Post on 29-Mar-2019
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
4.7.1 Relationship between strength and density of
chromium spiked samples 79
4.8 UCS of Stabilised/Solidified Copper spiked
sediment samples 82
4.8.1 Relationship between strength and density of
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
4.10.2.1 Relationship between strength and leachability of
chromium spiked sediment sample 97
4.10.3 Leachability of Copper spiked sediment 100
4.10.3.1 Relationship between strength and leachability of
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
throughout 28 days of curing 79
4.11 Relationship between strength and density of chromium
spiked sample 80
4.12 UCS of solidified/stabilised Copper spiked samples
throughout 28 days of curing 83
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
C.3 Qualitative analysis result of heavy metals in the
sediment using ICP-MS sample C 164
C.4 Qualitative analysis result of heavy metals in the
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
C.7 Quantitative analysis of some selected heavy metals
in sample C using ICP-MS 170
C.8 Quantitative analysis of some selected heavy metals
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
sediment for 7, 14 & 28 days 177
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
sediment for 7, 14 & 28 days 180
E.7 pH of the DIW leachates for Copper spiked
sediment for 7, 14 & 28 days 181
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
sediment for 7, 14 & 28 days 183
F.1 Relationship between leachability with pH of lead at
7 days in the TCLP 184
F.2 Relationship between leachability with pH of lead at
14 days in the TCLP 184
F.3 Relationship between leachability with pH of lead at
28 days in the TCLP 184
F.4 Relationship between leachability with pH of lead at
7 days in the SPLP 185
F.5 Relationship between leachability with pH of lead at
14 days in the SPLP 185
F.6 Relationship between leachability with pH of lead at
28 days in the SPLP 185
F.7 Relationship between leachability with pH of lead at
7 days in the DIW 186
F.8 Relationship between leachability with pH of lead at
14 days in the DIW 186
F.9 Relationship between leachability with pH of lead at
28 days in the DIW 186
F.10 Relationship between leachability with pH of
chromium at 7 days in the TCLP 187
F.11 Relationship between leachability with pH of
chromium at 14 days in the TCLP 187
F.12 Relationship between leachability with pH of
chromium at 28 days in the TCLP 187
F.13 Relationship between leachability with pH of
chromium at 7 days in the SPLP 188
F.14 Relationship between leachability with pH of
chromium at 14 days in the SPLP 188
F.15 Relationship between leachability with pH of
chromium at 28 days in the SPLP 188
xvii
F.16 Relationship between leachability with pH of
chromium at 7 days in the DIW 189
F.17 Relationship between leachability with pH of
chromium at 14 days in the DIW 189
F.18 Relationship between leachability with pH of
chromium at 28 days in the DIW 189
F.19 Relationship between leachability with pH of copper
at 7 days in the TCLP 190
F.20 Relationship between leachability with pH of copper
at 14 days in the TCLP 190
F.21 Relationship between leachability with pH of copper
at 28 days in the TCLP 190
F.22 Relationship between leachability with pH of copper
at 7 days in the SPLP 191
F.23 Relationship between leachability with pH of copper
at 14 days in the SPLP 191
F.24 Relationship between leachability with pH of copper
at 28 days in the SPLP 191
F.25 Relationship between leachability with pH of copper
at 7 days in the DIW 192
F.26 Relationship between leachability with pH of copper
at 14 days in the DIW 192
F.27 Relationship between leachability with pH of copper
at 28 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
K.2 Relationship of strength and leachability of lead in
the SPLP at 7, 14 and 28 days 239
K.3 Relationship of strength and leachability of lead in
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
K.5 Relationship of strength and leachability of Copper
in the SPLP at 7, 14 and 28 days 240
K.6 Relationship of strength and leachability of Copper
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
K.8 Relationship of strength and leachability of
Chromium in the SPLP at 7, 14 and 28 days 241
K.9 Relationship of strength and leachability of
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
4.4 Relationship between compressive strength and density of
lead spiked sample at 14 days 77
4.5 Relationship between compressive strength and density of
lead spiked sample at 28 days 78
4.6 UCS developments of chromium spiked sediment samples
throughout 28 days curing period 79
4.7 Relationship between compressive strength and density of
Chromium spiked sample at 7 days 81
4.8 Relationship between compressive strength and density of
Chromium spiked sample at 14 days 81
4.9 Relationship between compressive strength and density of
Chromium spiked sample at 28 days 82
4.10 UCS developments of Copper spiked sediment samples
throughout 28 days curing period 83
4.11 Correlation between compressive strength and density of
Copper spiked sample at 7 day 85
4.12 Relationship between compressive strength and density of
Copper spiked sample at 14 days 85
4.13 Relationship between compressive strength and density of
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
stabilized samples at 7 days 89
4.16 The effect of curing period on strength for Pb, Cr and Cu
stabilized samples at 7 days 89
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
4.27 Relationship between strength and leachability of copper in
the SPLP at 7, 14 and 28 days 104
4.28 Relationship between strength and leachability of copper in
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
4.36 The effect of pH on the leachability of lead in the DIW at 14
days 111
xxii
4.37 The effect of pH on the leachability of lead in the DIW at 28
days 111
4.38 The effect of pH on the leachability of Copper in the TCLP at
7 days 112
4.39 The effect of pH on the leachability of Copper in the TCLP at
14 days 112
4.40 The effect of pH on the leachability of Copper in the TCLP at
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
4.48 The effect of pH on the leachability of Chromium in the
TCLP at 14 days 115
4.49 The effect of pH on the leachability of Chromium in the
TCLP at 28 days 115
4.50 The effect of pH on the leachability of Chromium in the
SPLP at 7 days 116
4.51 The effect of pH on the leachability of Chromium in the
SPLP at 14 days 116
4.52 The effect of pH on the leachability of Chromium in the
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
leaching tests (TCLP, SPLP & DIW) 208
I Concentration of chromium at 7, 14 and 28 days after
leaching tests (TCLP, SPLP & DIW) 223
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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