارص ىه -اضافة املواد العضوية اوال اىل املواد مث يتلوها املعادن الثقيةلواد ىف حاةل السيناريو الثاىن )اسطح امل

لعضوية املش بعة عىل اسطح املواد حمل ادلراسة % للنحاس وهذا يدل عىل ان املعادن الثقيةل املمتصة عىل اسطح املواد ا78% للنيلك و 74% للزنك،

واد حمل ادلراسة كام هو ايضا ارتبطت ابملواد العضوية املضافة ىف حملول لكوريد الاكلس يوم. غري ان عند اضافة املعادن الثقيةل مع املواد العضويه معا اىل امل

% للنيلك( بغض النظر 17-8% للزنك، 19-11% للنحلس، 25-2نت ضئيةل للغاية )موحض ىف الفصل الرابع فان كفائة ازاةل املعادن الثقيةل من املواد اك

ادن عن نوع حملول لكوريد الاكلس يوم املس تخدم هل هو مذاب ىف املاء او مذاب ىف ماد عضويه. هذا يشري اىل ان عندما اضيف الك من حملول املع

للتنقية )الفصل الرابع( قد حدث نوع من الرتسيب وليس الامتصاص وعىل ذكل عند الثقيةل ىف نفس الوقت مع املاد العضوية اىل املواد املس تخدمة

ىت حتتوى عىل ااكس يد اس تخدام حماليل لكوريد الاكلس يوم ملعاجلة املواد احملمةل ابملعادن الثقيةل مل تمت العملية بشلك فعال. ومما يعزز ذكل ان الامعدة ال

ىف نزع املعادن الثقيةل مهنا ويرجع ذكل اىل اجلاذبيه الشديدة بني تكل الااكس يد و املعادن الثقيةل كام هو احلديد واملنجنزي قد اظهرت مقاومة شديدة

س يوم معروف. وعىل العكس من ذكل متاما فان الامعدة الىت حتتوى عىل معدن السمكتيت اظهرت اس تجابة قوية بعد معاجلاهتا مبحاليل لكوريد الاكل

ن الثقيةل متت ازالهتا من املواد اصخام وهذا ميكن تفسرية ابن معادن السمكتيت والىت تمتى اىل مجموعة معادن سلياكت الالومنيوم حيث مكية كبرية من املعاد

لس يوم بلك تفضل بشلك كبري معلية تبادل الاكتيوانت عىل سطحها اثناء معلية الامتصاص )الفصل الرابع( وبذكل عند اضافة حماليل الاكلس يوم فان الاك

ليوسني بعد هوةل تبادل مع املعادن الثقيةل )الفصل اصخامس(. سوف يكون لنتاجئ هذة ادلراسة أ ثر كبري ىف اماكنية معاجلة واعادة اس تخدام خامات الب س

مللوثة ىف لك مرة. اما اس تخدامه ىف معاجلة تنقية املياه من امللواثت وبذكل ميكن اس تخدامة بشلك دورى ىف معليات التنقية بعد معاجلتة وازاةل املعادن ا

د فان هذا سوف ابلنس بة للرتبة الزراعية فان ادلراسة احلالية اثبتت انه عندما تروى مبياه ملوثة مع وجود مواد عضوية موجودة ىف الرتبة نتيجة التسمي

سوف متتص نس بة عالية مهنا ىف تربة البلوسني. يؤدى اىل حترر نس بة كبرية من العنارص الثقيةل بعكس ان تكون املعادن الثقيةل ىه فقط املوجودة حفينئذ

ابملواد العضوية قادرة عىل ان تزيل و هذة ادلراسة ايضا اثبتت انه عندما تكون الرتبة محمةل بنس بة كبرية من معادن النحاس فان احملاليل الغنية ابلاكتيوانت

يلك فان احملاليل الغنية ابلاكتيوانت قادرة عىل ازاةل نس بة عاليو من هذة املعادن الثقيةل. نس بة عالية من هذة العنارص بيامن اذا اكنت الرتبة محمةل ابلزنك والن

هناك ترسيب وىف الك احلالتني هناك خطر قد يدامه املياه اجلوفية القريبة من سطح هذة الرتبة ابن تتلوث مبثل هذة العنارص. ولكن ىف حاالت ان يكون

العضوية عىل اسطح مكوانت الرتبة فهنا قياسات اخرى جيب اختبارها ىف املس تقبل اكس تخدام مذيب محىض قوى الزاةللالك من املعادن الثقيةل واملواد

هذة املعادن الثقيةل من الرتبة.

143 | الملخص باللغة العربية

-زيوهتا الفزي بصفات فريدة ىف مكوانخالصة ما مت اس تنتاجة ىف هذ ادلراسة )الفصل السادس( يوحض اماكنية تطبيق خامات البليوسن الطينية الىت تمت

رص. هذة ادلراسة اس تاطعت ان كيميائية واملعدنية ىف تنقية املياه امللوثة ابملعادن الثقيهل مما سوف يكون هل الاثر البالغ عىل الناحية العلمية والاجامتعية ىف م

عادن املكونة صخام البليوسني. أ يضا اوحضت ادلراسة جتيب عىل عدة تساؤالت هممة عن تفاعل املعادن الثقيةل حمل ادلراسة مع املواد العضوية وايضا مع امل

اس تخدامة مرة اخرى لعب ان توقيت ووجود املواد العضوية اثناء تنقية املياة او اثناء معاجلة خام البليوسني بعد تش بعة ابمللواثت اثناء معلية التنقية العادة

همم قد ابرزتة االااكس يد املوجودة ىف خام البليوسني مثل اجليوثيت والربنزييت ىف معلية دورا همم ىف سلوك امتصاص او اذاةل العنارص الثقيةل. أ يضا دور

امتصاص او ازاةل امللواثت خبام البليوسن.

املس تقبل اماكنية تطبيق واس تخدام خام البليوسني الزاةل الىت قدمهتا هذة ادلراسة تشلك الاساس اذلى سوف ييىن علية ىف البياانتبشلك عام فان

اسطة خام املعادن الثقيةل السامة من املياه ىف مرص. هذة ادلراسة مل ت الفراغ ىف معرفة مياكنيكية الامتصاص لالك من النحاس، الزنك والنيلك بو

وبينات جديدة مل تدرس البليوسني ىف وجود املواد العضوية وكذكل تأ ثري هذة النتاجئ عىل تنفيذ تكنولوجيا تنقية املياه ىف مرص. ادلراسة ايضا قدمت رؤية

ن ادلراسات من قبل عن تأ ثري توقيت وجود املواد العضوية عىل امتصاص امللواثت وتنقية املياه امللوثة. وىف نفس الس ياق فانة البد من معل املزيد م

ة الامتصاص هل.املس تقبلية خاصة حول اعادة اس تخدام خام البليوسني بعد معلية التنقية وتش بعة ابملعادن الثقيةل وحتسني سع

144 |

ACHKNOWLEGEMENTS | 145

Acknowledgments

I would like to convey my heartfelt gratitude and sincere appreciation to all people who

have supported and inspired me during my doctoral study, and I want to take some time

here to properly acknowledge them. First and foremost, I would like to express my

sincere gratitude and my special thanks to my promotors Prof. Dr. Pim de Voogt, Prof.

Dr. Karsten Kalbitz, and co-promoters Dr. Boris Jansen and Dr. John Parsons for their

endless guidance during my research at University of Amsterdam, you have been a

tremendous mentor for me. Your guidance helped me in all the time of research and

writing of this thesis. I could not have imagined having a better advisor and mentors for

my Ph.D study. I would like to thank you for encouraging my research and for allowing

me to grow as a research scientist. Your advice on both research as well as on my career

has been priceless. It gives me great pleasure to acknowledge the guidance, valuable

suggestions, constructive criticism, and incredible patience of my promoters and my co-

promoters. Your scientific excitement inspired me in the most important moments of

making right decisions and had significantly contributed to this thesis. Thank you for

trusting me.

A few words I want to dedicate to Dr. Boris Jansen in particular. You were giving me

intellectual freedom in my work and supporting my attendance at various conferences.

Thanks for being supportive and understanding during a difficult time, for the countless

hours of revisions and advice on my work and for vote of confidence and helping me

secure funding from the University of Amsterdam for the last year of my PhD research.

You certainly are a great mentor for me. Dear Boris, it was a great pleasure and honor to

work with you and I owe you lots of gratitude.

Additionally, I would like to thank my committee members, Prof. dr. E. Smolders, Prof.

dr. J. Sevink, Prof. dr. dr. P. de Ruiter, Prof. Dr. El-Shater, Prof. dr. ir. W. Bouten, Prof.

dr. El-Haddad and Dr. W. D. Gosling for their interest, time and effort they put into the

evaluation, thorough and critical judgment of this thesis.

I sincerely thank the chair of our group Prof. Dr. Peter de Ruiter, and members Dr. Erik

Cammeraat and Dr. Albert Tietema for their support. Throughout my Ph.D studies the

continuous and generous support from the technical staff of our ESS group is greatly

appreciated: Leo Hoitinga, Leen de Lange, Bert de Leeuw, Jorien Schoorl, Joke

Westerveld, John Visser and Peter Serne. I am earnestly thankful to Chiara Cerli (head

of our labs) and Dr. Katja Heister (Technische Universität München) for their helpful

146 | ACHKNOWLEGEMENTS

suggestions during development of the setup of column experiments. Also, I wish to

thank Yingkai Huang for technical help during the XRD analyses (Van der Waals-

Zeeman Institute, University of Amsterdam) and Norbert Geels for SSA measurements

(Van 't Hoff institute, University of Amsterdam). It was a great pleasure for me to share

the same room with my past and current Ph.D colleagues who contributed in very

diverse ways to my research.

Special thanks goes to the secretaries of our IBED institute which were always ready to

help me: Maria Dolorita, Tanya Noorlander, Mary Parra Tasayco, Pascale Thiery-van

der Bij and Saskia Heijboer, I am equally grateful.

I would also like to extend my sincere thanks to the Egyptian Higher Education

Ministry for their financial support of the first 2 years and to IBED for financial support

for the final year of my PhD project .

To all my friends, from here and there, thank you guys, you are awesome, I really like

you.

Almost but not least, a special thanks to my family. Words cannot express how grateful

I am: my late father, mother, brothers, sisters, mother-in law and father-in-law for all of

the sacrifices that you’ve made on my behalf. Your unflagging love and prayer for me

was what sustained me thus far. My mother was not happy to see me leave to The

Netherlands, but has never complained about it. Her deep faith, her prayers, and

supreme trust are always the most efficient motivation to accomplish my ultimate goal.

I have no suitable word that can fully describe my mother’s everlasting love to me.

Thanks for supporting me spiritually throughout writing this thesis and my life in

general. This thesis is dedicated to the soul of my father, may Allah forgive him and

grant him his highest paradise (Ameen).

And finally, the moment everybody has been waiting for: A big thank you for my wife

Sara Mohamed. This thesis would not have been what it is today without Sara. Being

together with Sara has a positive impact on my life in all aspects. I would like to express

appreciation to my beloved wife Sara who was always my support in the moments when

there was no one to answer my queries. My dearest wife, this is the end of our wish and

sorrows.

At the end, my endless love goes to my little boy Malek, who fills every day of our life

with joy and fun.

Amsterdam, 22 December 2016

LIST OF ABBREVIATIONS | 147

List of abbreviations

AH Ahaywa

BTC breakthrough curve

C carbon

CBD citrate bicarbonate dithionite

CEC cation exchange capacity

DOC dissolved organic carbon

DOM dissolved organic matter

HM heavy metal

IM initial mass

KO Kola

KW Kwamel

OM organic matter

PTF pedotransfer function

PZC point-of- zero charge

SOC soil organic carbon

SOM soil organic matter

SSA specific surface area

TC total carbon

TN total nitrogen

TOC total organic carbon

VC very coarse

VF very fine

WQ Wadi Qasab

XRD X-ray diffraction

148 | LIST OF PAPERS USED IN THIS THESIS

List of papers used in this thesis

I. Refaey, Y., Jansen, B., El-Shater, A., El-Haddad, A., Kalbitz, K. 2015 Clay

minerals of Pliocene deposits and their potential use for the purification of polluted

wastewater in the Sohag area, Egypt. Geoderma Regional 5, 215-225.

Laboratory work: Y. Refaey

Writing: Y. Refaey

Help in suggestion the sampling area: A. El-Shater and A. El-Haddad

Supervision and reviewing: K. Kalbitz and B. Jansen

II. Refaey, Y., Jansen, B., El-Shater, A., El-Haddad, A., Kalbitz, K. 2014. The role of

dissolved organic matter in adsorbing heavy metals in clay-rich soils. Vadose Zone

J., Vol. 13 No. 7.

Laboratory work: Y. Refaey

Writing: Y. Refaey

Help in suggestion the sampling area: A. El-Shater, A. El-Haddad

Supervision and reviewing: K. Kalbitz and B. Jansen

III. Refaey, Y., Jansen, B., Parsons, J., de Voogt, P., Bagnis, S., Markus, A., El-Shater,

A., El-Haddad, A., Kalbitz, K. 2016. Effects of clay minerals, hydroxides, and

timing of dissolved organic matter addition on the competitive sorption of Copper,

Nickel and Zinc: A column experiment. Revised version: Journal of Environmental

Management (ID: JEMA-D-16-02274).

Laboratory work: Y. Refaey and S. Bagnis

Modeling the data: Y. Refaey and A. Markus

Writing: Y. Refaey

Supervision and reviewing: P. de Voogt, K. Kalbitz, B. Janse, J. Parsons, A. El-Shater, and

A. El-Haddad

IV. Refaey, Y., Jansen, B., de Voogt, P., Parsons, J.B, El-Shater, A., El-Haddad, A.,

Kalbitz, K. 2016. The influence of organo-metal interactions on regeneration of

exhausted sorbent materials loaded with heavy metals. Under review (Pedosphere

Journal-ID: pedos201609457).

Laboratory work: Y. Refaey

Modeling the data: Y. Refaey

Writing: Y. Refaey

Supervision and reviewing: P. de Voogt, K. Kalbitz, B. Janse, J. Parsons, A. El-Shater, and

A. El-Haddad

CURRICULUM VITAE | 149

Curriculum Vitae

Yasser Refaey was born on 27 July 1982 in Sohag city, Egypt. He finished his

bachelor degree in Earth Sciences at the Sohag University in 2003. From 2003-

2004, he studied different pre-master courses in Geology. He worked as

assistant teacher at Sohag University from 2003 to 2012. For his master thesis

(2005-2008) he studied the mineralogical and geotechnical application of clay

minerals. During his master he was awarded a 6 months research-grant by the

Egyptian Ministry of Higher Education and Scientific Research to study the

advanced analytical tools used in identifying the different clay minerals at

theTechnical University of Munich, Germany. Before starting with his PhD

research he tutored several undergraduate courses for students at the geology

and chemistry departments of Sohag University. During his PhD at the

University of Amsterdam he co-supervised a master thesis and tutored an

undergraduate course (2015-2016).

150 |

Cover story:

The front cover shows a purification technique that was pictured on the wall of the tomb

of Amenophis II and Ramses in 1450 B.C. The figure pours a liquid into vases from the

cup and draws it off by the siphons. The drawing is modified (used with permission)

from a sketch that depicts a sedimentation apparatus and wick siphons in an American

Water Works Association book called The Quest for Pure Water: The History of Water

Purification from the Earliest Records to the Twentieth Century, authors Moses N.

Baker and Michael J. Taras, published in 1981.

152 |