solidification/stabilisation treatment of spiked sembrong river ...

SOLIDIFICATION/STABILISATION TREATMENT OF SPIKED SEMBRONG

RIVER SEDIMENTS USING CEMENT AND RICE HUSK ASH

MOHAMMED KABIR ALIYU

A thesis submitted in

fulfilment of the requirement for the award of the

Degree of Master of Civil Engineering

Faculty of Civil and Environmental Engineering

Universiti Tun Hussein Onn Malaysia

August 2015

v

ABSTRACT

Contaminated sediment represents a significant problem for the public health as well

as the environment. Solidification/Stabilisation (S/S) remediation technique was

employed in this study to treat river sediment spiked with three heavy metals. The

main objective of this research was to study the effect of replacing cement with rice

husk ash (RHA) on compressive strength and leaching of Pb, Cr and Cu from the

stabilised sediments. Artificially contaminated sediments were prepared by

individually spiking each sediment sample with solutions of Lead nitrate (Pb(NO3)2),

Copper sulphate (CuSO4.5H2O) and Potassium dichromate (K2Cr2O7) to achieve an

average of 1000 ppm target concentration of each element. Cement was added at

10% and rice husk ash at 5, 10 15 and 20% throughout to the total dry weight of

mixture, which were then cured at room temperature (27 ± 3oC) and humidity of 75 ±

5 % for 7, 14 and 28 days. Cylindrical samples were prepared with water - cement

ratio of 0.4. The effectiveness of the treatment was evaluated by performing

unconfined compressive strength (UCS) test on compacted samples and three

different leaching tests, namely Toxicity Characteristic Leaching Procedure (TCLP),

Synthetic Precipitation Leaching Procedure (SPLP) and Deionized Water Leaching

tests (DIW) at curing periods of 7, 14 and 28 days. X-ray diffraction (XRD) analysis

was used to study the reaction products and crystalline phases of the treated sediment

after 28 days in order to explain the mechanisms responsible for immobilization of

the heavy metals under study. The results showed that pH and strength were found to

have great influence on metal release. The UCS values of solidified samples at 7, 14

and 28 days exceeded the minimum landfill disposal limits of 0.34N/mm2 (340 kPa).

Similarly after 28 days of curing the concentration of the selected heavy metals in the

TCLP, SPLP and DIW leaching tests were also either undetected or below the

allowable leachability limits. Results have indicated that the partial replacement of

cement with RHA in the binder system has increased the strength and reduced

leachability of the treated compared to untreated sediment samples.

vi

ABSTRAK

Sedimen yang tercemar merupakan satu masalah yang penting kerana boleh

mempengaruhi kesihatan dan persekitaran. Teknik pemulihan Penstabilan/pemejalan

(P/P) telah digunakan dalam kajian ini untuk mengawal logam berat dalam sedimen

tercemar. Objektif utama kajian ini adalah untuk mengkaji kesan penambahan abu

sekam padi terhadap kekuatan dan kebolehan larut resapan daripada tiga logam berat

terpilih (Pb, Cu & Cr) dari sedimen yang telah distabilkan. Sedimen tercemar sintetik

telah disediakan dengan mencampurkan sampel dengan plumbum nitrat (Pb (NO3)2,

Kuprum sulfat (CuSO4.5H2O) dan Kromium (K2Cr2O3) untuk mencapai kepekatan

purata 1000 ppm. Simen ditambah pada 10% dan abu sekam padi pada 5, 10, 15 dan

20% untuk mengeraskan dan menstabilkan sedimen tercemar yang kemudiannya

diawet pada suhu bilik (27 ± 3oC) dengan kelembapan pada 75 ± 5 % selama 7, 14

dan 28 hari. Sampel silinder disediakan dengan nisbah air-simen 0.4. Keberkesanan

rawatan telah dinilai dengan melakukan ujian Kekuatan Mampatan Tak Terkurung

(UCS) dan tiga ujian larut lesap yang berbeza, iaitu Prosedur Larut Resap Ciri

Ketoksikan (TCLP), Prosedur Larut Resap Hujan Tiruan (SPLP) dan Ujian Larut

Resap Air Nyah ion (DIW) pada tempoh pengawetan 7,14 dan 28 hari. Hasil kajian

menunjukkan bahawa di antara semua parameter eksperimen yang dipertimbangkan,

pH dan kekuatan didapati mempunyai pengaruh yang besar terhadap pelepasan

logam. Nilai UCS sampel pejal pada 7, 14 dan 28 hari melebihi had minimum tapak

pelupusan 0.34 N/mm2 (340 kPa). Kepekatan tiga logam berat terpilih dalam ujian

TCLP, SPLP dan DIW selepas tempoh pengawetan 28 hari adalah sama ada tidak

dikesan atau dibawah had larut resap USEPA. Analisis sinar-X (XRD) telah

digunakan untuk menjelaskan mekanisma yang terlibat dalam pelumpuhan logam

berat yang dikaji. Keputusan telah menunjukkan bahawa penggantian sebahagian

simen dengan abu sekam padi dalam sistem bahan pengikat yang telah meningkatkan

dan mengurangkan kebolehan larut resapan semua sampel sedimen pejal berbanding

dengan sampel yang tidak dirawat.

vii

TABLE OF CONTENTS

DECLARATION ii

DEDICATION iii

ACKNOWLEDGEMENT iv

ABSTRACT v

TABLE OF CONTENTS vi

LIST OF TABLES x

LIST OF FIGURES xii

LIST OF ABBREVATIONS xiv

CHAPTER 1 INTRODUCTION 1

1.1 Background of the Research 1

1.2 Problem Statement 3

1.3 Objectives of the Research 4

1.4 Scope of the Research 4

1.5 Significance of Research 5

CHAPTER 2 LITERATURE REVIEW 6

2.1 Introduction 6

2.2 Contaminated Sediments 9

2.3 Heavy Metals 11

2.4 Sources of Heavy Metal Pollution 12

2.4.1 Natural sources 12

2.4.2 Anthropogenic sources 12

viii

2.5 Heavy Metal Effects 13

2.5.1 Lead 13

2.5.2 Chromium 14

2.5.3 Copper 15

2.6 Remediation techniques 17

2.6.1 Electrochemical remediation 18

2.6.2 Soil Washing 18

2.6.3 Chemical oxidation 19

2.6.4 Phytoremediation 19

2.7 Solidification/Stabilisation as a remediation

technology for the treatment of contaminated

sediments 20

2.7.1 Solidification 20

2.7.2 Stabilisation 21

2.8 Common Binders Used for the S/S Treatment

Technology 21

2.8.1 Cement 23

2.8.2 Rice husk ash as an additive in

Solidification/Stabilisation 24

2.8.3 Compressive Strength of Cement-RHA

Concrete 27

2.9 The Mechanisms of Heavy Metal Binding 29

2.9.1 Sorption 29

2.9.2 Complexation 29

2.9.3 Precipitation 30

2.10 Factors that can affect the Strength of

Stabilised Sediments 31

2.10.1 Organic Matter 31

ix

2.10.2 Sulphates 31

2.10.3 Sulphides 32

2.10.4 Compaction 32

2.10.5 Moisture Content 33

2.10.6 Temperature 33

2.10.7 Freeze-Thaw and Dry-Wet Effect 33

2.11 Solidified/Stabilised waste acceptance criteria 34

2.12 Factors that Influence Mobility of Heavy

Metals in Sediments 34

2.12.1 Effect of pH on the Immobilisation/Leaching

of Heavy metals 35

2.12.2 Effect of pH on mineral surface charge

development 36

2.12.3 Effect of complexing agents 37

2.13 LEACHING TESTS 37

2.13.1 Toxicity Characteristic Leaching Procedure

(Method 1311) 38

2.13.2 Synthetic Precipitation Leaching Procedure

(Method 1312) 38

2.13.3 Deionised water leaching test (DIW) 39

2.13.4 Leaching Mechanisms 39

2.13.5 Parameters controlling leaching tests under

laboratory conditions 40

2.14 Previous investigations of Heavy metals

retention and leachability using

Solidification/Stabilisation technology 44

CHAPTER 3 RESEARCH METHODOLOGY 47

3.1 Introduction 47

3.2 Raw materials used 52

x

3.2.1 Sediment samples 52

3.2.2 Cement 53

3.2.3 Rice Husk Ash (RHA) 53

3.3 X- ray fluorescence (XRF) Analysis 54

3.4 Sediment contamination by spiking 55

3.5 Binder System 56

3.6 Solidification/Stabilization Sample preparation 57

3.7 Unconfined Compressive Strength (UCS) Test 59

3.8 Leaching Tests 61

3.8.1 Toxicity characteristic leaching procedure,

Method 1311 (U.S. EPA 1992) 62

3.8.2 Synthetic Precipitation Leaching Procedure

Method 1312 (U.S. EPA 1994) 64

3.8.3 Deionised Water Leaching test (Control) 65

3.9 X-ray diffraction (XRD) 65

CHAPTER 4 ANALYSIS AND DISCUSSIONS 67

4.1 Introduction 67

4.2 Physical properties of the sediments 67

4.3 Chemical characteristics of Rice Husk Ash,

Sediments and Cement 68

4.4 Concentration of contaminants in the

sediments before and after spiking 71

4.5 Bulk densities of the stabilised sediments 72

4.6 Unconfined Compression Test 73

4.6.1 UCS of Stabilised/Solidified lead spiked

sediment samples 74

4.6.2 Relationship between strength and density of

Lead spiked sediment samples 76

xi

4.7 UCS of Stabilised/Solidified Chromium spiked

sediment samples 78


chromium spiked samples 79

4.8 UCS of Stabilised/Solidified Copper spiked

sediment samples 82


Copper spiked sediment 84

4.9 The effect of curing on strength gains due to

similar mix ratio for Pb, Cr and Cu at 7, 14

and 28 days 86

4.10 LEACHING TESTS 89

4.10.1 Leachability characteristics of lead stabilised

sediment 89

4.10.1.1 Relationship between strength and leachability of

lead spiked sediment samples 92

4.10.2 Leachability of chromium spiked sediment 95


chromium spiked sediment sample 97

4.10.3 Leachability of Copper spiked sediment 100


copper spiked sediment sample 102

4.10.4 Leachability of all the three spiked elements in

the TCLP, SPLP and DIW 105

4.10.5 The effect of pH on immobilisation/leaching

of stabilised lead, copper and chromium

spiked sediments 106

4.11 X- Ray Diffraction Analysis results 118

4.11.1 XRD analysis of the Sembrong river sediment 118

xii

4.11.2 XRD of Lead spiked sediments 119

4.11.3 XRD of Chromium spiked sediment 120

4.11.4 XRD of Copper spiked sediment 122

CHAPTER 5 CONCLUSIONS AND RECOMMENDATIONS 124

5.1 Introduction 124

5.2 Important Observations 124

5.3 Recommendations 127

APPENDICES 152

xiii

LIST OF TABLES

2.1 Common binders and additives used in S/S

treatment Technology 23

2.2 Binder/Additives that have been used in the process

of S/S of contaminated soil/sediments 27

2.3 Stabilised/Solidified waste acceptance criteria 34

2.4 Comparison of the three leaching tests 39

3.1 Chemical composition of cement 53

3.2 Summarised weight of contaminants spiked on the

sediments with their target concentrations 56

3.3 Mix Ratios for unconfined compression Test

(Cement + RHA + spiked sediments) 59

4.1 Properties of the Sembrong river sediment 68

4.2 Chemical compositions of sediment, cement and rice

husk ash using XRF 70

4.3 The initial concentration of contaminants in the

sediments before spiking using ICP-MS. 71

4.4 The recovered concentrations after spiking of Lead,

Chromium, and Copper in the spiked sediments sample

matrix using XRF 72

4.5 Bulk density of lead spiked sediment sample at 7, 14

and 28 days 73

4.6 Bulk density of Chromium spiked sediment sample at

7, 14 and 28 days 73

4.7 Bulk density of Copper spiked sediment sample at 7, 14

and 28 days 73

xiv

4.8 UCS of Solidified/Stabilised Lead spiked samples

throughout 28 days of curing 75

4. 9 Relationship between strength and density of lead

spiked sample 77

4.10 UCS of Solidified/Stabilised Chromium spiked samples


4.11 Relationship between strength and density of chromium

spiked sample 80

4.12 UCS of solidified/stabilised Copper spiked samples


4.13 Relationship between strength and density of copper

stabilised sample 85

4.14 The effect of curing period on the strength of Pb, Cr

and Cu at 7 days 87

4.15 The effect of curing period on the strength of Pb, Cr

and Cu at 14 days 88

4.16 The comparison on the effect of curing period on

strength of Pb, Cr and Cu at 28 days 88

4.17 Concentration of lead (mg/l) in the TCLP leachate 91

4.18 Concentration of lead (mg/l) in the SPLP leachate 91

4.19 Concentration of lead (mg/l) in the DIW leachate 91

4.20 Concentration of Chromium (mg/l) in the TCLP

leachate 96

4.21 Concentration of Chromium (mg/l) in the SPLP

leachate 96

4.22 Concentration of Chromium (mg/l) in the DIW leachate 97

4.23 Concentration of Copper (mg/l) in the TCLP leachate 101

4.24 Concentration of Copper (mg/l) in the SPLP leachate 101

4.25 Concentration of Copper (mg/l) in the DIW leachate 101

B.1 Concentration of Lead (Pb), Chromium (Cr), & Copper

(Cu) in ppm after spiking 157

B.2 Chemical composition of rice husk ash burnt at 700 158

B.3 Chemical composition of cement (ppm) 159

xv

C.1 Qualitative analysis result of heavy metals using

ICP-MS sample A 160

C.2 Qualitative analysis result of heavy metals in the

sediment using ICP-MS sample B 162


sediment using ICP-MS sample C 164


sediment using ICP-MS sample D 166

C.5 Quantitative analysis of some selected heavy metals

in sample A using ICP-MS 168

C. 6 Quantitative analysis of some selected heavy metals

in sample B using ICP-MS 169


in sample C using ICP-MS 170


in sample D using ICP-MS 171

D.1 Datasheet for the Lead spiked sediment sample 172

D.2 Datasheet for the Copper spiked sediment sample 173

D.3 Datasheet for the Chromium spiked sediment sample 174

E.1 pH of the TCLP leachates before and after filtration

of Lead spiked sediment for 7, 14 & 28 days 175

E.2 pH of the TCLP leachates for Chromium spiked

sediment for 7, 14 & 28 days 176

E.3 pH of the TCLP leachates for Copper spiked


E.4 pH of the SPLP leachates for lead spiked sediment

for 7, 14 & 28 days 178

E.5 pH of the SPLP leachates for Chromium spiked

sediment 179

E.6 pH of the SPLP leachates for Copper spiked


E.7 pH of the DIW leachates for Copper spiked


xvi

E.8 pH of the DIW leachates for lead spiked sediment

for 7, 14 & 28 days 182

E.9 pH of the DIW leachates for Chromium spiked


F.1 Relationship between leachability with pH of lead at

7 days in the TCLP 184






7 days in the SPLP 185






7 days in the DIW 186





F.10 Relationship between leachability with pH of

chromium at 7 days in the TCLP 187






chromium at 7 days in the SPLP 188





xvii


chromium at 7 days in the DIW 189





F.19 Relationship between leachability with pH of copper

at 7 days in the TCLP 190






at 7 days in the SPLP 191






at 7 days in the DIW 192





G.1 Concentrations of Lead (Pb) at 7, 14 and 28 days

after leaching tests (TCLP, SPLP & DIW) 193

H.1 Concentration of Chromium (Cr) at 7, 14 and 28

days after leaching tests (TCLP, SPLP & DIW) 208

I.1 Concentration of Copper (Cu) at 7, 14 and 28 days

after leaching tests (TCLP, SPLP & DIW) 223

J. 1 Calculations for the required quantity of each heavy

metal to be spiked 237

K.1 Relationship of strength and leachability of lead in

the TCLP at 7, 14 and 28 days 239

xviii


the SPLP at 7, 14 and 28 days 239


the DIW at 7, 14 and 28 days 239

K.4 Relationship of strength and leachability of Copper

in the TCLP at 7, 14 and 28 days 240


in the SPLP at 7, 14 and 28 days 240


in the DIW at 7, 14 and 28 days 240

K.7 Relationship of strength and leachability of

Chromium in the TCLP at 7, 14 and 28 days 241


Chromium in the SPLP at 7, 14 and 28 days 241


Chromium in the DIW at 7, 14 and 28 days 241

xix

LIST OF FIGURES

3.1 The location where sediment samples were taken (1°

52'.18.44'' N and 103° 06‘15.71'' E) 48

3.2 Sediment core sampler, (Beeker type, Netherlands) 48

3.3 Methodology flow chart (Physical and chemical tests) 50

3.4 Methodology flow chart (Mechanical test) 51

3.5 Methodology flow chart (Leaching tests) 51

3.6 Drying Sediment samples 52

3.7 Ground Rice husk ash 54

3.8 X-ray Fluorescence machine (S4 Pioneer, Bruker aXS

Germany) 55

3.9 ESM-989, Baker stand mixer, Japan 58

3.10 Plastic cylindrical mould 58

3.11 Unconfined compression test machine (LoadTrac II,

Geocomp, USA) 60

3.12 solidified specimen placed between the top and lower platens 61

3.13 500-mL plastic bottles with crushed sample mixed with

acetic acid ratio 1:20 before extraction 62

3.14 The End - over - End rotating extractor 64

3.15 Samples of filtered leachates in the chiller (< 4°C) before

metal analysis 65

3.16 X-ray diffraction equipment (D8 advance, Bruker, Germany) 66

4. 1 X-ray diffraction pattern of rice husk ash burnt at 700 ºC 70

4.2 UCS development of Lead spiked sediment samples

throughout 28 days curing period 75

4.3 Relationship between compressive strength and density of

lead spiked sample at 7 days 77

xx





4.6 UCS developments of chromium spiked sediment samples



Chromium spiked sample at 7 days 81





4.10 UCS developments of Copper spiked sediment samples


4.11 Correlation between compressive strength and density of

Copper spiked sample at 7 day 85


Copper spiked sample at 14 days 85


Copper spiked sample at 28 days 86

4.14. The effect of curing period on strength for Pb, Cr and cu

stabilized samples at 7 days 88

4.15 The effect of curing period on strength for Pb, Cr and Cu


4.16 The effect of curing period on strength for Pb, Cr and Cu


4.17 Effect of replacement of cement with RHA on the leachabilty

of lead (PbNO3) in TCLP, SPLP & DIW leaching tests 92

4.18 Relationship between strength and leachability of lead for the

TCLP at 7, 14 and 28 days 94

4.19 Relationship between strength and leachability of lead for the

SPLP at 7, 14 and 28 days 94

4.20 Relationship of compressive strength with leachability of lead

for the DIW at 7, 14 and 28 days 95

xxi

4.21 Effect of replacement of cement with RHA on the

leachability of stabilised Chromium (K2Cr2O7) in TCLP,

SPLP & DIW leaching tests 97

4. 22 Relationship between compressive strength and leachability

of Chromium in the TCLP at 7, 14 and 28 days 99

4.23 Relationship between compressive strength and leachability

of Chromium in the SPLP at 7, 14 and 28 days 99

4.24 Relationship between compressive strength and leachability

of Chromium in the DIW at 7, 14 and 28 days 100

4.25 Effect of replacement of cement with RHA on the leachabilty

of stabilised Copper (Cu) in TCLP, SPLP & DIW leaching

tests 102

4.26 Relationship between strength and leachability of copper in

the TCLP at 7, 14 and 28 days 104


the SPLP at 7, 14 and 28 days 104


the DIW at 7, 14 and 28 days 105

4.29 The effect of pH on the leachability of lead in the TCLP at 7

days 109

4. 30 The effect of pH on the leachability of lead in the TCLP at 14

days 109

4.31 The effect of pH on the leachability of lead in the TCLP at

28 days 109

4.32 The effect of pH on the leachability of lead in the SPLP at 7

days 110

4.33 The effect of pH on the leachability of lead in the SPLP at 14

days 110

4. 34 The effect of pH on the leachability of lead in the SPLP at 28

days 110

4.35 The effect of pH on the leachability of lead in the DIW at 7

days 111


days 111

xxii


days 111

4.38 The effect of pH on the leachability of Copper in the TCLP at

7 days 112


14 days 112


28 days 112

4.41 The effect of pH on the leachability of Copper in the SPLP at

7 days 113

4. 42 The effect of pH on the leachability of Copper in the SPLP at

14 days 113

4. 43 The effect of pH on the leachability of Copper in the SPLP at

28 days 113

4.44 The effect of pH on the leachability of Copper in the DIW at

7 days 114

4. 45 The effect of pH on the leachability of Copper in the DIW at

14 days 114

4.46 The effect of pH on the leachability of Copper in the DIW at

28 days 114

4.47 The effect of pH on the leachability of Chromium in the

TCLP at 7 days 115


TCLP at 14 days 115


TCLP at 28 days 115


SPLP at 7 days 116


SPLP at 14 days 116


SPLP at 28 days 116

4.53 The effect of pH on the leachability of Chromium in the DIW

at 7 days 117

xxiii

4. 54 The effect of pH on the leachability of Chromium in the DIW

at 14 days 117

4.55 The effect of pH on the leachability of Chromium in the DIW

at 28 days 117

4. 56 XRD patterns of Sembrong river sediment 119

4.57 XRD patterns of Lead spiked sediment for Pb28.100:0:0 and

Pb28d.70:10:20 after 28 days 120

4.58 XRD patterns of Chromium spiked sediment for

Cr28d.85:10:5 and Cr28d.75:10:15 after 28 days 121

4.59 XRD patterns of Copper spiked sediment for Cu28d.100:0:0

and Cu28d.75:10:15 after 28 days 123

A.1 Extracting sediment sample from the river using core sampler 152

A.2 Sediment sample completely taken out from the river 152

A.3 FL 50 Kiln furnace, Kilns & Furnaces Ltd, England 153

A.4 Ball mill grinder (pulverisette 6, Fritsch, Germany) 153

A.5 The customized compaction plastic mould and tools for the

preparation of unconfined specimens 154

A.6 Plastic curing box 154

A.7 The atomic absorption spectrometer (AAS) (AAnalyst 800,

Perkin Elmer, USA) 155

A.8 ICP-MS (ELAN 9000, Perkin Elmer, USA) 155

A.9 The vacuum pressure filtration 156

A.10 Holcim (Portland composite cement) 156

xxiv

LIST OF SYMBOLS AND ABBREVIATIONS

AAS - Atomic Absorption Spectroscopy

Al2O3 - Alumina

ASTM - American Society for Testing and Materials

C - cement

C2S - dicalcium silicate

C2SHx, C3S2Hx - Hydrated calcium silicate

C3AHx, C4AHx - hydrated calcium aluminates

C3S - tricalcium silicate

Ca (OH)2 - Calcium Hydroxide ( portlandite)

CaO - calcium oxide

cm - centimetre

C-S-H - Calcium Silicate Hydrate

CuSO4.5H2O - copper sulphate penta hydrate

DIW - Deionised Water Leaching Test

EPA - Environmental Protection Agency

Ggbs - Ground Granulated Blast Furnance Slag

Gs - specific gravity

ICDD - International Centre For Diffraction Data

xxv

ICP- MS - Inductively Couple Plasma Mass Spectroscopy

K2Cr2O7 - potassium dichromate

kg - kilogram

kPa - kilo Pascal

L/S - liquid to solid ratio

m - metre

mm - milimetre

Pb (NO3)2 - Lead nitrate

PFA - Pulverised Fuel Ash

qu - unconfined compressive strength

RHA - Rice Husk Ash

S - Sediment

S/S - Solidification/Stabilisation

SiO2 - silicon dioxide

SPLP - Synthetic Precipitation Leaching Procedure

TCLP - Toxicity Characteristic Leaching Procedure

UCS - Unconfined Compressive Strength test

USEPA - United States Environmental Protection Agency

UTHM - Universiti Tun Hussein Onn Malaysia

w - moisture content

ws - dry weight

ww - wet weight

XRD - X – Ray Diffraction

xxvi

XRF - X- Ray Fluorescence

μm - micro metre

xxvii

LIST OF APPENDICES

APPENDIX TITLE PAGE

A Tools and materials for the experimental 152

B Results of XRF tests 157

C Results of ICP-MS (Qualitative & Quantitative) 160

D TCLP Data sheet 172

E pH of the leachates before and after filtration 175

F Results showing the relationship of pH on the

The leachability of lead, chromium & copper 184

G Concentration of lead at 7, 14 and 28 days after

leaching tests (TCLP, SPLP & DIW) 193

H Concentration of chromium at 7, 14 and 28 days after


I Concentration of chromium at 7, 14 and 28 days after


J Calculations for the required quantity of metal

to be spiked 237

K Results for the relationship between strength and leachability

of lead, chromium and copper 239

CHAPTER 1

INTRODUCTION

1.1 Background of the Research

Solidification/Stabilisation (S/S) is a treatment technique by which contaminated

soils, sediments or waste are mixed with a binder and/or specific additives with the

aim of decreasing the mobility of the toxic contaminants by increasing the pH and

fully or partially binding the contaminants in the solid matrix (stabilisation). It is also

for improving the physical properties (strength, compressibility, permeability and

durability) of the final treatment products (solidification) (Antemir et al., 2010).

Solidification/Stabilisation technology was originally developed for the

treatment of nuclear waste in 1950s and later on different types of hazardous wastes

since 1970‘s (Conner, 1990). From around 1980s the technology was also applied for

the treatment of contaminated soil and sediments (Laugesen 2007). Interesting

example have been found in the case of treatment of contaminated sediments in

Norway using Stabilization/Solidification technologies as is the case for Trondheim

sediments (Arevalo, 2008).

Sediments can be defined as a collection of fine, medium and coarse grain

minerals and organic particles that are found at the bottom of lakes, ponds, rivers

streams, bays, estuaries, and oceans (Adams et al., 1992). Sediments are essential

components of aquatic and marine ecosystems where they provide habitat for a wide

variety of benthic organisms as well as juvenile forms of pelagic organisms.

Sediment has been described as the ―ultimate sink‖ or storage place for pollutants.

Unfortunately due to resuspension sediment can function as both a sink and a source

2

for contaminants in the aquatic environment (USEPA, 1997). Many toxic

contaminants that are barely detectable in the water column can accumulate in

sediments at much higher levels; the water column can continue to be contaminated

long after the source of pollutants was controlled. Sediments are the ultimate

reservoir for the numerous potential chemical and biological contaminants that may

be contained in effluents originating from urban, agricultural, and industrial lands

and recreational activities. Contaminated sediments in rivers and streams, lakes,

coastal harbours, and estuaries have the potential to pose ecological and human

health risks.

Most of the sediments in rivers, lakes, and oceans have been contaminated by

pollutants and most of the contaminants were released years ago while other

contaminants enter water every day. Some contaminants flow directly from industrial

and municipal waste dischargers, while others come from polluted runoff in urban

and agricultural areas. Still other contaminants are carried through the air, landing in

lakes and streams far from the factories and other facilities that produced them

(USEPA, 2003). There is always the threat of re-suspension of the contaminants into

the water column which presents its own set of environmental threats to wildlife and

humans (USEPA, 2003).

A wide variety of organic compounds and metals are discharged into

estuaries from industrial, agricultural, and urban sources. The contaminants are

adsorbed onto suspended particles and eventually settle to the sediments. There they

can exert toxic effects on the benthic community that lives in the sediments and can

indirectly affect human health as well.

Sediments are the major sinks for heavy metals released into the environment

by anthropogenic activities and unlike organic contaminants which are oxidized to

carbon (IV) oxide by microbial action, most metals do not undergo microbial or

chemical degradation (Kirpichtchikoya et al., 2006) and their total concentration in

soils persists for a long time after their introduction (Adriano, 2003). Heavy metal

contamination of sediments may pose risks and hazards to humans and the ecosystem

through direct ingestion or contact with contaminated sediments. Heavy metal

contamination of sediments continues to be a problem with few practical or

applicable remediation technologies, because of the importance of benthic organisms

such as worms, crustaceans, insect larvae, and microbes to the aquatic food chain.

Contaminated sediments introduce pollutants by the process of bioaccumulation

3

which invariably move up the food chain and eventually are consumed by humans.

In most agricultural- based countries, such as Malaysia, Philippines, Thailand and

Indonesia, the proper disposal of rice husk has become a problem, especially when

open burning is no longer permitted due to environmental concerns (Sarkawi & Aziz,

2003). The problems caused by irresponsible dumping of this agricultural waste can

be meaningfully reduced by finding its suitable engineering applications, such as

being used in stabilisation/solidification of contaminated soil/sediments. The lower

cost makes it an attractive alternative if adequate performance can be obtained.

In an effort to reuse the large quantities of the RHA, this study was carried

out to examine the possibility of utilising the waste material with cement in

contaminated sediment stabilization. The effectiveness of RHA in contaminated

sediment stabilisation was determined by conducting unconfined compressive

strength test (UCS) to measure the strength of the stabilised samples and leaching

tests to determine the leachability of inorganic contaminants (Heavy metals). X- ray

diffraction (XRD) was used to investigate the crystalline mineral phases responsible

for Lead, Chromium and Copper immobilisation in the stabilised sediment samples.

1.2 Problem Statement

Contaminated sediments represent a significant problem for the public‘s health as

well as for the environment especially in urban coastal regions (Vasconcelos et al.,

2007) as such environmentally motivated remediation efforts have become

increasingly relevant to solve the problem. Current researches on S/S treatment of

metal contaminated soils were concentrated primarily on using Portland cement

systems or combination of other established pozzolans such as pulverised fly ash

(PFA) and lime as reflected in studies conducted by Boardman (1999), Musta &

Kassim (2000), Dermatas & Meng (2001) and Wang & Vipulanandan (2001).

Although these S/S systems exhibit excellent treatment effectiveness, its applications

in Asian countries have weaknesses which include relatively high costs of cement

and lime as well as limited availability of mass amount of fly ashes. Besides, these

countries are experiencing difficulty in disposal of rice husk waste due to their

abundance. Concrete technologists are gradually finding applications in rice husk ash

(RHA) as an additive for producing high-strength concrete. However, few studies

4

have been conducted on usage of rice husk ash on S/S of contaminated

soils/sediments, So the use of rice husk ash, an indigenous agro-waste as a

supplementary binder to cement for treatment of contaminated sediments with heavy

metals not only marks a new innovation in the S/S technology but also assists in

easing disposal problem of rice husk heaps in Asian countries.

Much of the work on solidification/stabilisation of metals in soil/sediments

focused mainly on using cement alone or with other pozzolans, but very little

research has been done on the use of rice husk ash with cement for the treatment of

dissolved lead (Pb), Chromium (Cr) and Copper (Cu) in sediments.

1.3 Objectives of the Research

The objectives of this research are:

1. To determine the effect of adding 5%, 10%, 15% and 20% of rice husk ash

(RHA) incorporated into cement on contaminant immobilization of the spiked

Sembrong river sediments.

2. To determine the unconfined compressive strength of the solidified/stabilised

sediment sample and to establish the relationships between strength and

leachability.

3. To determine the potential leachability of the contaminants through

laboratory leaching tests (TCLP, SPLP and DIW) and to perform X-ray

diffraction (XRD) analysis so as to explain the mechanism that is responsible

for the immobilisation of the heavy metals under study.

4. To establish the relationship between the leachability of lead, chromium and

copper with change in pH.

1.4 Scope of the Research

The sediment used in this study was taken from Sembrong river at Parit Sempadan

near Universiti Tun Hussein Onn Malaysia (UTHM). This study is concentrated on

laboratory analysis of the Sembrong river sediment; the treatment of the spiked

sediments was done using Solidification/Stabilization treatment method using

cement and rice husk ash as binder. Test specimens were prepared with different mix

5

proportions of RHA (5%, 10%, 15% and 20%) and cement (10%) to total dry weight

of the mixture. The test specimens were then cured for 7, 14 and 28 days respectively

prior to tests. Three different leaching tests namely Toxicity characteristic leaching

procedure (TCLP), Synthetic precipitation leaching procedure (SPLP) and Deionised

water leaching DIW) were used to assess the leachability of the spiked contaminants

(Pb, Cr and Cu). Unconfined compressive strength test was carried out to compare

the strength of the stabilised sample before and after treating with different

concentration of cement and rice husk ash. X-ray diffraction (XRD) analysis was

also done to identify the mechanism responsible for the heavy metals immobilisation.

1.5 Significance of Research

The significance of the research was based on the need to enhance S/S of

contaminated sediments by adding rice husk ash to cement where optimum mix

proportions for cement- RHA would be established. In this study, the stabilised

specimens were observed in strength and also in contaminants leachability.

The results from this project will indicate whether rice husk ash may be

added as a substitution of cement to enhance the strength and reduced the

leachability of the heavy metals contaminated sediments. Considering the high cost

of cement, the utilization of RHA instead of dumping as a waste material provide a

significant contribution for the country‘s economy and solution of the environmental

pollution problem.

Furthermore, the characteristics of the stabilized sediment with cement-RHA

can contribute to knowledge and also the findings can be used as a reference by

future researchers in their research.

CHAPTER 2

LITERATURE REVIEW

2.1 Introduction

Sediments are the eventual reservoir for the numerous potential chemical and

biological contaminants that may be contained in effluents originating from urban,

agricultural, and industrial lands and recreational activities. Contaminated sediments

in rivers and streams, ponds, shoreline harbours, and estuaries have the potential to

pose ecological and human health risks. It has been shown in numerous studies in

which water quality criteria are not exceeded that adverse effects are possible in

aquatic organisms that reside or forage in or near sediment (Chapman, 1989). It is

widely understood that sediment contamination can have many detrimental effects on

an aquatic ecosystem, some of which may be readily evident and others more subtle

or unknown. In most receiving waters; however the effects are difficult to observe

and require the use of a variety of investigation and risk assessment tools, such as

benthic macro invertebrate community analyses, chemical testing, hydrodynamic and

sediment transport modelling, habitat analysis, and toxicity testing (Wenning &

Ingersoll 2002). Sediments house many contaminants and therefore pose the highest

risk to the aquatic environment as a source of pollution (Bervoets et al., 1994 and

Williamson et al., 1996).

Environmental pollution by heavy metals impacts negatively on human

health. Their remediation proves to be problematic due to the persistence and non-

degradability of heavy metals (Yuan et al., 2004). High concentrations of heavy

metals in biota can be linked to high concentration in sediments. The bioavailable

7

metal load in sediments may affect the distribution and composition of benthic

assemblages (Kress et al., 2004) and this can be linked to high concentration

recorded in living organisms (Pempkowiak et al., 1999).

Contaminated sediment investigations have features that make them more

complex than water evaluations and to a lesser degree, soil or terrestrial

investigations (National Research Council, 2001). The simple fact that sediments lie

under water makes measurement, observation, and mapping of contaminant and

ecosystem characteristics technically challenging and expensive. Sediments put

together contaminant input from many sources within a watershed or coastal region,

creating difficulties in tracking the potential sources of contamination. This can lead

to ubiquitous, regional ―background‖ levels of anthropogenic contaminants that are

difficult to separate from site specific sources (Crommentuijn et al., 2000).

For the same reasons, sediments are often contaminated with multiple

chemicals (Long et al., 1995), making risk assessment and management decision-

making difficult and complex. The hydrodynamics and geochemistry of aquatic

ecosystems are also quite different than those of terrestrial ecosystems. While soils

and groundwater can often be isolated from receptors during remediation, similar

isolation or removal approaches for contaminated sediments are more difficult to

implement successfully; sensitive aquatic biota are more likely, and at times

unavoidably, directly affected during the implementation of the remedy (USEPA

2002a), because the benthic community in direct contact with sediments is mostly

near the base of the aquatic food chain, clean up targets can be orders of magnitude

lower than those at most contaminated land sites. Together, these and other factors

often push the limits of assessment methods and clean up technologies for sediment

and can increase costs significantly over what may be needed to address similar

contaminant conditions in soil and groundwater. In addition, while the benefits of

ownership and clean-up of contaminated land, which can subsequently be sold or

developed (or both) to offset the costs of remediation, are clear, such benefits are less

obvious in aquatic ecosystems.

The potential to harm benthic organisms is not the only adverse impact of

contaminated sediments, they serve as diffuse sources of contamination to the

overlying water body; slowly releasing the contaminant back into the water column

(Marcus, 1991; DEC, 1989). Contaminated sediment comprise of a range of

materials that settle to the bottom of any water body. It includes the shells and

8

coverings of molluscs and other water animals, transported soil particles from

surface erosion, organic matter from dead and rotting vegetation and animals,

sewage, industrial wastes, organic materials, inorganic materials, and chemicals.

EPA defines sediment as soil, sand, and minerals washed from land into water

usually after rain (USEPA, 1988c). Current regulatory trends tend to separate

sediment/soil matrices from sludge. Surface waters from different part of the world

receive discharges of various liquid and solid wastes from industrial and municipal

operations, agricultural and urban runoffs, accidental spills, leaks, dumping of waste,

and precipitation carrying pollutants from the atmosphere.

Contamination is a concept that is not always clearly defined relative to

sediments. The mere presence of a foreign substance in sediment could be construed

as contamination. However, the presence of a foreign substance does not necessarily

mean it is harmful. Metals can be present in naturally occurring concentrations

(background levels) in species or forms that are not harmful to aquatic life. While

there are no naturally occurring background concentrations for synthetic organic

compounds, the presence of a synthetic organic compound does not necessarily

imply harm. Some evaluation must be made to estimate the potential risk to aquatic

life or human health that the compound will have (Pataki et al., 1999)

There are basically two sources of pollution, point and non-point sources. The

point sources are discharged from one particular location or point, such as municipal

and industrial plant. The Non-point sources are devoid of any particular or discrete

location, which include runoff from agricultural lands, soil entrainment and

atmospheric deposition and other sources such as spills, contaminated groundwater

infiltration, aquatic dumping (Van der Perk, 2006)

Modern agriculture is now becoming a nuisance to mankind. The

insecticides, pesticides, chemical fertilizers especially nitrate and phosphate are used

annually to boost agricultural production and these chemicals are washed down the

soil by rain and eventually end up to contaminate the ground and stream water ways.

Sembrong river is equally surrounded by these types of activities which are likely to

pollute the water way.

Many of these discharges contain toxic/hazardous materials that settle in

sediment and persist in the environment for long periods of time. These contaminated

sediments may affect human health and the environment and can cause losses of

important resources such as drinking water. Humans can be exposed to the

9

contaminants through such means as infiltration into drinking water, accumulation in

the food chain, and direct dermal contact. Animals of the benthic community can

absorb toxic substances from their surroundings. Contaminated sediment can be

lethal to them and affect the food chains of larger animals such as; fish, birds, and

man (EPA, 1993).

The EPA has defined a contaminant as: "Any solid, liquid, semisolid,

dissolved solid, gaseous material, or disease-causing agent which upon exposure,

ingestion, inhalation, or assimilation into any organism, either directly from the

environment or indirectly by ingestion through food chains, may pose a risk of or

cause death, disease, behavioural abnormalities, cancer, genetic mutations,

physiological malfunctions or physical deformations, in the organism or their

offspring" (EPA, 1992). This definition clearly explains that a contaminant is not

simply the presence of a foreign substance, but an element of harm to some

organism, species, population, or community must be involved. Sediment

contaminants primarily consist of heavy metals and persistent organic compounds

(EPA, 1990). Several factors strongly influence the extent and severity of

contamination by these toxic compounds. Fine-grained sediments high in organic

matter are better able to adsorb the pollutants than are coarser particles. The finer

particles are also more likely to be resuspended by currents and transported to

regions far from their point of origin. For these reasons, silty muds usually contain

the highest concentrations of contaminants.

2.2 Contaminated Sediments

One of the concerned environmental issues these days is contaminated sediments. As

explained above, the presence of aquatic organisms on and inside sediments makes

them important for their well-being and health. If sediments become contaminated,

they can pose a threat to sediment dwelling habitats (Ingersoll et al., 1995). Threats

can be of different types and intensities, e.g. damaged reproduction of fish and other

invertebrates, declined rate of aquatic organism growth, contaminants

bioaccumulation in aquatic plants or animals and even death. Smaller aquatic

organisms are located at the base of food webs. If they get contaminated through

these steps, they can die due to the toxicity of the sediments so larger organisms at

10

top of the food web lose their food (EPA, 1999), if they survive the contamination,

they easily can transmit it to bigger members of the food chain including terrestrial

animals and human. This is how fish, benthic organisms, birds and mammals can be

touched by the impacts of contaminated sediments through the connections of food

webs. Not only can human life be harmed by contaminated sediments indirectly but

also through direct ways. This can be achieved through direct exposure to

contaminated sediments via recreational activities, swimming in waters that have

contaminated sediments and more. As it is seen here, contaminated sediments in

aquatic environments are potential hazards to any related organisms, whether

sediment-dwelling and any species that depends on them or terrestrial organisms and

humans (Ingersoll et al., 1995; EPA, 1999).

Inorganic matter cannot be degraded but transformed into compounds having

more or less mobility or toxicity than their original form (Hamby, 1996). Due to

indestructibility of metals, remediation techniques are aim to extract, stabilize or

concentrate. Metals bind relatively fast to solid particles and often need to be

mobilised in order to extract them (NTV, 1993). Most metals are less mobile in

alkaline oxidized environments and therefore stabilisation of metals often means

adjusting pH by different kinds of improvements (cement, rice husk ash, lime, fly

ash) to slightly basic. Hence metals are immobilised but still present (Kumpiene et

al., 2007)

Heavy metals of major concern are Cd, Pb, Cr, Cu, Hg and arsenic

(metalloid). Chromium, copper, zinc, and nickel are reduced in a range of neutral to

slightly basic pH, while the solubility and mobility can increase in either very acidic

or very basic pH solutions. Arsenic is a metalloid that is more toxic at its reduced

state (As3+

) and Cr is more toxic at its oxidized state (Cr6+

). Providing reduced

alkaline environments or complexing materials forcing toxic metals to precipitate or

bind (Jakobsson et al., 1998). Adsorption is the process when a solved substance

binds onto a surface and is most governed by pH. In solution most metals present as

cations are adsorb more strongly as pH increases. In reduced anoxic environments

metal ions precipitate as sulfides, high pH means elevated concentrations of

hydroxide ions and metals precipitate as hydroxides or carbonates. Iron is an

abundant substance in soil and Cr (III) precipitates as hydroxides preferably with

iron in pH > 5, arsenate binds strongly with iron in soil. Humic matter adsorbs

cations at low pH e.g pH < 6 and is therefore an important source of cations, at

11

higher pH (> 6) cation adsorption by humic matter decreases as their solubility

increase with pH. Bioavailability of metals is lowered and extraction is possible

(Berggren et al., 2006). Inorganic compounds such as cyanides, fluorides and ozone

are remediated by decomposition, reduction and buffering (Jakobsson et al., 1998).

2.3 Heavy Metals

Heavy metal is a generic term used for metals and semimetals (metalloids) that are

associated with contamination. Their atomic densities are usually greater than 6

g/cm3 (Alloway, 1995; Wild, 1993; Van der Perk, 2006). Examples of such metals

include copper, lead, mercury, zinc, chromium, nickel, arsenic, tin, silver, and

cadmium. Heavy metals are also referred to as trace metals due to their relatively low

natural concentrations in soils, sediments, water, and organisms. In most

environments, heavy metals occur in their cationic forms, though some occur as

oxyanions for example, arsenate (AsO4 3-

) (Van der Perk, 2006). Whereas some

heavy metals such as Cd and Zn are less strongly adsorbed to soils and sediments.

Others such as Pb and Cu have been found to adsorb strongly and are released into

solution slowly when the ambient conditions are favourable. Unlike other

compounds, heavy metals are not biodegradable and many of them are toxic,

mutagenic, and carcinogenic. As a consequence, they accumulate in sediments and

pose a great threat to the environment especially when they encounter conditions that

increase their solubility and, when their concentrations in soils, sediments, water and

organisms exceed their acceptance levels (MacCarthy et al,. 1991; Volesky, 1994;

Clement et al., 1995; Volesky & Holan, 1995; Bozkurt et al., 2000). While present in

sediments, heavy metals could be occluded in amorphous materials; adsorbed on clay

surfaces or iron/manganese oxyhydroxides; precipitated as sulphides or oxides; or

complexed with organic matter (OM) (Tessier et al., 1979).

The releases of heavy metals to the environment started increasing

tremendously from the mid-19th century when industrialisation began. From this

period, enormous amount of heavy metals of deleterious effects have constantly been

released to the environment. These releases have occurred via several pathways such

as air, water and soil. Emissions via air are of enormous concern due to the large

12

quantities involved, the widespread dispersion, and the potential for extensive human

exposure (Jarup, 2003).

2.4 Sources of Heavy Metal Pollution

Pollution of sediments with heavy metals can occur in several ways. However, these

sources have been put into two main categories- natural and anthropogenic sources.

Anthropogenic inputs of heavy metals to the environment by far, exceed the natural

inputs.

2.4.1 Natural sources

Naturally, pollution of sediments with heavy metals occurs through weathering of

rocks. Weathering of rocks can occur through processes such as hydrolysis and

hydration reactions; oxidation and reduction reaction; dissolution and dissociation of

minerals; immobilization by precipitation; loss of mineral components via leaching

and volatilization; and chemical exchange processes such as cation exchange. Heavy

metals occur naturally in rocks as constituents. Through natural geological

weathering by any of the processes aforementioned, heavy metals can be released

into the environment. The concentrations of heavy metals due to natural geological

weathering are often referred to as background concentrations. Background

concentrations are not necessarily a threat to the environment but are considered so

only when their amounts exceed the acceptable limits in the environment. They could

serve as point source pollution or they may be transported to other places via surface

runoff or erosion, causing diffused pollution (Van der Perk, 2006).

2.4.2 Anthropogenic sources

It is no longer a matter of argument in many scientific debates that anthropogenic

activities are the main reasons for the observed increases of heavy metal

concentrations in sediments worldwide. Potential anthropogenic sources of heavy

metals worldwide include sewage sludge, application of fertilizers both of organic

and inorganic origins, leaching from building materials, industrial discharges and

13

disposals and atmospheric fallout (from smelting or from burning coal and gasoline).

Heavy metals released from anthropogenic activities are usually unstable and more

soluble and available than their natural forms (Dudley et al., 1991; Alloway, 1995;

Andersen et al, 1996; Van der Perk, 2006).

Anthropogenic releases of heavy metals to the environment increased greatly

in the 19th

century. This raised enormous concern worldwide for the adoption of

measures that would reduce their concentrations in the environment. With the

implementation of environmental regulations and improvement in technology, there

has been substantial reduction in the releases of heavy metals to the environment for

the past three decades (Van der Perk, 2006).

2.5 Heavy Metal Effects

Since heavy metals are not biodegradable and their high values can cause serious

problems like cancer for living organisms, they are a matter of high concern by the

environmental authorities. Among them, lead (Pb), cadmium (Cd) and mercury (Hg)

are of higher toxicity (Manahan, 2003). Long term exposure to toxic heavy metals,

can cause liver damage, lung disease, fragile bones and blood problems (Weiner,

2000). Heavy metals can bio accumulate in animals, fish, plants and humans

(Harrison & Mora, 1996), besides, as reported by Weiner (2010), some heavy metals

seem to be the major reason behind some specific cancers. Cancer cases that can be

attributed to environmental causes probably account for more than 60% of all

cancers, although the environment in this level, not only involves air, soil, sediments

and water, but also has food, drink, living habits, drugs and occupational exposure in

its domain (Zakrzewski, 2002). Some of the metals investigated in this thesis include

lead, chromium and copper.

2.5.1 Lead

Lead (Pb) is one of the most common contaminants found in soils and sediments as a

result of agricultural activities, urban activities and industrial activities such as

mining and smelting. It is toxic both to humans and animals and hence presents a

serious environmental and health hazard (Ma et al, 1995). The primary industrial

14

sources of lead (Pb) contamination include metal smelting and processing, secondary

metals production, lead battery manufacturing, pigment and chemical manufacturing,

and lead-contaminated wastes. Lead released to groundwater, surface water and land

is usually in the form of elemental lead, lead oxides and hydroxides, and lead metal

oxyanion complexes (Long et al., 1995). Some decades after the World War II, lead

became ubiquitous in sediments. This was due to the high usage of Pb-alkyls as

gasoline for automobiles. However, with the switch from using leaded to unleaded

automobile gasoline in the past few decades, studies have revealed that there has

been a substantial decline in the concentrations of lead in sediments (Bruland et al.,

1974; Barbeau et al., 1981; Gobeil & Silverberg, 1989). Lead usually occurs in

moderate amounts in the Earth‘s crust in the form of lead sulphide.

In natural solutions, lead reacts to form lead hydroxide, carbonate, and

phosphate. These compounds are less soluble reducing the mobility of lead greatly in

natural waters. In oxygenated seawater, dissolved inorganic lead carbonate is

predominant whereas in anoxic conditions, the sulphides of lead predominate

(Stumm & Morgan, 1981; Emerson et al., 1983; Van der Perk, 2006). The

concentrations of lead in natural waters is often low because lead sorbs strongly to

mineral and organic materials and it is also able to form complexes with manganese

oxide which can be precipitated in solution. Lead is an amphoteric metal in that its

hydroxides can be soluble in natural waters at high or low pH (Van der Perk, 2006).

Lead is one out of four metals that have the most damaging effects on human health.

The major health impacts of the lead include anaemia, rise in blood pressure, brain

damage, miscarriages, CNS, kidney and sperm damage. It is toxic to humans, and

especially hazardous to infants and children. Lead enters the body by inhalation,

ingestion or by skin contact. Lead can accumulate in the body over time causing

fatigue, headaches, vomiting, and seizures. Lead can have detrimental effects upon

hemoglobin production and kidney function (Bradl, 2004).

2.5.2 Chromium

Chromium is one of the less common elements and does not occur naturally in

elemental form, but only in compounds. Chromium is mined as a primary ore

product in the form of the mineral chromite, FeCr2O4. Major sources of Cr

15

contamination include releases from electroplating processes and the disposal of

chromium containing wastes (Long et al., 1995).

The wide spread use of chromium has resulted in the contamination of soils

and water (Wang & Vipulanandan, 2001). Chromium contamination is of great

concern due to its toxic, mutagenic, and carcinogenic nature. The chromium is

generated from steel and other alloy‘s production, chrome plating, pigments, and

leather tanning industries. Among the various forms of chromium, Cr (VI) is the

form which is commonly found at contaminated sites and most important one

because of its toxicity, solubility, and mobility characteristics (Katz & Salem, 1994).

It can cause skin rashes, nose irritation, stomach upsets, respiratory problems,

weakened immune systems, kidney and liver damage, lung cancer and sometimes

even death.

Chromium (Cr) is one of the most common skin sensitizers and often causes

skin sensitizing effect in the general public. A possible source of chromium exposure

is waste dumps for chromate-producing plants causing local air or water pollution.

Penetration of the skin will cause painless erosive ulceration (―chrome holes‖) with

delayed healing. These commonly occur on the fingers, knuckles, and forearms. The

characteristic chrome sore begins as a papule, forming an ulcer with raised hard

edges. Ulcers can penetrate deep into soft tissue or become the sites of secondary

infection, but are not known to lead to malignancy (Geller, 2001; Lewis, 2004).

Besides the lungs and intestinal tract, the liver and kidney are often target organs for

chromate toxicity (Rom, 2007). In natural waters, exposure to Cr has demonstrated

cumulative deleterious effects on fishes as a function of time (Velma et al., 2009,

Steinhagen et al., 2004). Chromium in its hexavalent oxidation state, which includes

chromates or dichromates, is widely recognised as potentially carcinogenic and

highly soluble in aqueous media (Huggins & Huffman, 2004) whereas trivalent Cr

(III) is less soluble and of much less concern to human health.

2.5.3 Copper

Copper in the aquatic environment is usually related to anthropogenic sources rather

than natural sources. Its industrial sources include mining, electroplating, petroleum

refining, metal works, and foundries, it‘s widely used in the manufacture of textiles,

16

electrical conductors and cooking utensils. Copper is the active ingredient in some

pesticides applied to reservoirs to inhibit fungal growth. Naturally, copper occurs in

the Earth‘s crust as free metal or in the +1 or +2 oxidation states. In oxidized

environments (e.g. oxygenated seawater) copper may occur as either Cu2+

or Cu1+

(cuprous) nevertheless, Cu1+

has a high tendency of undergoing disproportionate

reactions (Eq. 2.1) which may result in Cu2+

(cupric) predominating in oxic solution.

2 Cu Cu0 + Cu

2+ (2.1)

The Cu2+

formed, however, can be reduced again to Cu0 and Cu

1+ when reduced

conditions begin to prevail (Jacobs & Emerson, 1982). When present in sediments,

Cu can be adsorbed on surface of metal oxides, clay minerals, humic substances or

organo-mineral complexes; or be occluded in structures of secondary minerals or in

amorphous iron and manganese oxides. They could also be associated with

authigenic sulphides (Huang, 1993).

Copper sulfate is used as a drying agent in the anhydrous form, as an additive

for fertilizers and foods, and several industrial applications such as textiles, leather,

wood, batteries, ink, petroleum, paint, and metal, among others. It is used also as an

animal nutritional supplement. Copper sulfate is used as a fungicide, algaecide, root

killer, and herbicide in both agriculture and non-agricultural settings.

Copper ions are susceptible to complexation, especially with hydroxide and

carbonate ligands. In aerated natural waters containing dissolved carbonates, Cu2+

can react with the carbonates to form a strong CuCO3 (aq), which is usually the main

form of inorganic dissolved Cu though CuOH+ and Cu(CO3)2 can also be present

based on thermodynamic calculations (Stumm & Morgan, 1981). However, when the

pH under such conditions is above neutral, Cu(OH)3- complexes are formed. These

complexes are slightly soluble and can reduce the copper concentrations in water to

below 10μg/l. In the case where there is adsorption of Cu to sediments and soil

minerals or coprecipitated with ferric oxyhydroxide, the Cu concentration can even

decrease further. In the presence of sulphates under reduced conditions, Cu can react

to form strong insoluble sulphides (Jacob & Emerson, 1982; Van der Perk, 2006).

17

2.6 Remediation techniques

Remediation techniques are aim to extract, stabilise or concentrate metals due to the

indestructibility of metals. Metals binds relatively fast to solid particles and often

needs to be mobilized in order to extract them (NTV, 1993). Most metals are less

mobile in alkaline oxidized environments and therefore stabilization of metals often

means adjusting pH by different kinds of modifications (cement, rice husk ash, lime,

fly ash) to slightly basic. Hence metals are immobilized but still present (Kumpiene

et al., 2007)

The first step in the selection of remediation process is by characterizing the

site and sediment; the data obtained would allow coming to a decision whether the

sediment is contaminated and whether it poses a potential threat to human health or

the environment. If the sediment does not pose a threat, then no action is required. If

the sediment is contaminated and does pose a threat to human health or the

environment, then some action is required.

According to the USEPA 2001, the term treatment correspond to all

operation or operations that modify the composition of a dangerous substance or

contaminant by physico-chemical, thermal or biological actions in order to reduce

the toxicity, mobility or volume of the contaminated material. Remediation

technologies can be classified based on the place where the activity is being carried

out (INE, 2007). These are In-situ and Ex- situ. The first term is related to the

remediation activity that is being done in the same place or site which is polluted,

without the need of removing or excavating. The second one is about the activity

which needs excavating, dredging, removing or extracting the contaminated sediment

to be treated either on-site or off-site. Each type of technology has its own

disadvantages, it is certain that In-situ technology allows the treatment of the

contaminated soil without having to excavate or to remove, but it is also true that the

treatment requires greater time and represents bigger difficulties of verification in

the effectiveness of the treatment. While Ex-situ technology happens to be just the

opposite, it requires little treatment time, excavation and extraction of the

contaminated soil to carry out the treatment (Volke Sepúlveda et al., 2002).

In relation to the physico-chemical treatments, according to the EPA diverse

types of technologies can be found to apply in dependence of the polluting agent that

needs to be removed. The different technologies are discussed briefly.

18

2.6.1 Electrochemical remediation

Electrochemical remediation involves applying a low direct current (DC) or a low

potential gradient to electrodes that are inserted into the sediment/soil and encompass

the contaminated zone (Virkutyte et al., 2002). When DC electric fields are applied

to the contaminated sediment, migration of charged ions occurs. Positive ions are

attracted to the negatively charged cathode, and negative ions move to the positively

charged anode. Generally, in electrochemical remediation process, the development

of an acidic front is often couple with a successful remediation (Nystroem et al.,

2006). However, this acid addition also has some evident drawbacks. Achieving

these acidic conditions might be difficult due to higher sediment buffering capacity;

in addition, acidification of dredged sediment may not be an environmentally

acceptable method.

2.6.2 Soil Washing

Soil washing is a relatively simple and useful ex situ remediation technology that

involve the use of adding washing water, heavy metal can be transferred from the

sediment to wash solution. To enhance the performance of sediment washing,

various additives are employed, such as acid washing (e.g. H2SO4 and HNO3),

chelating agents (e.g. EDTA, DTPA and EDDS) or surfactants (e.g. rhamnolipid).

These additives can assist in the solubilisation, dispersal and desorption of metal

from dredged sediments. This technology is most appropriate for the weaker bound

metals in the form of exchangeable hydroxides; carbonates and reducible oxides

fraction, the most difficult ones to remove are not affected during the washing

process (Mulligan et al., 2001, Ortega et al., 2008). Additionally, fine grain

sediments are difficult to decontaminate through washing solutions, therefore

washing is most applicable to sands and gravels. Also when the waste mixture is

complex (e.g., metals with organics) formulating the washing fluid becomes difficult.

19

2.6.3 Chemical oxidation

Chemical oxidation is a remediation technique which makes use of adding chemical

substances that can oxidize organic contaminants present in soil and turn them into

carbon dioxide or into compounds that can be easily degraded. According to

Amarante (2000), there are two ways to inject the chemical substance, the first one is

near to the contaminated zone and with the groundwater extraction too or just

injecting with no extraction. There are some oxidants that are more frequently used;

just to mention some, ozone, hydrogen, permanganate, per sulphate, among some

others. The radical per sulphate is the most use just because it is more stable than the

other ones under different conditions, it reacts faster, etc. (Hamberg, 2009). This

technique has some limitations and it directly depends on some factors like the

amount or level of contamination, the organic matter content and how the particles

are distributed (Andreottola et al., 2009).

2.6.4 Phytoremediation

Phytoremediation is the use of plants to remove, impound, or detoxify pollutants.

This technology is widely viewed as an ecologically responsible alternative to the

environmentally destructive chemical remediation methods currently practiced

(Meagher, 2000). This technology is popularly applied in soil remediation, and also

shows some excellent remediation effects in some shallow rivers, lakes and wetlands.

Phytoremediation is comprised of two tiers, one by plants themselves and the other

by the root colonizing microbes, which degrades the toxic compounds to further non-

toxic metabolites. Generally, hydrophytes have the ability to uptake and accumulate

a variety of heavy metals by the action of phytochelatins and metallothioneins

(Suresh & Ravishankar, 2004). However, mass balances experiments prove that

metal uptake by hydrophytes were not high enough for phytoextraction. This

indicates that in hydro remediation, the direct uptake of hydrophytes is small, and the

indirect reactions, such as stimulation of microbial activity, redox

reactions/formation and precipitation of insoluble metal compounds in the

rhizosphere may play a relative important role (Clemente et al., 2005).

20

2.7 Solidification/Stabilisation as a remediation technology for the treatment

of contaminated sediments

Solidification and stabilization (S/S) is a technology whereby waste materials are

treated in a manner that alters the physico-chemical properties of the contaminants.

This reduces their spread via leaching thereby minimizing the threat they pose to the

environment. The process may involve chemical bonding or physical entrapment of

the hazardous compounds. Most applications of S/S either utilize Ordinary Portland

Cement as the sole binder or may be combined with other materials such as lime, fly

ash, blast furnace slag etc (Batchelor, 2006). Solidification/stabilization of

contaminants may be done in-situ or ex situ. The in-situ methods are usually

accomplished by injecting the binding agent into the contaminated site without

excavating the waste material. This method does not expose the waste material to

environment, thus pollution of other areas is greatly reduced. The ex-situ method on

the hand involves excavation of the contaminated material and mixing it with the

cement-based agent. The mixed material may then be returned to the ground at the

site of excavation or placed in a landfill. The area may thereafter be covered with

clean soil or pavement. Without proper handling, the ex-situ method of treating the

waste material can lead to contamination of other areas.

However, compared to the in-situ method, the ex-situ method can result in

proper mixing of the binder with the waste material that can greatly reduce leaching

of pollutants, which is the main objective of contaminant stabilization and

solidification (Barth & Wiles, 1989). Conner (1990) investigated the chemical

fixation and solidification of hazardous waste and identified high-unconfined

compressive strength, low permeability, and less interconnected pores in the

stabilized material as very important parameters for the success of the S/S. Batchelor

(2006) later supported this when he did a general review on stabilization and

solidification of waste materials.

2.7.1 Solidification

Solidification as defined by the Environmental Protection Agency (EPA) is a

technique that is employed to encapsulate the waste into a monolithic solid of high

21

structural integrity without any necessary chemical interactions between the

hazardous chemicals and the solidifying reagent (Conner, 1990; Glasser, 1997; Poon

et al., 2004). Solidification prevents the hazardous chemical from spreading in the

environment in that, it results in reduced surface area and low permeability of the

monolith. Advective flow through the waste material is greatly reduced (Cullinane &

Jones, 1986; Batchelor, 2006) thereby reducing contact between the hazardous waste

and other external substances that would otherwise enhance contaminant mobility.

Generally, permeabilities in the range of 10-5

to 10-9

cm/sec for cement-based

matrices and 10-6

- 10-7

cm/sec for pozzolanic-based waste forms have been

documented (Arniella & Blythe, 1990). Malviya & Chaudhary (2006) have also

proposed a minimum unconfined compressive strength of 50 psi (345 kPa) which is

the minimum strength to support any overburden pressure such as from vehicles. The

minimum allowable compressive strength of solidified samples must be greater than

50psi (345kPa) as a measure of adequate bonding level in a solidified sample.

2.7.2 Stabilisation

Stabilization is a process that leads to a reduction of the hazard potential of a waste

by converting the contaminants into their least soluble, mobile or toxic forms by

changing the chemical nature of the contaminants. The components of the binding

material react with the contaminants, culminating in changes in the chemical

reactivity of the contaminants. The changes in the chemical properties of the

hazardous substance is however dependent on the binding agent used and the result

of the chemical interaction between the binder and the waste form (Conner, 1990;

Glasser, 1997; Poon et al., 2004).

2.8 Common Binders Used for the S/S Treatment Technology

The use of a binder when treating a waste by stabilisation/solidification technique

has an advantage because of their capacity to provide physical solidity to the treated

product and chemical stabilisation. Suitable binders are selected for contaminants

and site specific based on a recognised design criterion. The high pH induced by the

addition of the most common binders, such as lime and Portland cement, results in

22

the precipitation of many contaminant species and a corresponding reduction in

mobility (Stabilisation). Secondly, the ability of the binder to set into a solid mass

encapsulating the contaminant results in a physical immobilization process

(Solidification). Many binders rely on the presence of free CaO for this, although the

use of additives is common for modification of the hydration/setting processes. The

design criteria have usually been depended on the properties of the end products and

required taking into account the nature of the material and contaminants that are

being treated.

Primary stabilizing agents are widely applied in the remediation of

contaminated soils/sediments worldwide. They are categorized as the ones that can

be used alone to carry on the stabilising action required. When a cementitious binder

is used the waste or sediment particle is encapsulate chemically and physically. The

most common binders used in S/S technology are: Portland cement, lime and

thermoplastic materials that include bitumen and sulphur polymer cement.

Secondary stabilizing agents are siliceous and aluminous materials which in

itself possess little or no cementitious value, but will, in finely divided form and in

the presence of moisture, chemically react with calcium hydroxide at ordinary

temperature to form compounds possessing cementitious properties (ASTM 595).

Sometimes only a small proportion of cement or lime is needed as an activator and

the secondary agent may comprise the major proportion of the binder. Secondary

materials may be added on S/S system for particular contaminants in quantities that

provide an economic binder system, without compromising technical properties.

Some of the most common secondary stabilizing agents are: Rice husk ash,

Pulverised fuel ash (PFA), Ground granulated blast furnace slag (ggbs), silica fume

etc (Bone et al., 2004). The common binders and additives used in the S/S treatment

technology are shown in Table 2.1.

23

Table 2.1 Common binders and additives used in S/S treatment Technology

No Binders/Additives Examples

1 Activated carbon

Powdered activated carbon (PAC),

Granular activated carbon (GAC)

2 Concrete additives

Accelerators, retarders, air trapping

agent, forming gas agent etc pozzolana

materials

3 Carbonates

Soda ash (sodium carbonate), calcium

carbonate, magnesium carbonate.

4 Ferrum and aluminium compounds Sodium aluminate (accelerator), fine

powder aluminum etc.

5 Neutralising agent

Lime, soda ash, flyash, sodium

hydroxide, magnesium hydroxide

6 Reducing agent

Blastfurnance slag, sodium borohydride,

light dehumidifiers,ferrous sulphate etc.

7 Organophyllic clays

Selected clays such as bentonite,

montmorillonite, kaolinite etc

8 Surface active agents (surfactants) Emulsifier, soap sulfonat, sodium

dodecylsulfate (SDS), etc

9 Sulfide (organic and inorganic)

Ferrous sulphate, sodium meta bisulfite,

sodium hydrosulfite, sulfite etc.

10 Micro silica Silica fume

11 Dissolved silica Sodium silicate, potassium silicate etc

12 Phosphate Calcium phosphate (Apatite)

2.8.1 Cement

Portland cement is the most commonly used binder for this process due to its cost

effectiveness, availability and compatibility with a variety of wastes (Spence et al.,

USEPA 1999, 2001, 2004). High pH of this binder is effective in immobilizing many

toxic metals, by precipitation and sorption reaction. In most cases, a solid is

produced with sufficient strength to support itself and a landfill cover.

Cement stabilisation has been widely used in order to improve the

engineering properties of soils (Broms 1999, Feng et al., 2001, Lorenzo & Bergado

2004, Xiao & Lee 2008). Cement is a hydraulic type stabilising agent. According to

Bergado et al., (1996), there are two major chemical reactions which are induced by

the addition of cement to clay and govern the soil cement stabilisation: The primary

hydration reaction of the cement and water, and the secondary pozzolanic reactions

between the limes released cementation agent and the clay minerals. The primary

hydration products are hydrated calcium silicates (C2SHx, C3S2Hx), hydrated calcium

aluminates (C3AHx, C4AHx) and hydrated lime Ca (OH)2. The first two of hydration

products listed above are the main cementitious products formed and the hydrated

lime is deposited as a separate crystalline solid phase. These cement particles bind

24

the adjacent cement grains together during hardening and form a hardened skeleton

matrix, which encloses unaltered soil particles.

Cement is known to create bonding in concrete through hydration. According

to Wilk & Germano (2001), Portland cement-based mix designs have been popular

S/S treatments and have been applied to a greater variety of contaminated materials

than any other binding agent (Conner, 1990). Cement is frequently selected for the

reagents ability to (1) chemically bind free liquids, (2) reduce the permeability of the

waste form, (3) encapsulate waste particles surrounding them with an impermeable

coating, (4) chemically fix hazardous constituents by reducing their solubility, and

(5) facilitate the reduction of the toxicity of some contaminants.

The above results are achieved due to the various kinds of physical and

chemical changes that happen when cement reacts with water. Portland cement when

mixed with water under goes hydration and forms physical bonds with contaminants.

This traps the contaminants into a matrix and eventually makes them less mobile.

Such techniques can also lower the permeability of the treated material and

significantly reduce the leachability. Hydration of cement on contact with water

generates calcium hydroxide Ca(OH)2 which further reacts with water and

contaminants, making them somewhat less soluble and therefore less leachable.

Portland cement on hydration also forms compounds of carbonates and silicates

which tend to bind with the contaminants making them less soluble.

2.8.2 Rice husk ash as an additive in Solidification/Stabilisation

Rice husk is the outer covering of the grain of rice plant with a high concentration of

silica, generally more than 80-85% (Siddique, 2007). It is responsible for

approximately 30% of the gross weight of a rice kernel and normally contains 80%

of organic and 20% of inorganic substances. Rice husk is produced in millions of

tons per year as a waste material in agricultural and industrial processes. After

burning rice husk, the RHA produced as a by-product, about 20% of its original

weight (Anwar et al., 2001, Chindaprasirt & Jaturapitakkal, 2009). RHA is a highly

pozzolanic material (Tashima et al., 2004). The non-crystalline silica and high

specific surface area of the RHA are responsible for its high pozzolanic reactivity.

RHA has been used in lime pozzolana mixes and could be a suitable partly

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