NOVEL TREATMENT AND VALUE-ADDED UTILIZATION OF OILY SLUDGE
USING IONIC LIQUID AND PYROLYSIS
by
Yuan Tian
B.E., Anhui University of Science and Technology, 2010
M.E., Beijing Institute of Technology, 2013

THESIS SUBMITTED IN PARTIAL FULFILLMENT OF
THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY
IN
NATURAL RESOURCES AND ENVIRONMENTAL STUDIES

UNIVERSITY OF NORTHERN BRITISH COLUMBIA
October 2019
© Yuan Tian, 2019

ABSTRACT
The effective treatment of oily sludge has been a challenging problem faced by the
petroleum industry worldwide. It is a semi-solid mixture of hydrocarbons, water, metallic ions,
and suspended fine solids. The recalcitrant nature of oily sludge makes the treatment a difficult
and costly task. The objective of this dissertation research was to develop environmentally
friendly and economically competitive techniques for oily sludge treatment. Three different
approaches were developed: ionic liquid (IL)-enhanced solvent extraction, co-pyrolysis with
wood waste, and value-added utilization of oily sludge as sorbent to remove lead (Pb2+) and
cadmium (Cd2+) from solution. Firstly, as compared to conventional solvent extraction, the ILenhanced solvent extraction not only improved oil recovery efficiency but also greatly reduced
solvent and energy consumption as well as shortening the treatment duration even at low IL
concentration. Secondly, co-pyrolysis of oily waste and hog fuel was conducted in a fixed bed
reactor. Three experimental parameters (pyrolysis temperature, reaction time, and hog fuel
addition) were explored to optimize both oil recovery and metal ion immobilization. The latter
was tested through sequential extraction techniques with high temperature pyrolysis leading to
metal ions fixed within the residues. The addition of hog fuel had a significant synergistic
effect on the distribution of metal ions in the various extraction fractions resulting in lower risk
index (RI) values. Thirdly, the oily sludge-derived char (OS500) obtained at 500 °C could
effectively remove Pb2+ from solution with the maximum sorption capacity of 373.2 mg/g
(based on a Langmuir model). Sorption of Pb2+ by the OS500 was mainly attributed to its
precipitation with carbonate (CO32‒) originating in OS500. The maximum sorption capacity
for Cd2+, using a Langmuir model, was 23.19 mg/g. Complexation and metal ion exchange
dominated Cd2+ sorption on OS500. The Pb2+ sorption capacity dramatically decreased as the
ii

pyrolysis temperature increased from 500 to 900 °C due to the decomposition of minerals
which could release CO32‒ at high temperature. The activated OS500 showed a higher sorption
capacity (90.06 mg/g) for Cd2+ than OS500 (23.95 mg/g) because conversion of barite (BaSO4)
to witherite (BaCO3) after chemical activation favored the precipitation of Cd-carbonate.

iii

CO-AUTHORSHIP
For all the chapters in this dissertation, I was the principal investigator including design of
studies, acquisition of data, and analysis of data. I wrote the manuscripts and was responsible
for incorporating comments and feedbacks in the revised manuscripts. Dr. Jianbing Li
supervised the experiments and contributed to experimental design, data analysis, and revision
of the manuscripts. He is included in authorship on all resulting publications. Dr. Bill McGill
and Dr. Todd Whitcombe contributed to the design and implementation of experiments and
helped review and improve the manuscripts, so they were included in some of my publications.
Dr. Ron Thring and Dr. Xiuyi Yan provided useful suggestions in co-pyrolysis experiment and
contributed some comments on the manuscripts, so they were included in related publications.

Publication and authorships from this dissertation (Published or prepared for submission):
Tian, Y., McGill, W.B., Whitcombe, T.W., Li, J., 2019. Ionic liquid-enhanced solvent
extraction for oil recovery from oily sludge. Energy Fuels 33, 3429–3438. DOI:
10.1021/acs.energyfuels.9b00224. (Chapter 3)
Tian, Y., Li, J., Yan, X., Whitcombe, T., Thring, R., 2019. Co-pyrolysis of metal contaminated
oily waste for oil recovery and heavy metal immobilization. J. Hazard. Mater. 373, 1–10. DOI:
10.1016/j.jhazmat.2019.03.061. (Chapter 4)
Tian, Y., Li, J., Whitcombe, T., McGill, W.B. Thring, R., 2019. Application of oily sludgederived char for lead and cadmium removal from aqueous solution. Submitted to Chem. Eng.
J. on Aug 27, 2019. (Chapter 5)
Tian, Y., Li, J., McGill, W.B., Whitcombe, T., 2019. Sorption of lead and cadmium on oily
sludge-derived chars prepared at different pyrolysis temperatures and chemical activation.
Prepared for submission to J. Hazard. Mater. (Chapter 6)

iv

TABLE OF CONTENTS
ABSTRACT .............................................................................................................................. ii
CO-AUTHORSHIP ................................................................................................................. iv
TABLE OF CONTENTS ...........................................................................................................v
LIST OF TABLES ................................................................................................................. viii
LIST OF FIGURES ...................................................................................................................x
GLOSSARY .......................................................................................................................... xiii
ACKNOWLEDGEMENT ..................................................................................................... xvi
Chapter 1 GENERAL INTRODUCTION ...............................................................................1
1.1. Background .................................................................................................................... 1
1.2. Statement of the problem ............................................................................................... 2
1.3. Objectives of the study ................................................................................................... 5
1.4. Organization of the dissertation ..................................................................................... 6
Chapter 2 LITERATURE REVIEW ........................................................................................7
2.1. Sources of oily sludge .................................................................................................... 7
2.2. Characteristics of oily sludge ......................................................................................... 8
2.3. Toxicity of oily sludge ................................................................................................. 12
2.4. Techniques for treating oily sludge .............................................................................. 15
2.5. Ionic liquid-enhanced solvent extraction ..................................................................... 21
2.5.1. Solvent extraction .................................................................................................. 21
2.5.1.1. Selection of solvent ............................................................................................ 21
2.5.1.2. Parameters affecting extraction performance ..................................................... 23
2.5.2. Ionic liquid ............................................................................................................. 24
2.6. Pyrolysis ....................................................................................................................... 31
2.6.1. Pyrolysis in oily sludge treatment ......................................................................... 31
2.6.2. Effects of parameters on liquid production ........................................................... 32
2.6.3. Effects of parameters on gas production ............................................................... 34
2.6.4. Effects of parameters on solid production ............................................................. 36
2.6.5. Co-pyrolysis........................................................................................................... 40
2.7. Summary of literature review ....................................................................................... 43
Chapter 3 IONIC LIQUID-ENHANCED SOLVENT EXTRACTION FOR OIL
RECOVERY FROM OILY SLUDGE ....................................................................................46
v

Abstract ............................................................................................................................... 46
3.1. Introduction .................................................................................................................. 47
3.2. Materials and methods ................................................................................................. 49
3.2.1. Materials and reagents ........................................................................................... 49
3.2.2. Single factor experiments ...................................................................................... 50
3.2.3. Optimization of solvent extraction and IL-enhanced solvent extraction ............... 53
3.2.4. Laboratory analysis................................................................................................ 55
3.2.4.1. Hydrocarbon analysis using gas chromatograph ................................................ 55
3.2.4.2. Characterization of recovered oil and residual sand/clay ................................... 56
3.3. Results and discussion.................................................................................................. 57
3.3.1. Single factor experiments ...................................................................................... 57
3.3.2. Multifactor orthogonal experiments ...................................................................... 60
3.3.2.1. Solvent extraction alone ..................................................................................... 60
2.3.2.2. IL-enhanced solvent extraction .......................................................................... 64
2.3.3. Characteristics of the recovered oil and residual sand/clay ................................... 67
3.4. Summary................................................................................................................... 75
Chapter 4 CO-PYROLYSIS OF METAL CONTAMINATED OILY WASTE FOR OIL
RECOVERY AND HEAVY METAL IMMOBILIZATION..................................................76
Abstract ............................................................................................................................... 76
4.1. Introduction .................................................................................................................. 77
4.2. Materials and methods ................................................................................................. 79
4.2.1. Materials ................................................................................................................ 79
4.2.2. Experimental set up ............................................................................................... 80
4.2.3. Experimental design .............................................................................................. 81
4.2.4. Chemical analysis and heavy metal species characterization ................................ 85
4.2.5. Determination of potential ecological risk index (RI) ........................................... 88
4.3. Results and discussion.................................................................................................. 89
4.3.1. Response surface model and statistical analysis.................................................... 89
4.3.2. TPH recovery rate .................................................................................................. 91
4.3.3. Properties of solid residues .................................................................................... 96
4.3.4. Heavy metal speciation characteristics .................................................................. 98
4.3.5. Potential ecological risk assessment .................................................................... 103
4.4. Summary .................................................................................................................... 106
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Chapter 5 APPLICATION OF OILY SLUDGE-DERIVED CHAR FOR LEAD AND
CADMIUM REMOVAL FROM AQUEOUS SOLUTION..................................................107
Abstract ............................................................................................................................. 107
5.1. Introduction ................................................................................................................ 108
5.2. Materials and methods ............................................................................................... 110
5.2.1. Materials and reagents ......................................................................................... 110
5.2.2. Char preparation .................................................................................................. 111
5.2.3. Char characterization ........................................................................................... 112
5.2.4. Batch sorption studies .......................................................................................... 114
5.2.5. Post-sorption analysis .......................................................................................... 115
5.3. Results and discussion................................................................................................ 118
5.3.1. Characteristics of char ......................................................................................... 118
5.3.2. Sorption kinetics .................................................................................................. 122
5.3.3. Sorption isotherms ............................................................................................... 126
5.3.4. Post-sorption characteristics and sorption mechanisms ...................................... 134
5.3.4.1. XRD and FTIR results ...................................................................................... 134
5.3.4.2. Sequential extraction results ............................................................................. 137
5.4. Summary .................................................................................................................... 140
Chapter 6 SORPTION OF LEAD AND CADMIUM ON OILY SLUDGE-DERIVED
CHAR PREPARED AT DIFFERENT PYROLYSIS TEMPERATURES AND CHEMICAL
ACTIVATION .......................................................................................................................142
Abstract ............................................................................................................................. 142
6.1. Introduction ................................................................................................................ 143
6.2. Materials and methods ............................................................................................... 145
6.2.1. Materials and reagents ......................................................................................... 145
6.2.2. Chars preparation and characterization ............................................................... 146
6.2.3. Preliminary test .................................................................................................... 147
6.2.4. Post-sorption analysis .......................................................................................... 147
6.3. Results and discussion................................................................................................ 148
6.3.1. Characteristics of chars ........................................................................................ 148
6.3.2. XRD and FTIR results ......................................................................................... 155
6.3.3. Sequential extraction results ................................................................................ 160
6.4. Summary .................................................................................................................... 168
vii

Chapter 7 CONCLUSION AND RECOMMENDATIONS .................................................170
7.1. Synthesis and conclusion ........................................................................................... 170
7.2. Limitations and future research .................................................................................. 174
References ..............................................................................................................................177
APPENDIX A ........................................................................................................................208
APPENDIX B ........................................................................................................................214
APPENDIX C ........................................................................................................................219
APPENDIX D ........................................................................................................................223

LIST OF TABLES
Table 2.1. Heavy metal content in oily sludge generated from various refineries ................. 11
Table 2.2. Summary of technologies for the oily sludge treatment ....................................... 16
Table 2.3. Summary of ionic liquids used in bitumen recovery from oil sands..................... 29
Table 3.1. Characteristics of the oily sludge .......................................................................... 50
Table 3.2. Single-factor experimental conditions .................................................................. 51
Table 3.3. Orthogonal array L9 (34) experimental design and factor levels ........................... 54
Table 3.4. Results of the orthogonal experiments for solvent extraction alone ..................... 61
Table 3.5. Results of the orthogonal experiments for IL-enhanced solvent extraction ......... 65
Table 3.6. Properties of extracted oil and residual sand/clay after extraction in different
systems ........................................................................................................................ 74
Table 4.1. Soil texture and synthetic oily waste properties ................................................... 80
Table 4.2. Coded (𝑥𝑖) and un-coded (𝑋𝑖) Box-Behnken design matrix with the experimental
results .......................................................................................................................... 84
Table 4.3. ANOVA of the regression model for TPH recovery ............................................ 90
Table 4.4. ANOVA of the regression model for RI in solid residue...................................... 91
Table 4.5. Total C, H, N, and pH in the feedstock and the solid residue after pyrolysis ....... 97
Table 5.1. Properties of oily sludge..................................................................................... 111
Table 5.2. Physicochemical properties of OS500a .............................................................. 120
Table 5.3. The fitting results of Pb2+ and Cd2+ sorption on OS500 using the pseudo first
order kinetic (PF-order), pseudo second order kinetic (PS-order), Elovich, and
intraparticle diffusion (IPD) models. ........................................................................ 126
Table 5.4. Langmuir (L), Freundlich (F), Sips (S), Redlich-Peterson (RP), Temkin (T), and
Dubinin-Radushkevich (DR) isotherm parameters of Pb2+ and Cd2+ sorption on
OS500. ...................................................................................................................... 131
Table 5.5. Summary of the literature data on sorption of Pb2+ and Cd2+ to different biochars
or other sorbents ........................................................................................................ 132
Table 5.6. The release of K+, Ca2+, Mg2+, and Ba2+ during Pb2+ and Cd2+ sorption by OS500
................................................................................................................................... 139
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Table 6.1. Preliminary sorption results of chars.................................................................. 150
Table 6.2. Physicochemical properties of charsa ................................................................. 151
Table 6.3. The release of K+, Ca2+, Mg2+, and Ba2+ during Pb2+ and Cd2+ sorption by OSDCs
................................................................................................................................... 163
Table 6.4. Quantitative analysis of sorption mechanisms of Pb2+/Cd2+ onto OSDCs.......... 164
Table A1. Mass balance calculation of the F2, F3, and F4 fractions of TPH for orthogonal
experiments of solvent extraction alone ................................................................... 208
Table A2. Mass balance calculation of the F2, F3, and F4 fractions of TPH for orthogonal
experiments of IL-enhanced solvent extraction process ........................................... 210
Table A3. ANOVA for TPH, F2, F3, and F4 recovery in solvent extraction alone ............ 212
Table A4. ANOVA for TPH recovery in IL-enhanced solvent extraction process ............. 212
Table B1. Concentrations of extractable fractions of heavy metals in the oily waste (SRC)
and its solid residues ................................................................................................. 214
Table C1. Sequential extraction results of OS500-B ........................................................... 219
Table C2. Sequential extraction results of OS500-Pb and OS500-Cd................................. 219
Table C3. Recovery of sequential extraction results of OS500-Pb and OS500-Cd ............. 220
Table D1. Sequential extraction results of OSDCs before Pb2+/Cd2+sorption ..................... 223
Table D2. Sequential extraction results of Pb-loaded/Cd-loaded OSDCs ........................... 224
Table D3. Recovery of sequential extraction results of Pb-loaded/Cd-loaded OSDCs ....... 225
Table D4. Pearson correlation coefficient matrix between OSDCs properties and Pb2+/Cd2+
sorption mechanism .................................................................................................. 226

ix

LIST OF FIGURES
Figure 3.1. The phases formed by mixing, (a) oily sludge with [Emim][BF4] and
cyclohexane in the proportion of 1:3:2 (by weight), and (b) oily sludge with
cyclohexane in the proportion of 1:2 (by weight). ...................................................... 52
Figure 3.2. Influence of different operational parameters on TPH recovery: (a) So/Sl ratio
(shaking speed: 200 rpm, extraction duration: 30 min, IL/Sl ratio: 1:1 mL/g), (b)
shaking speed (So/Sl ratio: 2:1 mL/g, extraction duration: 30 min, IL/Sl ratio: 1:1
mL/g), and (c) extraction duration (So/Sl ratio: 2:1 mL/g, shaking speed: 200 rpm,
IL/Sl ratio: 1:1 mL/g). Significant (P < 0.05) differences are shown by different
letters: lowercase letter for the solvent extraction, and uppercase letter for the ILenhanced solvent extraction. ....................................................................................... 58
Figure 3.3. Influence of IL/sludge (IL/Sl) ratio on TPH recovery (So/Sl ratio: 2:1 mL/g,
shaking speed: 200 rpm, extraction duration: 30 min). Different lowercase letters (a,b)
indicates significant (P < 0.05) differences................................................................. 60
Figure 3.4. Effect of independent factors on (a) TPH recovery, (b) F2 recovery, (c) F3
recovery, and (d) F4 recovery of the solvent extraction alone process. ..................... 63
Figure 3.5. Effect of independent factors on TPH recovery of the IL-enhanced solvent
extraction process........................................................................................................ 65
Figure 3.6. GC chromatogram of petroleum hydrocarbons in (a) recovered oil using solvent
extraction alone, (b) recovered oil from top oil/solvent layer using IL-enhanced
solvent extraction, and (c) recovered oil from oil/solvent-IL interface using ILenhanced solvent extraction. ....................................................................................... 68
Figure 3.7. Distribution of PHCs fractions in crude oil, recovered oil (RO) using solvent
extraction alone, recovered oil in top oil/solvent layer using IL-enhanced solvent
extraction, and recovered oil in oil/solvent-IL interface using IL-enhanced solvent
extraction..................................................................................................................... 69
Figure 3.8. FTIR spectra of the (a) oily sludge, (b) recovered oil, and (c) residual sand/clay
for solvent extraction alone. ........................................................................................ 70
Figure 3.9. FTIR spectra of the (a) recovered oil from top oil/solvent layer, (b) recovered oil
from oil/solvent-IL interface, (c) residual sand/clay, and (d) [Emim][BF4] in ILenhanced solvent extraction. Residual sand/clay was treated with water and washed 3
times. ........................................................................................................................... 72
Figure 4.1. Schematic diagram of fixed-bed pyrolysis reactor (1 - N2 gas cylinder, 2 - flow
meter, 3 - thermal couple, 4 - quartz tube, 5 - sample ark, 6 - tube furnace, 7 temperature controller, 8 - ice-water bath, 9 - Tedlar gas bag)................................... 81
Figure 4.2. Gas chromatogram of (a) oil from untreated oily waste and (b) recovered oil after
pyrolysis treatment (run #11). ..................................................................................... 93
Figure 4.3. The statistical evaluation of the effect of single factor on TPH recovery rate, (a)
pyrolysis temperature (reaction time: 75 min, hog fuel ratio: 10%), (b) reaction time
(pyrolysis temperature: 500 °C, hog fuel ratio: 10%), and (c) hog fuel addition
(pyrolysis temperature: 500 °C, reaction time: 75 min). ............................................ 94

x

Figure 4.4. Contour plot of TPH recovery (%) as function of pyrolysis temperature and
reaction time at hog fuel addition = 10%. ................................................................... 96
Figure 4.5. Response surface graph of TPH recovery as function of (a) reaction time and hog
fuel addition at pyrolysis temperature = 500 °C , and (b) pyrolysis temperature and
hog fuel addition at reaction time = 75 min. ............................................................... 96
Figure 4.6. Fraction distribution of metals in oily waste before and after pyrolysis treatment.
................................................................................................................................... 100
Figure 4.7. Response surface graph of risk index (RI) as function of (a) pyrolysis
temperature and reaction time (at hog fuel addition = 10%), (b) pyrolysis temperature
and hog fuel addition (at reaction time = 75 min), and (c) reaction time and hog fuel
addition (at pyrolysis temperature = 500 °C)............................................................ 105
Figure 5.1. (a) Nitrogen adsorption/desorption isotherms and (b) pore size distribution of the
OS500. ...................................................................................................................... 121
Figure 5.2. The influence of contact time on (a) Pb2+ and (b) Cd2+ sorption onto OS500 and
the fit of the data by pseudo first order, pseudo second order, and Elovich equation
(0.1g char in 30 mL solution (0.01 M NaNO3), initial Pb2+ (Cd2+) concentration 1600
(1000) mg/L; room temperature 24 °C; initial solution pH 5). ................................. 125
Figure 5.3. Isotherms of (a) Pb2+ and (b) Cd2+ sorption onto OS500 (0.1g char in 30 mL
solution (0.01 M NaNO3); contact time 128 h; room temperature 24 °C; initial
solution pH 5). .......................................................................................................... 130
Figure 5.4. XRD patterns of OS500 before (OS500-B) and after Pb2+/Cd2+ sorption (OS500Pb and OS500-Cd). Minerals with peaks labeled: M, Muscovite 2M2,
K0.77Al1.93(Al0.5Si3.5)O10(OH)2; B, Barite, BaSO4; Q, Quartz, SiO2; C, Calcite, CaCO3;
D, Dolomite, CaMg(CO3)2; Ce, Cerussite, PbCO3; H, hydrocerrusite,
Pb3(CO3)2(OH)2. ....................................................................................................... 135
Figure 5.5. Comparisons between solution pH before and after (a) Pb2+ sorption and (b) Cd2+
sorption in equilibrium studies.................................................................................. 136
Figure 5.6. FTIR spectra of OS500 before (OS500-B) and after Pb2+/Cd2+ sorption (OS500Pb and OS500-Cd). ................................................................................................... 137
Figure 5.7. Sequential extraction results for (a) OS500-Pb, and (b) OS500-Cd (WS: water
soluble fraction, EX: exchangeable fraction, CB: fraction bound to carbonate, OX:
fraction bound to Fe-Mn oxides, OM: fraction bound to organic matter, RS: residual
fraction). .................................................................................................................... 138
Figure 6.1. (a) Nitrogen adsorption/desorption isotherms and (b) pore size distribution of the
OSDCs prepared at different pyrolysis temperatures. .............................................. 153
Figure 6.2. (a) Nitrogen adsorption/desorption isotherms and (b) pore size distribution of the
non-activated OSDC and activated OSDC obtained at 500 °C. ............................... 154
Figure 6.3. XRD patterns of (a) OS300, (b) OS500, (c) OS700, (d) OS900, and (e)
OS500N100-8M before and after Pb2+/Cd2+ sorption............................................... 158
Figure 6.4. FTIR spectra of (a) OS300, (b) OS500, (c) OS700, (d) OS900, and (e)
OS500N100-8M before and after Pb2+/Cd2+ sorption............................................... 159
Figure 6.5. Sequentially extracted (a) Pb2+ and (b) Cd2+ amounts in the OSDCs after Pb2+ or
Cd2+ sorption. Fractions of (c) Pb2+ and (d) Cd2+ in each step of sequential extraction
xi

(WS: water soluble fraction, EX: exchangeable fraction, CB: fraction bound to
carbonate, OX: fraction bound to Fe-Mn oxides, OM: fraction bound to organic
matter, RS: residual fraction). ................................................................................... 162
Figure B1. FT-IR spectra of the feedstock and solid residues after pyrolysis (SRC: solid
residue sample from control experiment without pyrolysis treatment; SROx: solid
residue from oily waste pyrolysis alone at X (°C) temperature and reaction time = 75
min; SROHx: solid residue from co-pyrolysis of oily waste and hog fuel at X (°C)
temperature and reaction time = 75 min and hog fuel addition = 20% .................... 218
Figure C1. Determination of pHPZC for OS500. Thick straight line represents pHinitial =
pHfinal line. ................................................................................................................. 220
Figure C2. Intraparticle diffusion plot for the sorption of (a) Pb2+ and (b) Cd2+ on OS500
(0.1g char in 30 mL solution (0.01 M NaNO3), initial Pb2+ (or Cd2+) concentration
1600 (1000) mg/L; room temperature 24 °C; initial solution pH 5). ........................ 221
Figure C3. Isotherms of (1) Pb2+ and (2) Cd2+ sorption onto OS500 (0.1g char in 30 mL
solution (0.01 M NaNO3); contact time 128 h; room temperature 24 °C; initial
solution pH 5). .......................................................................................................... 222
Figure D1. XRD patterns of OSDCs before Pb2+/Cd2+ sorption (OSDC-B). Minerals with
peaks labeled: M, Muscovite 2M2, K0.77Al1.93(Al0.5Si3.5)O10(OH)2; B, Barite, BaSO4;
Q, Quartz, SiO2; C, Calcite, CaCO3; D, Dolomite, CaMg(CO3)2; O, Calcium oxide,
CaO; Cl, Celsian, BaAl2Si2O8; Bo, Barium oxide hydrate, BaO2 (H2O)8; F,
Ferritungstite, (W, Fe)(O,OH)3; W, Witherite, BaCO3............................................. 227
Figure D2. FTIR spectra of OSDC before Pb2+/Cd2+ sorption (OSDC-B). ......................... 228

xii

GLOSSARY

AAEMs

Alkali and alkaline earth metallic species

AC

Activated carbon

ANOVA

Analysis of variance

API

American Petroleum Institute

ASTM

American Society for Testing and Materials

ATR

Attenuated total reflection

BET

Brunauer-Emmett-Teller

BJH

Barrett-Joyner-Halenda

BP

British Petroleum

BTEX

Benzene, toluene, ethylbenzene, and xylenes

CCME

Council of Ministers of the Environment

CEC

Effective cation exchange capacity

DP

Average pore diameter

DCM

Dichloromethane

DR

Dubinin-Radushkevich

DTG

Differential thermogravimetric analysis

F

Freundlich

FFU

Flocculation/flotation unit

FSOS500

Ferric-activated char

FTIR

Fourier transform infrared spectroscopy

GC-FID

Gas chromatography with a flame ionization detector

HCs

Hydrocarbons

HHV

High heating value

IARC

International Agency for Research on Cancer

ICP-OES

Inductively coupled plasma optical emission spectrometry

ILs

Ionic liquids

IL/Sl

Ionic liquid/sludge ratio

IPD

Intraparticle diffusion model

IUPAC

International union of pure and applied chemistry

L

Langmuir

MCL

Maximum contaminant level
xiii

NSO

Nitrogen sulfur oxygen

O/W

Oil-in-water

OSDC

Oily sludge-derived char

OSX

Oily sludge-derived char produced from pyrolysis at X (°C)
temperature

OSX-B

Oily sludge-derived char produced from pyrolysis at X (°C)
temperature before sorption

OSX-Cd

Oily sludge-derived char produced from pyrolysis at X (°C)
temperature after cadmium (Cd2+) sorption

OSX-Pb

Oily sludge-derived char produced from pyrolysis at X (°C)
temperature after lead (Pb2+) sorption

OS500N24-2M

Activated char obtained from hydrothermal treatment of
OS500 with 2M NaOH solution at 24 °C

OS500N100-2M

Activated char obtained from hydrothermal treatment of
OS500 with 2M NaOH solution at 100 °C

OS500N100-8M

Activated char obtained from hydrothermal treatment of
OS500 with 8M NaOH solution at 100 °C

OS500N900

Activated char obtained from chemical activation of OS500
with NaOH at 900 °C

OSHA

Occupational Safety and Health Administration

PAHs

Polycyclic aromatic hydrocarbons

PF-order

Pseudo first order

pHPZC

The point of zero charge

PHCs

Petroleum hydrocarbons

PS-order

Pseudo second order

PV

Total pore volume

Qt

Total extracted Pb2+/Cd2+ in sequential extraction (mg/g)

Qp

The amount of Pb2+/Cd2+ sorbed through precipitation (mg/g)

Qcf

The amount of Pb2+/Cd2+ sorbed through complexation with
free carboxyl/hydroxyl functional groups (mg/g)

Qcm

The amount of Pb2+/Cd2+ substituted with metal ions to form
complexes (mg/g)

RI

Potential ecological risk index

RP

Redlich-Peterson
xiv

RSM

Response surface methodology

SBET

The Brunauer-Emmett-Teller (BET) surface area

S

Sips

S/N

Signal to noise

So/Sl ratio

Solvent/sludge ratio

SRC

Original oily waste

SROX

Solid residue after pyrolysis at X (°C) temperature

SROHX

Solid residue from the co-pyrolysis at X (°C) temperature

T

Temkin

TGA

Thermogravimetric analysis

TOC

Total organic carbon content

TPH

Total petroleum hydrocarbon

TSA

Total surface acidity

TSB

Total surface basicity

UHQ

Ultra high purity

USEPA

US Environmental Protection Agency

VOCs

Volatile organic compounds

W/O

Water-in-oil

XRD

X-ray diffraction

xv

ACKNOWLEDGEMENT
First of all, I would like to express my gratitude and appreciation for my supervisor Dr. Jianbing Li
whose guidance, support and encouragement have been invaluable throughout this study. Without his
help and advice, I would not have been able to complete this research. Further, I would like to thank
my supervisory committee members: Dr. Liang Chen, Dr. Bill McGill, Dr. Ron Thring, and Dr. Todd
Whitcombe for their insightful comments, guidance, and discussions.
Special thanks to Dr. Joselito Arocena, for sharing his expertise. I would also like to thank Dr. Xiuyi
Yan from China University of Petroleum. She continuously provided encouragement and was always
willing to help me with my research. I would like to express my thanks to Dr. Art Fredeen, Dr. Neil
Hanlon, Dr. Ken Otter, and Dr. Ellen Petticrew as my course supervisors in my first year of PhD. They
encouraged me to take a broader view of my research and this later helped me to identify my research
questions.
This research was financially supported by the Natural Sciences and Engineering Research Council
of Canada and Fisheries and Oceans Canada’s Multi-Partner Oil Spill Research Initiative (MPRI).
Additional support from the UNBC Graduate Teaching Assistantship, Graduate Conference Travel
Award, Research Project Awards, and PhD Dissertation Completion Award were much appreciated.
Many thanks to Dr. Hossein Kazemian, Charles Bradshaw, and Erwin Rehl for their support and
professional advice with sample analysis at UNBC Northern Analytical Laboratory Service. I would
like to thank Conan Ma working at UNBC Chemstore for his support with sample delivery and lab
materials supply. I would also like to thank Dr. Umesh Pashotam for his patience and helpful
instructions as the supervisor of my teaching assistant position at UNBC. I would like to extend my
thanks to several UNBC faculty and staff: Jenia Blair, Dr. Liz Dunn, Dr. Daniel Erasmus, John
Orlowsky, Dr. Mike Rutherford, Doug Thompson, and Dr. Youqin Wang who lend me their lab
equipment and supplies, provided me assistance in sample preparation, or helped me in other ways.
My gratitude also extends to the graduate students at UNBC for sharing their knowledge, skills, and
thoughts; they are Talaat Bakri, Dr. Guangji Hu, Huan Liu, Marzieh Mortezapour, Jingjing Shi, and
Zeyi Tong. I would also like to extend my gratitude to all my friends in China and Canada for always
being supportive, inspiring, and understanding. The friendship from my friends has helped me to gather
my strength to pass through many hard times during my PhD study.
I would like to dedicate this work to my families for their unconditional supports and love. Special
thanks go to my mother, who is a strong woman and my role model, and my grandfather, who has
taught me to be brave and not afraid of any challenges in my life.

xvi

Chapter 1
GENERAL INTRODUCTION
1.1. Background
A considerable amount of oily sludge is generated by the activities involved in petroleum
industry such as drilling, production, transport, processing and distribution. It is estimated that
more than 11,000 tons of oily sludge is generated in refineries daily (da Silva et al., 2012). Oily
sludge can accumulate in a refinery at various points, such as pump, desalter, tank, and pipeline
(Kuriakose and Manjooran, 2001).
The properties of oily sludges vary depending on the source producing them, the nature of
the crude oil, the refinery processing scheme, and the equipment and reagents used in
wastewater treatment (Elektorowicz and Habibi, 2005; Mazlova and Meshcheryakov, 1999).
Generally, an oily sludge is a complex water-in-oil (W/O) or oil-in-water (O/W) emulsion,
typically consisting of 10–56 wt% oil, 30–85 wt% water, and 13–46 wt% solids (API, 2010;
Mazlova and Meshcheryakov, 1999). The stability of the emulsion depends on a film formed
at the oil-water interface that acts as a barrier to prevent droplet coalescence (Sullivan and
Kilpatrick, 2002). A variety of compounds including natural emulsifiers such as asphaltenes
and resins, carboxylic organic acids and fine inorganic particles play important roles in
enhancing emulsion stabilization (Kilpatrick, 2012). Oily sludge is classified as a hazardous
waste since it contains toxic compounds such as benzene, toluene, ethylbenzene, and xylenes
(BTEX), polycyclic aromatic hydrocarbons (PAHs), and some heavy metals (e.g., cadmium,
lead, and mercury) (USEPA, 2012). The improper disposal of oily sludge can lead to
environmental pollution and pose serious threats to public health and the environment (Hejazi
et al., 2003; Kuang et al., 2011; Robertson et al., 2007). Most of the oily sludge is
1

conventionally disposed by landfill or incineration (van Oudenhoven et al., 1995). However,
the use of a landfill is associated with various disadvantages, such as the need for large space,
low degradation efficiency, a long treatment period, and the risk of harmful compounds
leaching into the environment (da Silva et al., 2012; Renou et al., 2008). Moreover, more
restrictive disposal regulations (e.g., Ontario Environmental Protection Act Regulation 347)
have been enacted over the past few decades limiting the use of land disposal for hazardous
waste (da Silva et al., 2012; Revised Regulations of Ontario, 1990). Incineration can greatly
reduce the volume of waste but the costs of processing oily sludge by incineration are usually
very high due to the need for flue gas treatment (Ball et al., 2012; van Oudenhoven et al., 1995).
Over the last decade or so, various treatment methods, such as ultrasonic treatment, electrokinetics,

solvent

extraction,

pyrolysis,

oxidation,

land-farming,

bio-slurry,

and

solidification/stabilization, have been explored as methods to deal with oily sludge (Ferrarese
et al., 2008; Hu et al., 2015, 2017; Jasmine and Mukherji, 2019; Karamalidis and Voudrias,
2007; Marin et al., 2005; Yang et al., 2005; Zhang et al., 2012). Most of these technologies do
not achieve a reasonable compromise between satisfying environmental regulations and
treatment costs. Some of these technologies, such as bioslurry, are effective in removing
petroleum hydrocarbons (PHCs) but are very expensive. Others, such as land-farming, have a
relatively low cost but fail to meet the requirements for current or future regulations. Therefore,
the development of new technologies for treating oily sludge with high efficiency and low
costs is needed.
1.2. Statement of the problem
Recycling is the most desirable strategy for handing oily sludge since it enables the
petroleum industry to re-use valuable oil by reprocessing, reformulating or recovering energy
2

through burning (Elektorowicz and Habibi, 2005). Recycling of oily sludge can minimize the
disposal of wastes outside an industrial zone, reduce overall contamination, and decrease the
use of natural nonrenewable resources (Elektorowicz and Habibi, 2005). Solvent extraction is
a commonly used approach to recycling hydrocarbons from oily sludge. In this technology,
solvent is mixed with an oily sludge to ensure complete miscibility allowing semi-volatile and
non-volatile organic compounds to be removed from soil/water matrices effectively. Solvent
extraction can recover most useable PHCs from oily sludge, but this method needs a large
amount of organic solvent to achieve a promising oil recovery rate. The major concern of
applying solvent extraction to oily sludge is the toxicity of the organic solvents used to both
the process operators and the environment. Moreover, the volatile and flammable nature of
organic solvents makes them a potential explosion hazard (Schmid et al., 1998). There is a
growing interest in environmentally friendly solvents with high efficiency in the past few years.
Ionic liquids (ILs) are purely ionic materials with melting temperature below 100 °C and have
been considered as “green, eco-friendly solvents” because of their unique properties (Rogers
and Seddon, 2003). ILs have a very low vapor pressure, outstanding thermal stability, and are
non-flammable (Guzmán-Lucero et al., 2010). As a consequence of their unique properties,
researchers working in the petroleum industry have been examining potential applications for
ILs in the process of refining, recovery, or upgrade of oil (José-Alberto and Jorge, 2011).
Studies on the use of ILs in environmental remediation are still scarce, especially oil recovery
from oily sludge.
Pyrolysis is a thermo-chemical decomposition process of a material in a closed oxygen-free
reactor system with elevated temperatures (usually 400–600 °C). The application of pyrolysis
technology to oily sludge is attractive because pyrolysis can effectively recover the oil from

3

waste (Schmidt and Kaminsky, 2001) while also immobilizing heavy metal ions (Devi and
Saroha, 2014). Experiment on the co-pyrolysis of oily sludge with wood waste has also shown
that the overall yield of oil was greater than the yield obtained by the pyrolysis of the oily
sludge and the wood waste separately indicating a synergistic effect (Hu et al., 2017). Some
of the oil properties, such as high heating value (HHV) and carbon content, have been found
to be improved by co-pyrolysis (Q. Cao et al., 2009). In addition to the potential advantages of
increasing oil yield and improving oil quality, co-pyrolysis with other waste material could
even enhance the heavy metal immobilization in solid residues (Debela et al., 2012; Jin et al.,
2017). To study the fate of heavy metals during the thermal treatment of waste, several articles
have focused on the total heavy metal content and leaching behavior of residues after treatment
(Jin et al., 2014; Thurnau, 1996). Their results can hardly be used for examining the mobility
and eco-toxicity of heavy metals, which depend on their specific chemical forms or binding
patterns (He et al., 2010).
Growing interest has been shown in the application of oily sludge-derived char (OSDC) to
the removal of pollutants from wastewater in recent years. For example, Mohammadi and
Mirghaffari (2015) produced activated carbon (AC) from oily sludge through a chemical
method and they found a surface area of 327.96 m2/g. To further improve the adsorption
capability of an AC derived from oily sludge, Wang et al. (2017b) added a solvent de-oiling
step before activation. The surface area of the AC prepared with the de-oiling step reached
3292 m2/g. As a result, oily sludge could be a potential material for producing effective
adsorbent. However, there is a lack of information on the roles of different sorption
mechanisms for OSDC. Also limited information is available on how the preparation condition

4

(i.e., pyrolysis temperature and chemical activation) and resulting OSDC properties translate
into the heavy metal sorption capacities.
1.3. Objectives of the study
The main objective of this research is to develop novel technologies for oily sludge
management with added economic and environmental benefits. Based on literature review in
Chapter 2, the specific objectives include:


Evaluating the effect of IL (i.e., 1-ethyl-3-methyl-imidazolium tetrafluoroborate
([Emim][BF4])) on oil recovery from oily sludge by comparing to solvent extraction
without IL addition. Several experimental factors including solvent/sludge ratio,
shaking speed, extraction duration, and IL/sludge ratio were studied for their effects on
oil recovery. The optimal conditions for oil recovery in the two processes (i.e., solvent
extraction and IL enhanced solvent extraction) were also investigated.



Investigating the pyrolysis of oily waste containing heavy metals for both oil recovery
and heavy metal immobilization. Hog fuel (i.e. wood waste) from lumber production
was used as an additive, aiming to explore the reduced bioavailability and toxicity of
heavy metals in the pyrolysis char. The metal ion chemical speciation in the solid char
residue was examined to evaluate the ecological risk posed by oily waste treatment.
The effects of three selected experimental variables (pyrolysis temperature, reaction
time, and wood waste addition) on the process performance were studied using a
response surface methodology (RSM).



Evaluating the application potential of OSDC as a cost-effective sorbent for the
removal of Pb2+ and Cd2+ in aqueous solution. The mechanism of Pb2+ and Cd2+
sorption by OSDC was investigated using batch sorption studies, Fourier transform
5

infrared spectroscopy (FTIR), and X-ray diffraction (XRD). The contribution of
different mechanisms to the overall sorption of Pb2+ and Cd2+ by OSDC was also
quantified using a modified sequential extraction test.


Examining how the preparation condition (i.e., pyrolysis temperature and chemical
activation) of OSDC affected the heavy metal sorption capacities. To relate the heavy
metal removal performance of OSDCs to their properties, the adsorption of nitrogen,
pH drift method, Boehm titration, and element analysis were used to characterize the
surface features. The sorption mechanisms of heavy metal ions on activated and nonactivated OSDC were investigated by qualitative methods (i.e., FTIR and XRD) and
quantitative method (i.e., sequential extraction).

1.4. Organization of the dissertation
The dissertation is structured as follows: in chapter 2, literatures were reviewed and
summarized; in chapters 3–6, three different approaches to treat oily sludge was developed,
including IL-enhanced solvent extraction (chapter 3), co-pyrolysis with wood waste (chapter
4), and value-added utilization of oily sludge as sorbent to remove Pb2+ and Cd2+ from solution
(chapters 5 and 6); in chapter 7, the conclusion of this research and recommendations for future
research were provided.

6

Chapter 2
LITERATURE REVIEW
2.1. Sources of oily sludge
Oily sludge is one of the most significant solid wastes generated in petroleum industry. The
petroleum industry is usually divided into three major components: upstream, midstream, and
downstream. The upstream includes the exploration and production of crude oil. The
midstream includes transportation and storage of crude oil, while the downstream includes
refining and processing of crude oil. According to a study conducted by the American
Petroleum Institute (API), in 1995, US onshore and coastal oil-field exploration and production
operations generated approximately 278,180 tons of oily sludge from tank bottoms (ICF
Consulting, 2000). In the upstream operation, oily sludge might also originate from oil debris,
dehydration waste, used oils, etc (Hu et al., 2013; ICF Consulting, 2000).
During the refining process, impurities such as sulfur, nitrogen, and metal ions are separated
from petroleum. These impurities along with the residual hydrocarbons, mineral particles,
water, and chemicals form oily sludge (Kriipsalu et al., 2008). The demand for refined
petroleum products has been increasing worldwide due to rapid industrialization. According
to British Petroleum’s (BP) annual report, the refining throughput worldwide in 2018 was
13,188,076 m3/day, representing an 11% increase in throughput compared to 2008 (11,889,030
m3/day) (BP, 2019). In general, a higher refining throughput could bring about greater oily
sludge production (Hu et al., 2013). The total amount of oily sludge produced by the European
refineries in 1993 was 856,274 tons, which corresponded to 0.16 wt% of refinery throughput
(van Oudenhoven et al., 1995). Even based on optimistic estimates and assuming oily sludge

7

production is only 0.1 wt% of refinery throughput, more than 11,000 tons of oily sludge is
produced daily worldwide (da Silva et al., 2012).
Within refineries, the major sources of oily sludge include slop oil emulsion solid, separator
sludge, flocculation/flotation unit (FFU) sludge, and tank bottom sludge (Shie et al., 2000).
Depending upon the sources of the crude oil and the point in the processing where the oily
sludge is generated, it can vary significantly in its oil, water, and solid content. For example,
the oily sludge (i.e., separator sludge and FFU sludge) from wastewater treatment tends to
contain a higher percentage of water (API, 2010). In contrast, tank bottom sludge contains a
higher percentage of the solids (API, 2010). Other factors such as the type of crude oil, the
refinery processing scheme, and the chemicals and equipment used in the wastewater treatment
can also affect the composition of oily sludge (Mazlova and Meshcheryakov, 1999). As a
consequence of all the factors affecting its production, oily sludges can range across a range of
compositions, typically oil 10–56%, water 30–85%, and solids 13–46% by weight. (API, 2010;
Mazlova and Meshcheryakov, 1999).
2.2. Characteristics of oily sludge
The physical properties of an oily sludge, such as water solubility, boiling point, and vapor
pressure, are significantly affected by the oil, water, and solids content (API, 2010). For
instance, as the oil content of oily sludge decreases, the overall boiling point decrease, whereas
the volatility increase. Moreover, properties such as polarity and molecular weight of
hydrocarbons can be used to predict the other related properties in oily sludge (API, 2010).
The aromatic constituents in oily sludge have a higher boiling point, higher water solubility,
and lower vapor pressure than the aliphatic constituents (Gustafson et al., 1997). The
hydrocarbons with higher molecular weight have a higher boiling point, lower water solubility,
8

and lower vapor pressure as compared to those with lower molecular weight (Gustafson et al.,
1997).
The chemical composition of oily sludge is complex and depends on the type of petroleum
and its source. The oil phase of oily sludge ranges from 52–67% aliphatics, 15–31% aromatics,
7–8% nitrogen sulfur oxygen (NSO) containing compounds, and 10–12% asphaltenes (Mishra
et al., 2001; Mrayyan and Battikhi, 2005; Tahhan and Abu-Ateih, 2009). According to
Kriipsalu et al. (2008), the most abundant compounds among the aliphatic and aromatic groups
are alkenes, cyclic alkenes, benzene, toluene, ethylbenzene, and xylenes. Small amounts of
phenols and polycyclic aromatic hydrocarbons (PAHs) such as naphthalene, fluorene,
phenanthrene, and chrysene can be also found in oily sludge (Kriipsalu et al., 2008). Generally,
oily sludge is identified as either a water-in-oil (W/O) or an oil-in-water (O/W) emulsion
(Elektorowicz et al., 2006; Long et al., 2013). An emulsion is the mixture of two immiscible
liquids in which one of the liquids is stably suspended in the other (Kilpatrick, 2012). The fine
particles and natural emulsifiers such as asphaltenes and resins accumulate and form a film at
the oil-water interface, acting as a barrier to prevent droplet coalescence (Ali and Alqam, 2000).
Asphaltenes are condensed polyaromatic hydrocarbons that contain heteroatoms such as N, S,
and O polar groups (Sullivan and Kilpatrick, 2002). Resins are organic molecules consisting
of a hydrophobic end (alkyl chains) and a hydrophilic end (polar functional groups) (Sullivan
and Kilpatrick, 2002). These amphiphilic substances (i.e., asphaltenes and resins) play
important roles in stabilizing emulsion. In addition, long-chain carboxylic acids and fine
inorganic particles also contribute to the stability of emulsions of oily sludge. The emulsion
stabilization mechanisms associated with these three types of components were reviewed by
Kilpatrick (2012).

9

Heavy metals such as copper, lead, zinc, nickel, chromium, etc. are also present in oily
sludge in ionic form. It was reported that some metals (i.e., zinc, calcium, and magnesium) in
crude oil usually exist in combination with naphthenic acid as soaps, while other metals (i.e.,
vanadium, copper, and nickel) exist in the form of oil-soluble complexes such as
metalloporphyrins (Ali and Abbas, 2006). The heavy metal ion content in oily sludge varies
most likely due to the diversity of crude oil and differences in the methods for oil-water
separation and oil storage (Kriipsalu et al., 2008). High concentrations of heavy metal ions in
oily sludge have been reported in some studies, i.e., 500 mg/kg copper (Cu), 565 mg/kg lead
(Pb), 2100 mg/kg zinc (Zn), 1300 mg/kg nickel (Ni), 480 mg/kg chromium (Cr), 120 mg/kg
mercury (Hg), 2400 mg/kg vanadium (V), 3900 mg/kg calcium (Ca), 622 mg/kg magnesium
(Mg) (Table 2.1).

10

Table 2.1. Heavy metal content in oily sludge generated from various refineries

Type of oily
sludge
Tank bottom

Copper Lead
mg/kg mg/kg
214
565

Tank bottom

2–26

Tank bottom

55

Tank bottom

14

Open storage
500
impoundments
FFU
34–
370
Refinery oily
0.03
sludge
Refinery oily
32–
sludge
120
Refinery oily
sludge
a

13–93

The data was not obtained.

Zinc
Nickel Chromium Mercury Vanadium Calcium Magnesium Reference
mg/kg mg/kg mg/kg
mg/kg
mg/kg
mg/kg
mg/kg
769
125
119
–a
–
–
542
(da Rocha et al.,
2010)
34–
74–
0–19
4–18
2–6
–
–
–
(Alfutaisi et al.,
148
759
2007)
49
91
16
14
–
6
3900
622
(Jasmine and
Mukherji, 2015)
20
92
12
7
35
–
1056
96
(Shie et al.,
2000)
270
690
480
480
–
2400
–
–
(Admon et al.,
2001)
29–
520–
120–
12–110
19–120 66–380
–
–
(Kriipsalu et al.,
400
2100
1300
2008)
–
0.60
–
0.04
–
–
–
–
(Mishra et al.,
2001)
0.001– 7–80
17–25 27–80
–
–
–
19–24
(Bhattacharyya
0.12
and Shekdar,
2003)
4–9
65–84 4–8
1–3
–
9–26
310–
105–130
(Mazlova and
360
Meshcheryakov,
1999)

11

2.3. Toxicity of oily sludge
Oily sludge was classified as hazardous waste because of the presence of toxic components
such as benzene, toluene, ethylbenzene, and xylenes (BTEX), and PAH (USEPA, 2012). The
improper disposal of oily sludge may pose serious threats to public health and the environment.
The volatilization of oily sludge can cause volatile organic compounds (VOCs) in the
atmosphere, leading to human health risk to the onsite workers and surrounding communities.
Among these risks, the aromatic VOCs, including benzene, toluene, ethylbenzene, and isomers
of xylene (BTEX) are considered a primary health concern (Hejazi et al., 2003). Exposure to
BTEX can cause a variety of health problems such as cancer, liver and kidney damage, and
impairment of the central nervous system (Hejazi et al., 2003). The US Occupational Safety
and Health Administration (OSHA) has established 8 h work exposure limits for BTEX: 1ppm
(3.19 mg/m3) for benzene, 200ppm (753.6 mg/m3) for toluene, 100ppm (434 mg/m3) for
ethylbenzene, and 100ppm (434 mg/m3) for xylene (NIOSH, 2018). The Ministry of
Environment in British Columbia (BCMOE) has established the following threshold limits for
disposal of oily waste in landfill: 25 mg/kg for benzene, 150 mg/kg for toluene, 250 mg/kg for
ethylbenzene, and 250 mg/kg for xylene (BCMOE, 2016). When BTEX values are above these
limits, the oily waste must be pretreated before landfill.
The petroleum hydrocarbons (PHCs) and heavy metals from oily sludge may migrate
through the soil and contaminate the terrestrial environment. PHCs contamination could reduce
water retention capacity (i.e., wettability) of soils, leading to desiccation of surface soils on
slops and to severe erosional problems (McGill et al., 1981). The hydrophobic films formed
near the soil surface due to the accumulation of degradation products may reduce the water
availability while limiting water and gas exchange between the soil and the atmosphere
12

(Robertson et al., 2007). High concentration of PHCs may create nutrient-deficient conditions,
inhibit seed germination, reduce plant yields, slow plant growth, and directly kill plants
(McGill et al., 1981). It has been noted that microbial activity is stimulated by the addition of
PHCs to soil, which was probably due to the adaptation and growth of the microbial petroleumdecomposing population (McGill et al., 1981). The presence of PHCs increased the total
microbial number in soil; however, the population diversity is lower in PHC-contaminated soil
compared to unpolluted soil (McGill et al., 1981).
The petroleum hydrocarbons (PHCs) and heavy metal ions may leach into the groundwater
that is linked with other aquatic system. The fate of PHCs once they are released into aquatic
environment depends on the hydrologic conditions (e.g., tides and winds) of the receiving
water, the soil and water temperature, and the composition of PHCs (Hall et al., 1978). The
light volatile compounds are lost through volatilization (Wake, 2005). The dissolved/dispersed
hydrocarbon in water was due to the presence of the compounds with water solubility such as
aromatics (González et al., 2006). Some PHCs can be degraded by chemical and biotic
pathways, while others may incorporate into sediments and beach (Hall et al., 1978). The
effects of PHCs on marine organisms are broad and sometimes persistent, including growth
inhibition, respiratory distress, reduced reproduction, and bio-magnification of hydrocarbon
compounds through the food web (Wake, 2005).
Aromatic hydrocarbons are generally more toxic than aliphatics. Therefore, the quantity
and composition of aromatic hydrocarbons are of special concern in the evaluation of toxicity
of PHC mixtures. Polycyclic aromatic hydrocarbons (PAHs) are chemical compounds
containing two or more benzene rings in a linear, cluster, or angular arrangement. PAHs with
four or more benzene rings are often mutagenic and/or carcinogenic (Pacheco and Santos,

13

1997). The International Agency for Research on Cancer (IARC) classifies some PAHs as
known, possibly, or probably human carcinogens (IARC WGECRH, 2010). Among these are
benzo(a)anthracene, chrysene, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene,
indone(1,2,3-cd)pyrene, and dibenzo(a,h)anthracene (IARC WGECRH, 2010). Humans and
other ecological receptors can absorb PAHs through inhalation, ingestion, or direct dermal
contact (K.-H. Kim et al., 2013). Health effects from exposure to PAHs include eye and skin
irritation, nausea and vomiting, inflammation, decreased immune efficiency, cataracts, kidney
and liver damage, skin, lung, bladder, and gastrointestinal cancers, and DNA damage (K.-H.
Kim et al., 2013). The Canadian Council of Ministers of the Environment (CCME) set the soil
quality guideline value for benzo(a)pyrene, the most carcinogenic PAH, at 0.6 mg/kg or 5.3
mg/kg, based on incremental lifetime cancer risks (ILCR) of 10–6 or 10–5, respectively (CCME,
2008a). According to the US EPA, the maximum contaminant level (MCL) for benzo(a)pyrene
in drinking water is 0.0002 mg/L (USEPA, 2015).
Heavy metal ions from oily sludge pose a risk to public health because of their toxicity,
persistence, high solubility, mobility, and biological accumulation. Some heavy metals (e.g.,
copper, cobalt, iron, manganese, and zinc) are essential for living organisms but excessive
levels of essential metals are harmful. The non-essential metals (e.g., cadmium, lead, and
mercury) undergo bio-magnification and have the potential to cause serious health effects, such
as liver and kidney damage, brain and nerve damage, and cancers (Wu et al., 2016). The US
EPA established strict MCL goals for heavy metal ions in drinking water: 1.3 mg/L for Cu,
0.015 mg/L for Pb, 0.1 mg/L for Cr, and 0.002 mg/L for Hg (USEPA, 2015). The coexistence
of PHCs and heavy metals in oily sludge contaminated soil could have complex impacts on its
microbial community. Some studies have demonstrated that heavy metal ions can retard the

14

rate of oil biodegradation in the environment due to the presence of toxic metal ions inhibiting
the activity of the organic-degrading microorganisms (AL-Saleh and Obuekwe, 2005; Said and
Lewis, 1991). However, the stimulating effects of heavy metal ions on microbial processes
have also been documented (Chen et al., 1997; Shen et al., 2005). For example, Shen et al.
(2005) reported the interaction between Zn and phenanthrene enhanced urease activity over an
incubation time of 21 days.
2.4. Techniques for treating oily sludge
A plausible solution in oily sludge management should consider the implementation of the
3Rs (Reduce, Reuse, and Recycle) policies and strategies (Demirbas, 2011):
a) Reduce: action that reduces the amount of oily sludge generated;
b) Re-use: technology that focuses on the re-use of oily sludge without treatments that
modify its physicochemical properties; and
c) Recycle: technology that allows a profit to be generated from oily sludge by altering
its physicochemical properties (da Silva et al., 2012).
The selection of the oily sludge treatment will always be site specific and rely on factors
such as the amount and type of oily sludge available, the cost of treatment, time, and legal and
environmental considerations. Comprehensive reviews of physical, chemical, and biological
processes specifically for oily sludge treatment are available in the literature (da Silva et al.,
2012; Hu et al., 2013). A summary of oily sludge treatment methods is shown in Table 2.2.

15

Table 2.2. Summary of technologies for the oily sludge treatment

Technology

Description

Advantages

Disadvantages

Separation

Solids/liquids separation by floatation or
centrifugation

Easy to operate, low
space requirements

Viscosity reducing pretreatment requirement, noise
problem for centrifugation

Landfill

The isolation of oily sludge from air and water is
achieved through the use of thick layers of
impermeable clay and synthetic liner

Relatively low cost,
large treatment capacity

Slow process, requiring a large
space, leachates may need
proper treatment

Electrokinetic (EK) The application of a low level direct current
electric potential through electrodes, which are
placed into the contaminated soil, ionic
contaminations are transported to the oppositely
charged electrode by electromigration, soluble
contaminants may be removed by electroosmotic
flow

Cost-effective, has the
potential to remove coexisting heavy metal and
organic pollutants from
soils

Low removal efficiency for
non-polar organic pollutants,
long remediation time

Solvent extraction

Sufficient solvent is mixed with the oily sludge
to ensure complete miscibility, semi-volatile and
non-volatile organic compounds can be
effectively removed from soil/water matrices

Simple, fast, effective,
energy recovery

Large volume of solvent
consumed, high cost,
depending upon solvent may
not be environmentally sound

High temperature oxidation which converts oily
sludge into gaseous products and ash

Waste reduction,
complete removal of

The generation of off-gases
containing PAH and BTEX,
auxiliary fuel is required

Physical methods

Chemical
methods
Incineration

16

Technology

Description

Advantages

Disadvantages

PHCs, heat value can be
reused
Oxidation

PHCs are degraded through chemical (e.g.,
Fenton’s reagent and ozone) or other advanced
oxidation reactions (e.g., photocatalytic
degradation)

Relative insensitive to
external disturbances
(e.g., temperature), short
treatment duration

High cost, a large volume of
chemical oxidant is needed

Stabilization/
S/S involves the addition of binders to achieve:
Rapid treatment, able to
Solidification (S/S) (a) stabilization of oily sludge by converting
immobilize heavy metals
contaminants into a less toxic and/or less soluble
form, and (b) solidification of the oily sludge into
a durable matrix with high structural integrity

Risk associated with long term
stabilization, an increase in
the volume of waste

Pyrolysis

Thermal desorption of organic materials into a
solid (char) and volatile matter (liquid and gases)
in an inert environment

Fast, energy recovery,
waste reduction, and less
secondary pollution, able
to immobilize heavy
metals

High energy consuming, high
capital, maintenance, and
operational costs

Land-farming

Oily sludge is scattered and mixed into the thin
soil layer (≤1.5 m) in a controlled manner
allowing the microbiota in the ground to act as a
degrading agent

Low cost, simple
operation, low energy
consumption, large
treatment capacity

Slow process, not efficient in
degrading the heavy
components of petroleum,
VOCs emission problem, risk
of underground water
pollution

Biopile

The piling of oily sludge contaminated soil into
piles, aerobic microbial activity is situmlated by
aeration and the addition of minerals, nutrients,
and moisture

Faster and more effective Higher cost and smaller
than land-farming, VOCs treatment capacity than landemissions are controlled farming

Biological
methods

17

Technology

Description

Advantages

Disadvantages

Bio-slurry

Bio-slurry dissolves oily sludge into an aqueous
phase, and microbially degrading the PHCs, it
allows the addition of nutrients, oxygen,
microorganisms, and surfactants

Fastest biodegradation
approach, VOCs
emission is controlled

High cost, the fine solids after
bio-slurry treatment still need
dewatering and disposal

Phytoremediation

Phytoremediation uses plants and their associated
microorganisms to remove pollutants from the
environment or to make them harmless, plants
take up, accumulate, stabilize, and/or degrade
pollutants

Low cost, aesthetically
pleasing, minimum
environmental
disturbance

Slow process, climate may
limit plant growth, disposal of
the plants used is required,
contaminants may enter the
food chain

(Source: da Silva et al., 2012; Hu et al., 2013; Khan et al., 2004)

18

Physical separation is often used as a pretreatment technique prior to other techniques, such
as landfilling and incineration (van Oudenhoven et al., 1995). In the past few decades, landfill
has been one of the most commonly used methods for the disposal of oily sludge (van
Oudenhoven et al., 1995). The disposal in landfills needs much space and it is still possible for
the harmful compounds, especially PAHs and heavy metals, to leach out of the treated soil and
contaminating the surrounding ecological system (Renou et al., 2008). Moreover, the use of
landfills is being gradually prohibited or restricted because of more restrictive disposal
regulations (da Silva et al., 2012; Elektorowicz and Habibi, 2005). According to the Hazardous
Waste Regulation issued by BCMOE, oily sludge containing more than 10 % oil (by weight)
cannot be accepted at any landfills in British Columbia (BCMOE, 2016). Consequently, the
use of landfill is only considered when other techniques, such as re-using or recycling, are not
readily available.
Another common method of disposing of oily sludge is incineration. Although incineration
can greatly reduce the volume of waste, it has certain limitations such as the need for auxiliary
fuel, the emission of hazardous off-gases (PAH, BTEX, dioxins, and furans), and high cost (da
Silva et al., 2012). To meet air pollution regulations, the off-gases must be cleaned before being
discharged into the atmosphere. Capital equipment costs for an incinerator that processes
between 3 and 10 tons h–1 range from US$ 3.5–5 million (Ball et al., 2012). The operational
costs associated with remediating petroleum-contaminated solids by incineration can range
from US$140 to US$700 per tonne (Ball et al., 2012; van Oudenhoven et al., 1995).
Biological treatment by land-farming or by using a biopile may have large treatment
capacity and low operational cost, but they are often ineffective in removing high molecular
weight hydrocarbons (Khan et al., 2004). Additionally, the microbial degradation process in

19

land-farming and biopile can be very time-consuming (from 6 months to 2 years) (Khan et al.,
2004). In terms of a bio-slurry treatment, the biodegradation process occurs in a confined and
highly controlled system. The conditions such as temperature, moisture, pH, oxygen, nutrients,
supplementation of microorganisms, and addition of surfactant are controlled, which results in
an increase in the rate of biodegradation (Ball et al., 2012). The treatment times for bio-slurry
can range from less than 1 month to more than 6 months (Khan et al., 2004). The major obstacle
to applying this technology to field-scale oily sludge treatment is the high cost. For example,
Castaldi (2003) reported the cost of operating a bio-slurry process for oily sludge was US$625
per ton. In comparison, an estimate for land-farm bioremediation of hydrocarbon-contaminated
soil (based on 2700 m3 of material) was reported to be US$10.3 per ton (Line et al., 1996).
In summary, some treatments such as incineration are effective in totally removing PHCs
from oily sludge but have the highest costs. Other treatments such as land-farming have lower
costs but their degradation process can be very time-consuming and/or they fail to meet the
requirements for the reduction of large molecular weight species and metal ions in the oily
sludge. As the basic theme of United Nations Conference on Environment and Development
(UNCED) in 1992 stated, “Sustainable development” is commonly understood as development
that “meets the needs of the present without compromising the ability of future generations to
meet their own needs”. A sustainable technique for oily sludge management needs to be both
environmentally effective and economically affordable (Morrissey and Browne, 2004). Most
of the available techniques do not satisfy a sustainable development approach. There is a
demand for the development of new techniques for recycling and treating oily sludge with
added environmental and economic benefits.

20

2.5. Ionic liquid-enhanced solvent extraction
2.5.1. Solvent extraction
2.5.1.1. Selection of solvent
Solvent extraction is a simple process achieved by mixing an oily sludge and a solvent at
proportions which ensure adequate miscibility of the oil in the solvent (Al-Zahrani and Putra,
2013). The main advantages of solvent extraction are low energy requirement, high efficiency
of oil recovery, short treatment duration, and the possibility of solvent reuse (Hu et al., 2013).
In solvent extraction process, the solvent molecules tend to diffuse into the oil phase since the
viscosity of the solvent is much lower than that of the oil (X. Li et al., 2012). The soluble
materials in the oily sludge readily mix into the solvent until they reach an equilibration
between the two media. However, the insoluble fractions tend to stay on the surface of the sand
grains, forming a porous network structure (Funk, 1979). Additionally, some lighter fractions
can be attached to the surface of insoluble asphaltenens, which resulted in the increase of the
amount of soluble materials in the network at equilibrium (X. Li et al., 2012).
From the thermodynamic point of view, solvent extraction of oily sludge can be seen as the
mixing/dissolving of solvent and solute (i.e., oil components). The solubility parameter is an
important physical property which characterizes the overall capability of a solvent to dissolve
a solute. In 1950, Hildebrand and Scott first introduced the solubility parameter and
theoretically described the relationship between solubility, heat of evaporation, and the
enthalpy (heat) of mixing (Hildebrand and Scott, 1950). The solubility parameter is defined as
the square root of the cohesive energy density:
∆Hv −RT 1⁄2

δ=(

V

21

)

(2.1)

where δ is the Hildebrand solubility parameter, ∆Hv is the heat of evaporation, R is the gas
constant, T is the absolute temperature, V is the molar volume of the solute. According to the
Hildebrand solution theory (eq. 2.2), if solvent is miscible with solute, which means enthalpy
of mixing is small and their solubility parameters would tend toward equality (δ1 = δ2 ).
∆H = Vm φ1 φ2 (δ1 − δ2 )2

(2.2)

where ∆H is the enthalpy of mixing, Vm is the molar volume of the solution, φ1 and φ2 are the
volume fractions of the solvent and solute, respectively, and δ1 and δ2 are the solubility
parameters of the solvent and solute, respectively.
In the middle 1960s, Hansen proposed the splitting of the Hildebrand solubility parameter
into three components derived from different types of cohesive energy (i.e., dispersion, polarity,
and hydrogen bonding) (Hansen, 2007):
δ = (δ2d + δ2p + δ2h )

1⁄2

(2.3)

where δd is the dispersion force parameter, δp is the polar force parameter, and δh is the
hydrogen-bonding force parameter. Based on the theory of Hanson solubility, good solvents
always possess parameter distributions (i.e., dispersion force, polar force, and hydrogenbonding force) in a specific range depending upon the solute (Hansen, 2007). Wang et al. (2014)
reported good solvents for the dissolution of bitumen generally have relatively high values for
δd and δh and a low value for δp .
Therefore, in the solvent extraction of oil from oily sludge, the dissolving capacity of
solvent is a combination of the Hildebrand solubility parameter ( δ ), and the parameter
distribution of the acting force parameter (δd , δp , and δh ). According to Wang et al. (2014), a
solvent with a solubility parameter close to that of asphaltene and a parameter distribution
close to that of toluene is desirable for the bitumen recovery from oil sands.
22

2.5.1.2. Parameters affecting extraction performance
Extraction performance is also affected by operational factors such as temperature, solventto-sludge ratio, extraction duration, and stirring rate. The increase of temperature can decrease
the viscosity of the oil and increase its solubility in a solvent, thereby increasing the oil
recovery. However, high temperature means a large energy consumption and the loss of PHCs
and solvent.
An increase in the solvent-to-sludge ratio will enhance the interaction between the oil and
solvents, as well as increase the concentrations gradient between liquid-solid phases, which
may result in a higher oil recovery. Additionally, the quality of recovered oil can be improved
with increasing solvent-to-sludge ratio, which is evidenced by the reduction in ash, carbon
residue, and asphaltene content in recovered oil when more solvent is used (Zubaidy and
Abouelnasr, 2010). The aspheltene content in the extracted oil generally has a positive
relationship with the concentration of oil in the solvent phase, which is likely due to the
mechanism that it is the extracted oil and not the solvent that extracts the asphaltene from oily
sludge (Wang et al., 2014; Zubaidy and Abouelnasr, 2010). With increasing solvent-to-sludge
ratio, the concentration of oil in the solvent phase decreased (if the increase of the solvent mass
exceeded the increase of extracted oil), which contributed to a lower asphaltene content (Wang
et al., 2014). There is also a possibility that an increasing solvent-to-sludge ratio can decrease
the oil recovery (Wang et al., 2014; Wu and Dabros, 2012). At high solvent-to-sludge ratio,
some asphaltenes self-associate and form larger aggregations, leading to a lower oil recovery
(Wu and Dabros, 2012).
The optimal extraction duration is affected by temperature, solvent-to-sludge ratio, and the
stirring rate. As the extraction duration increases, an extraction equilibrium is reached and any
23

further increases in time are unlikely to significantly improve oil recovery. Finding the optimal
extraction time is important in order to save costs for an extraction process. The oil recovery
from oily sludge can be promoted by increasing the stirring rate. The diffusion of the dissolved
materials into the solvent phase can be facilitated by stirring or agitation. In addition, external
shear forces break the porous network structure into smaller particles (X. Li et al., 2012). These
particles tend to be suspended in the solvent phase thus increasing the interfacial area between
bitumen and solvent thereby leading to a higher rate of mass transfer.
2.5.2. Ionic liquid
The utilization of solvent is essential to solvent extraction. However, it also limits the
extension of this technology to field-scale oily sludge treatment. The cost of recovering solvent
from the spent or extracted sand/clay is prohibitive, especially for high-boiling-point solvents.
Moreover, organic solvent exposure is associated with various health hazards, such as
reproductive damage, liver and kidney damage, cancer, and toxicity to the nervous system
(NIOSH, 1977). Ionic liquids (ILs) are organic salts that are composed of organic cations with
organic or inorganic anions and exist in liquid state at or close to room temperature. ILs possess
high thermal stabilities and negligible vapor pressure making them favorable in the
applications of petroleum industry (José-Alberto and Jorge, 2011). More importantly, ILs can
be customized by choosing the appropriate combination of anion and cation for specific
applications (Bera and Belhaj, 2016).
Various studies have been reported on the application of ILs to reduce the precipitation of
asphaltenes and other organic compounds while enhancing the stability of crude oils
(Boukherissa et al., 2009; Hu and Guo, 2005; Subramanian et al., 2015). Hu and Guo (2005)
studied the effect of the structures of ILs on the inhibition of asphaltnene precipitation in CO224

injected reservoir oils. It was found that inhibition abilities of ILs p-alkylpiridinium chloride
([Cnpy][Cl]) decreases with the increase of alkyl chain length from n=4 to n=12 (Hu and Guo,

2005). The ILs with high anionic charge density and low cation charge density were effective
in inhibiting asphaltene precipitation (Hu and Guo, 2005). According to the mechanism they
proposed, the local non-neutrality of the charge density promotes the acid-base interaction
between ILs and the asphaltene, thus breaking intramolecular associations within the
asphaltenes (Hu and Guo, 2005). Boukherissa et al. (2009) evaluated imidazolium ILs
consisting of either a boronic acid or a propenyl group with bromide anions as dispersants for
petroleum asphaltenes. They observed that the viscosity of the oil was reduced from 16 Pa·s
to 3 Pa·s as the IL/asphaltenes ratio increased from 0 to 2 (2 wt% asphaltene at nheptane/tolune = 2) (Boukherissa et al., 2009). The presence of boronic acid favored the
interaction between the Lewis-acid moiety of the ILs and the basic regions of the asphaltene
molecules, which consequently reduced the viscosity (Boukherissa et al., 2009). Subramanian
et al. (2015) found small amounts of ILs (between 1ppm and 10 ppm), dissolved in toluene,
reduced the viscosity of crude oil up to 35 % (with reference to the oil containing toluene).
They attributed this effect to the intermolecular interactions, such as π-π, aliphatic, and acidbase interactions between the functionalized molecules and the asphaltene molecules
(Subramanian et al., 2015).
Some ILs have similar surface activity and micelle formation capacity as surfactants and
are thus considered to be good candidates for enhanced oil recovery process (Bera and Belhaj,
2016). Guzmán-Lucero et al. (2010) evaluated 10 ILs as de-emulsification agents to break
water-in-oil (W/O) emulsions. All of the ILs tested showed good demulsifying activities in
water-in-medium-crude-oil emulsions. The demulsification efficiency were greatly affected by

25

the structure of ILs. It was observed that the ILs with longer alkyl chain or more hydrophobic
anion provided higher separation efficiency of W/O emulsions (Silva et al., 2013). Lemos et
al. (2010) studied the effect of 1-octyl-3-methylimidazolium tetrafluoroborate ([Omim][BF4])
and 1-octyl-3-methylimidazolium hexafluorphosfate ([Omim][PF6]) on the demulsification of
a W/O emulsion. They pointed out that the greater anion size of [Omim][PF6] hindered the
aggregation of this IL in the medium, which resulted in a greater demulsification efficiency
(Lemos et al., 2010). In addition, the use of microwave irradiation can dramatically increase
efficiency to the demulsification of W/O emulsion by ILs. For example, very high
demulsification efficiency (beyond 95%) was obtained in 30 min by the joint use of
trioctylmethylammonium chloride (TOACl) and microwave irradiation (Guzmán-Lucero et al.,
2010).
Apart from the oil-water separation, ILs are also used in the extraction of bitumen from oil
sands. Painter et al. (2010a) evaluated the extra heavy oil production from Utah tar sands using
ILs. Tar sands were mixed with three hydrophilic ILs, [Bmmim][BF4], [Bmim][CF3SO3], and
[Emim][Ac] (Table 2.3), and toluene to give a rise to a three-phase system. The tested ILs and
toluene were effective in breaking up large particles of tar sands and giving more than a 90%
yield of bitumen (Painter et al., 2010a). The cleanest separation between the phases was given
by [Bmmim][BF4] while extractions using [Bmim][CF3SO3], and [Emim][Ac] resulted in
some incorporation of IL into the bitumen phase (Painter et al., 2010a). Williams et al. (2010)
reported the use of the IL [Bmmim][BF4] allowed for a cleaner separation compared to solvent
extraction alone. The clay bands were clearly displayed in the infrared spectrum of the toluene
extract of low-grade oil sands (Williams et al., 2010). Such phenomenon was attributed to the
solids in the weathered oil sands possessing a strong hydrophobicity, resulting in an adhesive

26

force between the bitumen and fines (Liu et al., 2005). However, there were no mineral fines
in the extracts utilizing IL, while some clay fines remained suspended into the middle IL phase
(Williams et al., 2010). This might be because the electrostatic interaction between the IL and
the mineral fines are much more favorable than those between bitumen and fines (Williams et
al., 2010).
Nevertheless, ionic liquids with high viscosity, like [Bmmim][BF4], could strongly adhere
to the inner wall of pipelines and blocked the equipment increasing the difficulty of its
industrial application. In this regard, some authors applied ILs with low viscosity (e.g.,
[Emim][BF4]) to enhance solvent extraction (Li et al., 2011; Sui et al., 2016; Tourvieille et al.,
2017). Li et al. (2011) reported the maximum bitumen recovery was 86% at the acetone-to-nheptane ratio of 2:6 (v/v) in solvent extraction alone. With IL [Emim][BF4] enhancement (1.3
mL/g of IL/oil sand ratio), an additional improvement of 10% of bitumen recovery was
observed (Li et al., 2011). Since 92 % (by weight) bitumen recovery was obtained even when
the IL was reused 10 times, the solvent extraction technology using composite solvent
enhanced by IL was encouraged to be extended from bench scale to the oil sand industry by Li
et al. (2011). Sui et al. (2016) examined the effect of operational parameters (i.e., IL-to-oil
sands ratio, agitation time, and conditioning time) on bitumen recovery from oil sands using
IL enhanced solvent extraction. It was found that bitumen recovery increased as IL-to-oil sands
ratio and agitation time increased, whereas the conditioning time had a negligible effect on
bitumen recovery. Tourvieille et al. (2017) assessed the impact of variables (i.e., temperature,
contact time, solvent-to-oil sands ratio, IL-to-oil sands ratio, water-to-IL ratio, solvent, and IL
types) on bitumen recovery and bitumen quality. It was observed that the IL-to-solids ratio
(0.75–2 g/g) and type of IL had no significant effect (P > 0.05) on bitumen recovery

27

(Tourvieille et al., 2017). However, types of IL used does play a significant role in aluminum
(Al) content in recovered bitumen. [Amim][DCA] leads to an increase of Al content in bitumen,
whereas [EtNH3][NO3] was significant in decreasing Al content in bitumen (Tourvieille et al.,
2017). [Amim][DCA] was less capable in retaining inorganic material because of its longer
hydrocarbon tail and increased π-π interactions (Tourvieille et al., 2017). Conversely, the
shorter hydrocarbon tail and higher viscosity for the [EtNH3][NO3] may be responsible for its
efficiency in capturing fines from bitumen (Tourvieille et al., 2017). Table 2.3 provides the
structures and physical properties of some of the ionic liquids employed in various studies of
oil recovery from tar sands.

28

Table 2.3. Summary of ionic liquids used in bitumen recovery from oil sands

Ionic liquid

1-butyl-2,3-dimethylimidazolium tetrafluoroborate
([Bmmim][BF4])

Structure

N

Viscosity

CH3

F
F

B

CH 3

N

F

F

Density

N

CH3

N

O

F
F
F

S

O

(g/cm )

at 25 °C

at 25 °C

456
(Bou Malham
and Turmine,
2008)

1.19
(Bou Malham
and Turmine,
2008)

>90% bitumen
recovery, high
viscosity, clean
separation (no IL and
fines in the bitumen)

(Li et al.,
2011;
Painter et
al., 2010a)

75.3
(Ge et al.,
2008)

1.30
(Ge et al.,
2008)

>90% bitumen
recovery, less clean
(small amounts of IL
was found in the
bitumen)

(Painter et
al., 2010a)

141.1
(Nazet et al.,
2015)

1.09
(Nazet et al.,
2015)

>90% bitumen
recovery, less clean
(small amounts of IL
was found in the
bitumen)

(Painter et
al., 2010a)

66.5
(Huddleston et
al., 2001)

1.25
(Seddon et al.,
2002)

Up to 95% bitumen
recovery, relatively
low viscosity, clean
separation (no IL and
fines in the bitumen)

(Li et al.,
2011; Sui et
al., 2016)

O

CH 3

1-ethyl-3-methyl-imidazolium
acetate ([Emim][Ac])

N
N

CH3
O
O

CH3

CH3

1-ethyl-3-methyl-imidazolium
tetrafluoroborate
([Emim][BF4])

N

CH3
F

N
CH3

F
B

F

References

(mPa·s)

CH 3

1-butyl-3-methyl-imidazolium
trifluoromethanesulfonate
([Bmim][CF3SO3])

Performance

3

F

29

Ionic liquid

ethylammonium nitrate
([EtNH3][NO3])

Structure

O
H 3C

NH 3 O

N
O

1-Allyl-3-methylimidazolium
dicyanamide ([Amim][DCA])

N
N

CH3
N
CH2

N

Viscosity

Density

(mPa·s)

(g/cm3)

at 25 °C

at 25 °C

32.0
(Greaves et al.,
2006)

1.22
IL type had no
(Greaves et al., significant effect on
2006)
bitumen recovery, low
aluminum content in
the bitumen

(Tourvieille
et al., 2017)

14.0
(Tourvieille et
al., 2017)

–

(Tourvieille
et al., 2017)

N

30

Performance

IL type had no
significant effect on
bitumen recovery,
high aluminum
content in the bitumen

References

2.6. Pyrolysis
2.6.1. Pyrolysis in oily sludge treatment
Pyrolysis is a thermal disposal method which converts organic materials into a carbon-rich
solid (char) and volatile matter (liquid and gases) in an inert environment. It has the unique
characteristic of being able to break down large molecules into smaller ones (“cracking”). In
pyrolysis, the emission of hazardous materials (e.g., NOX, SO2, and heavy metals) may be
decreased relative to straight incineration (Li et al., 1999). Moreover, the reduced air flow rate
and lower temperature make emission controls easier for pyrolysis than for incineration (Li et
al., 1999). The most important advantage of pyrolysis is that it can generate solid, liquid, and
gaseous fractions all of which can be used as fuels in subsequent applications or to drive the
pyrolysis process while the liquid fuels are easily stored and transported (Martínez et al., 2014).
The solid product of pyrolysis (i.e., char) has the potential ability to enhance soil quality and/or
immobilize metal ions due to its unique properties (e.g., retention of water and nutrients,
porous structure, and cation exchange capacity) (Mullen et al., 2010; Uchimiya et al., 2010).
Pyrolysis is a promising strategic approach to the sustainable management of oily sludge
having the ability to not only minimize the solid waste, but also to yield valuable products.
Schmidt and Kaminsky (2001) reported that about 70–84 % of the oil can be separated from
oily sludge during the pyrolysis in fluidized bed reactors at temperatures from 460–650 °C.
Shie et al. (2000) studied the kinetic and thermal conversion behavior of oily sludge pyrolysis
at various heating rates (i.e., 5.2, 12.8, and 21.8 °C/min) by thermogravimetric analysis (TGA).
The TGA curves were represented by the two- and three-reaction models relatively well (Shie
et al., 2000). As the heating rate increased, the peak value of reaction rate increased, and the
TGA curve shifted to higher temperatures (i.e., the weight loss occurred at higher temperatures)
31

(Shie et al., 2000). According to the differential thermogravimetric analysis (DTG) curve of
oily sludge, the major mass loss of the oily sludge was observed in a temperature range of 100–
550 °C with peak mass losses at temperatures of 110–150 and 150–550 °C probably due to the
decomposition of light PHCs and complex PHCs, respectively (Hu et al., 2017). At
temperatures above 550 °C, the decomposition of inorganic compounds (such as CaCO3,
MgCO3, etc.) might take place (Liu et al., 2009; Wang et al., 2007).
Chang et al. (2000) investigated the major products obtained from the pyrolysis of oily
sludge at 600 °C. It was observed that the mass fraction of solid residues, liquids, and gases in
the final product was 13.10 wt%, 69.93 wt%, and 3.57 wt%, respectively (Chang et al., 2000).
The liquid product consisted of 66.58 wt% light gas oil with a boiling range between 204–
338 °C, indicating the distillation characteristics of liquid product was close to that of diesel
oil (Chang et al., 2000). Their results also indicated that the major gaseous products excluding
N2, were CO2 (50.88 wt%), hydrocarbons (HCs) (25.23 wt%), H2O (17.78 wt%), and CO (6.11
wt%) while the HCs gases mainly consisted of low molecular weight organic species,
especially C1–2 (51.61 wt%). The research by Liu et al. (2009) has shown that around 80% of
total organic carbon content (TOC) in oily sludge can be converted into HCs by a pyrolysis
process, with significant HCs yields occurring in the temperature range of 327–450 °C. They
found that a higher heating rate resulted in a higher peak value of HCs evolution rate and a
higher temperature for its occurrence.
2.6.2. Effects of parameters on liquid production
Qin et al. (2015) found oil yield increased initially with pyrolysis temperature with a
maximum oil yield (59.2 wt% of feed oily sludge) occurring at 500 °C, which was consistent
with other studies (Hu et al., 2017; Shen and Zhang, 2003). The decrease in oil yield above
32

500 °C may be attributed to the secondary decomposition reactions which break the oil into
lighter, gaseous hydrocarbons (Qin et al., 2015). Liu et al. (2008) studied the effects of
temperature on the composition of oil produced from pyrolysis of oily sludge in a fixed bed
reactor. It was found that as the temperature increased from 350 to 550 °C, the light oil content
increased from 49.8 to 69.2 wt%, whereas the heavy oil content decreased from 31.2 to 10.8
wt%. This result is in agreement with another pyrolysis study carried out in a fluidized bed
reactor (Shen and Zhang, 2003). According to Liu et al. (2008), the alkene and aromatic content
in the oil product increased with rising temperature. The decarboxylation reaction
accompanied by cracking might be responsible for the rise in the alkene compounds (Jindarom
et al., 2007b). The formation of aromatic compounds at high temperatures was mainly
attributed to Diels-Alder reactions (Park et al., 2008; Zhang et al., 2011).
Shen and Zhang (2003) found oil yield decreased from a maximum of 30 to 21 wt% with
increasing gas residence time from 1.5 to 3.5 s due to the occurrence of secondary cracking
reactions. Their results also indicated that the proportion of lighter compounds increased when
the gas residence time increased. Similarly, Wang et al. (2007) found the operation of holding
at 400 °C for 20 min decreased the amount of heavy oil product and increased the amount of
gasoline and diesel oil. Thangalazhy-Gopakumar et al. (2015) reported high heating rates
supplied more thermal energy to overcome the temperature gradient inside the oily sludge
leading to an increase in the gas and liquid fractions. Wang et al. (2007) also found that more
oil was produced as the heating rate increased from 5 to 20 °C/min. They attributed this effect
to the preferable formation of free radicals at high heating rates.
Shie et al. (2004) found chemical additives (i.e., sodium and potassium compounds)
enhanced the reaction rate when the temperatures were in 377–437 °C probably due to the

33

dispersion of the inorganic additives throughout the mixture and their catalytic properties. Ma
et al. (2014) evaluated quartz sand as a solid heat carrier in the pyrolysis process of oily sludge
and they found oil yield increased from 16.63 wt% at the sludge/sand ratio of 1:0 to the
maximum yield of 28.87 wt% at sludge/sand ratio of 1:2. They suggested this could be due to
the higher heat transfer coefficient of quartz sand (i.e., 1.03 W/m·K) than that of the gas
medium (i.e., 0.25 W/m·K) (Cheng et al., 2016a; Ma et al., 2014). Studies on the addition of
oily sludge ash in the pyrolysis of oily sludge have been carried out with TGA (Shie et al.,
2004) and stirred tank reactor (Cheng et al., 2016a, 2016b). It was reported that the addition of
oily sludge ash not only decreased the carbon residue of the oil product but also increased the
light oil/heavy oil ratio. These improvements might be due to the high contents of Fe and Ca
in the oily sludge ash (Shie et al., 2004). Results from other studies also indicated that additives
of metal oxides (Al2O3, Fe2O3, and CaO), metal salts (Fe2(SO4)3H2O and AlCl3) and zeolite
catalysts (ZSM-5, Y-zeolite) affect the reaction efficiency, yield and quality of oil product
(Pánek et al., 2014; Shie et al., 2004; Wang et al., 2017a).
2.6.3. Effects of parameters on gas production
Generally, an increase in pyrolysis temperature increases gas yield, as reported by Liu et
al., (2008), Qin et al., (2015) Shen and Zhang, (2003), and Wang et al., (2007). According to
Qin et al. (2015), the increase of the gaseous products at 400–500 °C was mainly due to the
vaporization of oil, while the increase of the gaseous products at 500–600 °C was attributed to
the secondary cracking or thermal decomposition of the oils. Liu et al. (2008) found the content
of HCs and H2 increased as the temperature increased from 350 to 550 °C resulting from the
further cracking of organic compounds at higher temperatures. In contrast, the concentrations
of CO and CO2 decreased with temperature increasing from 350 to 550 °C. This might be
34

because the decomposition of carbonyl and carboxyl functional groups occurred at lower
temperatures (<450 °C) (Liu et al., 2008). In contrast, Qin et al. (2015) showed an increase in
CO concentration when the temperature was above 550 °C. This might result from the reaction
shown in eq. (2.4) occurring at high temperatures (Qin et al., 2015). Moreover, high
temperatures (>550 °C) may also be responsible for the reduction of CmHn (C2H4, C2H6, and
C3H8) (Qin et al., 2015; Wang et al., 2007). Typical reactions are shown in eq. (2.5) and (2.6)
(Leung et al., 2002).
C + H2 O → CO + H2

∆H298K =132 KJ/mol

(2.4)

C2 H6 → C2 H6 + H2

(2.5)

C2 H4 → CH4 + H2

(2.6)

Shen and Zhang (2003) observed that an increase in the gas residence time from 1.5s to 3.5
s led to an increase in the gas fraction due to secondary cracking reactions. Wang et al. (2007)
studied the changes of C1–6 hydrocarbons content utilizing an interval holding at 400 °C for 20
min. They found the concentration of C1–3 hydrocarbons had increased while the concentration
of C4–6 hydrocarbons significantly decreased regardless of the final pyrolysis temperature of
either 500 or 700 °C.
Wang et al. (2007) found the gas yield increased slightly from 15.3 to 17.8 wt% when
increasing the heating rate from 2.5 to 10 °C/min. Inguanzo et al. (2002) investigated the effect
of heating rates on the characteristics of the gas product. When the heating rate was 5 °C/min,
a maximum heating value of gas (i.e., 25 MJ/m3) was obtained at 455 °C with the maximum
release of C2H4 and C2H6. However, the maximum heating value of gas (i.e., 20 MJ/m3) at the
heating rate of 60 °C/min appeared at 600 °C.

35

Wang et al. (2007) studied the effect of adding catalyst (i.e., FCC (a fluid catalytic cracking
catalyst), NaOH, and Cu(NO3)2·3H2O) on the gas yield and composition. They found pyrolysis
of oily sludge with these three additives resulted in a higher production of gas and an increase
in the light hydrocarbons (i.e., C1–3) content, suggesting the additives were favorable to induce
the formation of free radicals. Ma et al. (2014) found gas yield increased from 2.89 wt% at the
sludge/sand ratio of 1:0 to 28.25 wt% at the sludge/sand ratio of 1:3 due to the high heat transfer
coefficient in sand. Cheng et al. (2016a) found the addition of oily sludge ash increased the H2
content from 26.62 wt% (no addition) to 34.05 wt%. The increase of H2 content could be
attributed to the catalytic effect of CaO contained in the oily sludge ash (Cheng et al., 2016b).
It was reported that the active site of CaO can promote the breakage of the C‒H bond and
intensify the dehydrogenation during pyrolysis (Yongbin et al., 2004).
2.6.4. Effects of parameters on solid production
The yield of solid from pyrolysis has been found to decrease with increasing pyrolysis
temperature (Liu et al., 2008; Qin et al., 2015; Shen and Zhang, 2003). This might be attributed
to the de-volatilization of the hydrocarbons and partial gasification of the carbonaceous
residues in the char at high temperatures (Zhang et al., 2011). Shen and Zhang (2003) observed
that an increase in the gas residue time led to a decrease in solid yields. Other authors studying
the influence of the temperature and gas residue time have found similar trends for the yield of
solids (Fonts et al., 2008). However, the heating rate was reported to have an insignificant
influence on the solid yield from the pyrolysis of oily sludge (Punnaruttanakun et al., 2003;
Wang et al., 2007).
The HHV of chars given in the existing literature varied between 7.02–34.4 MJ/kg due to
the different origins of the oily sludge and pyrolysis conditions (Liu et al., 2008; Qin et al.,
36

2015; Thangalazhy-Gopakumar et al., 2015). Differenting from the chars generated from
biomass, the sludge chars generally have a higher ash content (Karayildirim et al., 2006; Zhang
et al., 2015). This might be because most of the inorganic content such as silicon, aluminum,
and iron remained in the char while the organic content was reduced during pyrolysis of oily
sludge (Zhang et al., 2015). An HHV near to 7 MJ/kg and high ash content make them
unattractive for use in incineration. For this reason, some authors advocate the use of oily
sludge-derived char (OSDC) for adsorbent production. For example, Mohammadi and
Mirghaffari (2015) produced char from oily sludge by pyrolysis and found the BrunauerEmmett-Teller (BET) surface area of the char was relatively low (i.e., 3.63 m2/g). To improve
the surface area, they used potassium hydroxide (KOH) as an activator to produce activated
char with a much higher surface area (i.e., 327.97 m2/g). It was observed that the maximum
sorption amount for Cd2+ by the activated char (i.e., 19.23 mg/g) was higher than that of nonactivated char (i.e., 1.49 mg/g). Wang et al. (2017b) produced high quality activated char from
oily sludge by adding a solvent de-oiling step before activation, with the resulting surface area
reaching 3292 m2/g. After the de-oiling treatment, the methylene blue adsorption capacity
significantly increased from 267 mg/g to 969 mg/g.
Many efforts have been made to investigate the production of sewage sludge-derived
biochars and their sorption mechanisms (Smith et al., 2009). Jindarom et al. (2007a) developed
a potential adsorbent for dye removal by pyrolysis of sewage sludge at 750 °C under a CO2
atmosphere. The maximum adsorption capacities of acid yellow 49, basic blue 41, and reactive
red 198 dyes were reported to be 116 mg/g, 588 mg/g, and 25 mg/g, respectively and the
electrostatic interactions and dispersive interactions were deemed the dominant mechanisms
for dye adsorption. Seredych and Bandosz (2006) found the sewage sludge-derived biochar

37

obtained at 650 °C had the Cu2+ removal capacity of 63.48 mg/g, which is higher than that of
the biochar obtained at 900 °C (34.01 mg/g). They found the major sorption mechanism of
Cu2+ on biochar obtained at 650 °C was via a cation exchange reaction, leading to the release
of Mg2+, Ca2+, and Zn2+ during sorption process. However, high temperature pyrolysis at
950 °C resulted in biochar with a high degree of mineralization and therefore Cu2+ was mainly
sorbed via surface complexation and precipitation such as the formation of copper hydroxide.
Lu et al. (2012) found the biochar derived from pyrolysis of sewage sludge at 550 °C
effectively removed Pb2+ from acidic solution (pH 5) with the capacity of 30.88 mg/g. Their
results also showed that the Pb2+ sorption involved the coordination with organic carboxyl and
hydroxyl functional groups (accounting for 38.2%) and the precipitation or complexation on
mineral surface (accounting for 61.8%).
Pyrolysis temperature plays a significant role in changing biochar properties and its removal
efficiency for contaminants. For example, T. Chen et al. (2014) studied the influence of
pyrolysis temperature (500–900 °C) on properties and heavy metal adsorption ability of
sewage sludge biochar and found the best result was obtained for biochar pyrolyzed at 800 °C,
with the Cd2+ removal capacity of more than 15 mg/g. It was found that with the increasing
temperature, the ash content and microstructure development of biochar was promoted but the
concentrations of surface functional groups remained low. Generally, the biochar produced
from sewage sludge pyrolysis at low temperature was acidic whereas at high temperature it
was of a basic nature (Hossain et al., 2011; Zhang et al., 2015). Some authors have attributed
this behavior to the reduction of acidic surface groups during pyrolysis (Gascó et al., 2005). It
was found that aromatic ratios of H/C and O/C decreased with increasing pyrolysis temperature,
suggesting the reduction of surface functional groups on the biochars (T. Chen et al., 2014;

38

Zhang et al., 2015). As a result, biochar samples obtained at higher temperatures were less
polar and had greater aromaticity compared to those prepared at lower temperatures. Ahmad
et al. (2012) reported better adsorbent properties of biochar produced at higher temperature as
a consequence of the higher aromaticity and lower polarity.
During pyrolysis, the carbon content structurally rearranges and condenses to biochar with
a higher density. Further activation can promote pore development leading to a biochar which
is less dense and more porous (Lima et al., 2009). Activated char can be produced by either
physical activation or chemical activation or a combination of both. Jindarom et al. (2007a)
found the CO2 activation enhanced the formation of micropores in sewage sludge char. The
surface area was increased to 60.7 m2/g for CO2 activated char (at 750 °C), which is 1.8 times
that of non-activated char (at 750 °C). This may be attributed to the fact that CO2 gasification
could remove carbon atoms from the interior of the particle, leading to the expansion of opened
micropores or the opening up of closed micropores (Jindarom et al., 2007a). Chemical
activation of sewage sludge-derived biochars can be performed using a pyrolysis process
emplying a dehydrating reagents such as KOH (Lillo-Ródenas et al., 2008), NaOH (Zou et al.,
2013), ZnCl2 (Zhang et al., 2005), and H3PO4 (Ros et al., 2006). All these agents could enlarge
the BET surface area of biochars and eventually lead to better sorption capacity (Ros et al.,
2006; Zhang et al., 2005). Chemical activation with ZnCl2 promoted the aromatization of the
carbon skeleton, while inhibiting tar formation, thus improving the porosity of materials
(Gupta and Garg, 2015). The presence of H3PO4 during activation can increase the yield of
porous biochar by promoting depolymerisation, dehydration, and redistribution of constituent
biopolymers (Kumar and Jena, 2016). Zhang et al. (2005) found the BET surface area of H3PO4
activated biochar (289 m2/g) was about two times higher than that of non-activated biochar

39

(137 m2/g). However, Ros et al. (2006) found chemical activation with H3PO4 was ineffective
for porosity development of sewage sludge-derived biochar with BET surface areas of 6–17
m2/g. They attributed the difference to the higher activation temperature employed by Zhang
et al. (2005). Zou et al. (2013) characterized the activated biochar from sewage sludge by
chemical activation with NaOH and found the BET surface area and pore volume of activated
char was about 3 times higher than that of non-activated char (121.3 m2/g and 0.102 cm3/g) at
the same temperature (600 °C). They also found more oxygen-containing functional groups on
the surface of NaOH activated biochar facilitated hydrogen bonding with organic pollutants,
and enhanced its adsorption capacity. During the activation process, the oxidation of carbon
into carbonate and formation of intercalated compounds, such as Na2SiO3, can cause
gasification and emission of steam and carbon-dioxide, resulting in the formation of pores (Zou
et al., 2013). A scheme for NaOH activation mechanism is presented in eq. (2.4) to (2.10) (Zou
et al., 2013):
C + H2 O → CO + H2

(2.4)

CO + H2 O → CO2 + H2

(2.7)

2NaOH + CO2 → Na2 CO3 + H2 O

(2.8)

Na2 CO3 → Na2 O + CO2 or 2NaOH → Na2 O + H2 O

(2.9)

Na2 O + SiO2 →Na2 SiO3

(2.10)

2.6.5. Co-pyrolysis
A great number of studies have been reported on the investigation of the co-pyrolysis of
biomass with other materials, such as coal (Yuan et al., 2012), oily sludge (Hu et al., 2017),
waste tires (Q. Cao et al., 2009), and plastics (Önal et al., 2014). For example, Brebu et al.
(2010) studied the co-pyrolysis of pine cone with synthetic polymers and showed significant
40

synergies in pyrolysis product yields. It was observed that experimental oil yield (i.e., 52.3
wt%) from the co-pyrolysis was higher than the anticipated theoretical yield (i.e., 47.1 wt%)
calculated based on the yields from the pyrolysis of the pine cone and polystyrene alone. Such
synergetic effect in co-pyrolysis has also been reported in other studies (G. Chen et al., 2014;
Hu et al., 2017; Kar, 2011; Önal et al., 2014; Song et al., 2014). The synergetic effect may be
a result of the high H/C and O/C molar ratios in biomass which can generate a large number
of hydrogen and hydroxyl radicals. These radicals could promote the cracking reaction of
aromatic compounds within the oily sludge (Yuan et al., 2012). Meanwhile, some radicals
species such as amyl, hydroxyl, and methoxyl could combine with each other to form organic
compounds (e.g., aldehydes, carboxylic acids, and ketones), and thus increase the yield of
bottom oil (derived from biomass) (Zuo et al., 2014). Önal et al. (2014) observed that the
produced oil from co-pyrolysis with almond shells (polyethylene/almond shell = 1:1 w/w) had
higher carbon (26% higher) and hydrocarbon contents (78% higher), lower oxygen content (86%
less) and a high heating value (HHV) (38% higher) than those of biomass oil. Q. Cao et al.
(2009) found co-pyrolysis of sawdust and waste tires had the advantage of inhibiting the
formation of PAH produced by pyrolysis of the tires alone. The improvement of oil quality in
aspects of elemental composition, HHV, and content of PAHs reflected a synergistic effect
exist during co-pyrolysis of the biomass and tire (Cao et al., 2009).
Debela et al. (2012) investigated the immobilization of heavy metals in contaminated soil
by co-pyrolysis with woody biomass and found heat treatment without the addition of biomass
can be effective in immobilizing heavy metals such as Pb and Cd. Their results also showed
the additional biomass enhanced further declines in Zn and Cd leaching. Kistler et al. (1987)
observed the metals (i.e., Cr, Ni, Cu, Zn, Cd, and Pb) in the sewage sludge after pyrolysis at

41

500 °C was not mobilized by acidic leachates (50 g of char with 2070 mL of pH 3.2 HCl). This
may be attributed to the well-buffered neutral to alkaline properties of char (Kistler et al., 1987).
Devi and Saroha (2014) studied the transformation behavior of heavy metal ions (i.e., Cr, Cu,
Ni, Zn, Pb, and Cd) during the pyrolysis of paper mill effluent treatment plant sludge at
different temperatures (200–700 °C). It was observed that the mobile and bioavailable heavy
metal ion fractions were transformed into relatively stable fractions after pyrolysis and most
of the heavy metal ions were found to exist in the residual form in the biochar pyrolyzed at
700 °C (Devi and Saroha, 2014). Similarly, Islam and Park, (2017) found the hydrothermal
treatment (270 °C) significantly decreased the bioavailable fraction of Pb in contaminated soil
from 41.33 to 14.66 %, while the Pb content in the non-bio-available fraction increased from
2.90 to 15.76%. They explained Pb immobilization may result from mechanisms such as cation
exchange and precipitation.
Shi et al. (2013b) reported that the addition of rice husk during the hydrothermal treatment
of sewage sludge exhibited a reduction effect on the leaching toxicity of heavy metals in sludge
residues. It was observed that at the rice husk/sludge ratio of 1/1.75, the concentrations of
leachable Cr, Ni, Cu, and Cd decreased by 17%, 89%, 95%, and 93% respectively. More
importantly, addition of rice husks led to further transformation of the metals within residual
fractions. They postulated that a biomass-sewage sludge matrix was formed during
hydrothermal treatment with rice husk addition in which the metals were further entrapped by
the enlarged structure of rice husk or chelated with some functional groups, resulting in an
enhanced immobilization effect on the metals. A similar cumulative effect of biomass addition
on lowering the contamination risk of heavy metals in biochar has been reported by other
authors (H. Chen et al., 2014; Jin et al., 2017).

42

2.7. Summary of literature review
Oily sludge is one of the most significant solid waste generated in petroleum industry
during the process of drilling, production, transport, storage, and refining (da Silva et al., 2012).
Such sludge usually contains various pollutants, such as heavy metals (da Rocha et al., 2010),
PAHs (Kriipsalu et al., 2008), and BTEX (Kriipsalu et al., 2008), posing risks to human health
and the environment (Kuang et al., 2011; Robertson et al., 2007). The effective treatment of
oily sludge has attracted widespread attention due to its hazardous and recalcitrant nature (Hu
et al., 2013). A variety of remediation technologies are available for oily sludge treatment
(Table 2.2). However, no single technology is universally applicable because properties of oily
sludge from different sources vary. Moreover, each method is associated with different
advantages and disadvantages. Some treatments are effective in removing PHCs but are very
expensive. Other treatments have relatively low cost but require long treatment duration and/or
fail to meet the requirements for the reduction of large molecular weight species and metal
ions in the oily sludge. Future research should focus on the development of new techniques for
recycling and treating oily sludge with high efficiency and low costs.
Solvent extraction is a simple and commonly used approach to recycle petroleum
hydrocarbons from oily sludge. Extraction performance depends on solvent type, temperature,
solvent-to-sludge ratio, extraction duration, and stirring rate. The utilization of solvent is
essential to solvent extraction. However, it also limits the extension of this technology to fieldscale oily sludge treatment due to the high cost of solvent recovery and toxicity of the organic
solvent. In the past years, growing interest has been shown in the application of ILs to recover
bitumen from oil sands (Painter et al., 2010b; Sui et al., 2016; Tourvieille et al., 2017). The
literature suggests that ILs are potentially good candidates for the enhancement of the
43

liberation of oil from solid matrices during solvent extraction. However, limited studies have
been reported on the use of ILs in environmental remediation, especially oil recovery from oily
sludge.
Pyrolysis is a promising strategic approach to the sustainable management of oily sludge,
having the ability to not only minimize the solid waste, but also to yield valuable products (i.e.,
liquid, gas, and char). A number of operating variables may affect the yield and composition
of the liquid and gas product, such as pyrolysis temperature, reaction time, heating rate, and
chemical additives (Qin et al., 2015; Shie et al., 2003; Wang et al., 2007). Due to the different
origins of the oily sludge and pyrolysis conditions, the HHV of chars can range from 7.02
MJ/kg to 34.4 MJ/kg (Liu et al., 2008; Qin et al., 2015; Thangalazhy-Gopakumar et al., 2015).
An HHV near to 7 MJ/kg and high ash content make chars unattractive for use in incineration.
Therefore, some authors advocate the use of oily sludge-derived char (OSDC) for adsorbent
production (Mohammadi and Mirghaffari, 2015; Wang et al., 2017b). However, there is a lack
of information on the roles of different sorption mechanisms for OSDC. Pyrolysis conditions
such as temperature and chemical activation play important roles in changing biochar
physiochemical properties and its removal efficiency of contaminants (Ahmad et al., 2012; T.
Chen et al., 2014; Zhang et al., 2005; Zou et al., 2013). However, very limited research tried
to establish the relationship among properties of OSDC, metal sorption behavior, and pyrolysis
conditions. Therefore, investigating the exact mechanisms for Pb2+ and Cd2+ sorption on OSDC
prepared at different pyrolysis temperature and activation methods is also necessary. The
addition of biomass during pyrolysis of oily sludge resulted in a synergistic effect on oil yield
(Hu et al., 2017). Some of the oil properties such as HHV and carbon content were found to be
improved by co-pyrolysis (Q. Cao et al., 2009). Some efforts have been made to investigate

44

the cumulative effect of biomass addition on lowering the contamination risk of heavy metals
in sewage sludge residues (H. Chen et al., 2014; Jin et al., 2017; Shi et al., 2013b). However,
there has been limited research investigating co-pyrolysis effect of oily sludge and sawdust on
heavy metal immobilization.

45

Chapter 3
IONIC LIQUID-ENHANCED SOLVENT EXTRACTION FOR OIL RECOVERY
FROM OILY SLUDGE1
Abstract
An ionic liquid (IL), 1-ethyl-3-methyl-imidazolium tetrafluoroborate ([Emim][BF4]), was
investigated for its effectiveness on enhancing oil recovery from hazardous crude oil tank
bottom sludge using solvent. A range of solvent (cyclohexane)/sludge ratio (2–8 mL/g),
shaking speed (100–400 rpm), extraction duration (10–120 min), and IL/sludge ratio (0.25–1.0
mL/g), were examined. The addition of IL (i.e., 1 mL/g of IL/sludge ratio) increased total
petroleum hydrocarbon (TPH) recovery by 9.4% (from 84.4 ± 2.4% to 93.8 ± 2.3% at the
solvent/sludge ratio of 8 mL/g). An orthogonal experimental design was subsequently applied
to optimize extraction conditions. The use of IL (i.e., 0.1 mL/g of IL/sludge ratio) with solvent
yielded higher TPH recovery (above 95%) at shorter extraction duration (10 min), lower
solvent/sludge ratio (4:5 mL/g), and lower energy consumption (100 rpm). The recovered oil
had similar calorific value but a higher F3 fraction compared with crude oil. The results
suggested the IL-enhanced solvent extraction with lower solvent consumption is an effective
approach for oily sludge treatment.
Keywords: oily sludge, ionic liquid, solvent extraction, oil recovery, total petroleum
hydrocarbon

This work has been published as Tian, Y., McGill, W.B., Whitcombe, T.W., Li, J., 2019. Ionic
liquid-enhanced solvent extraction for oil recovery from oily sludge. Energy Fuels 33, 3429–3438.
DOI: 10.1021/acs.energyfuels.9b00224.
1

46

3.1. Introduction
The petroleum industry produces a large quantity of oily sludge during its oil exploration
and processing activities (da Silva et al., 2012; Xu et al., 2009). Such sludge is a complex
water/oil emulsion typically including 30–50% oils, 10–20% solids and 30–50% water by
weight (Ramaswamy et al., 2007). It also contains some aromatic hydrocarbons (e.g., benzene)
and polycyclic aromatic hydrocarbons (PAHs) that are mutagenic and carcinogenic (IARC
WGECRH, 2010, 1987). Therefore, oily sludge has been classified as a hazardous waste in
many countries. Its improper treatment can pose a serious threat to the environment and
humans (Hejazi et al., 2003; Kuang et al., 2011; Robertson et al., 2007). Most conventional
methods for dealing with oily sludge are landfills or incineration. Due to stringent regulations
in recent years, landfilling is being gradually prohibited or restricted (da Silva et al., 2012;
Elektorowicz and Habibi, 2005). Although incineration reduces a large volume of oily sludge,
it requires a substantial amount of auxiliary fuel and generates a large volume of off-gases
containing PAHs (C.-T. Li et al., 1995). Recycling is an alternative option to handle oily sludge
by providing the possibility for reusing valuable oil for reprocessing and recovering energy
(Elektorowicz and Habibi, 2005; Hu et al., 2013). Solvent extraction is a commonly used
approach to recycle hydrocarbons from oily waste by mixing with a desirable amount of
solvent (Hu et al., 2016). Many studies reported that a variety of variables could affect oil
recovery in solvent extraction processes (e.g., temperature, pressure, solvent-to-waste ratio,
mixing, and the property of solvent) (Fisher et al., 1997; Taiwo and Otolorin, 2009; Zubaidy
and Abouelnasr, 2010). The application of solvent extraction to oily sludge treatment is
desirable, but challenged by its unsatisfactory extraction efficiency, large volume of solvent
consumption, and long extraction duration (Hu et al., 2013; Li et al., 2011).
47

Ionic liquid (IL) is purely ionic material with a melting temperature below 100 °C. It has
been considered as “green and eco-friendly solvent” because of its unique properties (e.g.,
thermal and chemical stability) (José-Alberto and Jorge, 2011). In the past few years, growing
interest has been shown in the application of ionic liquid (IL) to recover bitumen from oil sands
(Hogshead et al., 2011; Painter et al., 2010b; Sui et al., 2016; Tourvieille et al., 2017). Painter
et al. (2010a) demonstrated that more than 90% (by weight) of bitumen was recovered from
oil sands when using an imidazolium-based IL and toluene, while the use of IL could allow for
a cleaner separation. Li et al. (2011) investigated the effect of IL on bitumen recovery from oil
sands using a composite solvent (acetone and n-heptane), and found that the bitumen recovery
was increased by up to 10% (by weight) by adding IL (i.e., 1.3 mL/g of IL/oil sands ratio). A
recent study by Sui et al. (2016) found that bitumen recovery from oil sands using IL-enhanced
solvent extraction increased with IL-to-oil sands ratio and agitation time. As a result, IL can
be potentially effective candidates for enhancing the liberation of oil from solid matrices during
solvent extraction. However, few studies have been reported on the use of IL in oily sludge
treatment. The objective of this study was then to propose an IL-enhanced solvent extraction
method for oil recovery from oily sludge. Several factors including solvent/sludge ratio,
shaking speed, extraction duration, and IL/sludge ratio were examined for their impacts on oil
recovery. The optimal conditions of oil recovery for two processes (i.e. solvent extraction and
IL-enhanced solvent extraction) were also investigated.

48

3.2. Materials and methods
3.2.1. Materials and reagents
The oily sludge used in this study was obtained from a crude oil tank bottom in an oil
refinery plant in western Canada. The appearance of the oily sludge was black and sticky.
Table 3.1 lists its characteristics. The total petroleum hydrocarbon (TPH) extraction followed
the US EPA 3540C method (USEPA, 1996a). The water content was analyzed based on the
ASTM D2974-00 Method A (ASTM, 2000), and the solid content was calculated according to
the measured TPH and water contents. Dichloromethane (DCM) and cyclohexane used for
sample extraction were of HPLC grade (>97%, VWR International). Previous studies on ILenhanced bitumen extraction from oil sands showed that among various ionic liquids, the use
of 1-ethyl-3-methyl-imidazolium tetrafluoroborate ([Emim][BF4]) and 1-butyl-2,3-dimethylimidazolium tetrafluoroborate ([Bmmim][BF4]) led to high-efficiency IL-free bitumen
recovery (Li et al., 2011; Painter et al., 2010b, 2010a; Sui et al., 2016). Moreover, [Emim][BF4]
has a much less viscosity (0.06 Pa·s) than that of [Bmmim][BF4] (0.46 Pa·s at 298 K), which
allows an easier industrial operation (i.e., no equipment blockage). Therefore, [Emim][BF4] at
HPLC-grade (≥98%)) purchased from Sigma-Aldrich was used as the ionic liquid (IL) to be
investigated in this study. Silica gel (70‒230 mesh, VWR international) activated at 105 °C for
12 h was used to clean up the extraction solution. Anhydrous sodium sulfate (VWR
international) dried at 400 °C for 4 h was used to remove moisture in the extract.

49

Table 3.1. Characteristics of the oily sludge

Parameter

Concentration

TPH

58.7 ± 4 % (by mass)

Water content*

22.0 ± 5 % (by mass)

Solid content

19.3 ± 1 % (by mass)

on a wet-weight basis

*

3.2.2. Single factor experiments
Single factor experiments were designed to evaluate the influence of IL ([Emim][BF4])
addition on the oil recovery by solvent extraction with cyclohexane. Among various solvents,
cyclohexane was reported to have a high bitumen recovery since it has a medium solubility
(i.e., 16.8 MPa1/2) which is close to that of asphaltene (i.e., 20.1 MPa1/2) (Wang et al., 2014).
Additionally, cyclohexane has a low boiling point (i.e., 80.7 °C) which allows easy solvent
recovery, and less toxicity than many other solvents. Our previous study (Hu et al., 2015) also
indicated that cyclohexane could achieve a high oil recovery when treating petroleum refinery
wastewater treatment pond sludge. Therefore, cyclohexane was selected as the solvent for
investigation in this study. A total of four factors were examined, including solvent/sludge
(So/Sl) ratio, shaking speed, extraction duration, and IL/sludge (IL/Sl) ratio. It was reported in
our previous study and other authors’ work that the extraction performance could be
significantly affected by these factors (Hu et al., 2015, 2016; X. Li et al., 2012; Sui et al., 2016).
The levels of experimental factors are presented in Table 3.2.

50

Table 3.2. Single-factor experimental conditions

Number

So/Sl ratio

Shaking speed

(mL/g)

(rpm)

Extraction
duration (min)

IL/Sl ratio
(mL/g)

1

2:1, 4:1, 6:1, 8:1 200

30

0

2

2:1, 4:1, 6:1, 8:1 200

30

1:1

3

2:1

100, 200, 300, 400

30

0

4

2:1

100, 200, 300, 400

30

1:1

5

2:1

200

10, 30, 60, 120

0

6

2:1

200

10, 30, 60, 120

1:1

7

2:1

200

30

1:4, 2:4, 3:4, 4:4

Solvent extraction using cyclohexane alone was conducted by adding the desired volume
of cyclohexane (Table 3.2) to a 15-mL centrifuge tube containing oily sludge (0.5 g). The
mixture was then shaken under the designed condition (Table 3.2) by using a Talboy 3500
Orbital shaker, then centrifuged for 8 min at 4704g. The extract cleanup and quantification
followed the method of Zhang et al. (2012) and CCME (CCME, 2001). The solvent in the
cleaned extract was evaporated in a rotary evaporator. The residue was weighed and designated
as recovered TPH (TPHr). The TPH recovery rate was calculated as follows:
R = TPHr/(Co×Mo)×100%

(3.1)

where 𝑅 is the TPH recovery rate (%); TPHr (mg) is the mass of recovered TPH; Co (mg/g) is
the TPH concentrations in the oily sludge sample (Co = 5.87 × 105 mg/kg, as listed in Table
3.1); and Mo (g) is the mass of oily sludge sample.
In terms of solvent extraction with IL addition, preliminary experiments showed that the IL
layer stayed between the top oil/solvent layer and the sand/clay layer after mixing oily sludge
51

with IL and solvent. At the oil/solvent-IL interface, a black film containing suspended oil and
some mineral fines was formed as shown in Figure 3.1. Similar phenomenon has also been
reported in other studies (Sui et al., 2016; Tourvieille et al., 2017). Given this observation, the
recovered TPH in IL-enhanced solvent extraction treatment was calculated as follows:
TPHr = TPHrt +TPHri

(3.2)

where TPHr (mg) is the mass of recovered TPH; TPHrt (mg) is mass of recovered TPH from
top oil/solvent layer; and TPHri (mg) is mass of recovered TPH from oil/solvent-IL interface.

Solvent/oil
mixture
Film at the
interface

Solvent/oil
mixture

Ionic liquid
Residual
sand/clay

a

b

Residual
sand/clay

Figure 3.1. The phases formed by mixing, (a) oily sludge with [Emim][BF4] and cyclohexane in the
proportion of 1:3:2 (by weight), and (b) oily sludge with cyclohexane in the proportion of 1:2 (by
weight).

In IL-enhanced solvent extraction treatment, a sample of oily sludge, IL, and cyclohexane
in proportion (Table 3.2) was placed into a 15-mL centrifuge tube. Extraction, centrifugation,
cleanup and evaporation for determining the recovered TPH mass from top oil/solvent layer
(TPHrt) were similar as described above. After separation of the top oil/solvent layer from the
tube, a small amount of fresh cyclohexane solvent (500 µL) was added and then gently mixed

52

with the oil/solvent-IL interface. The mixture was centrifuged at 4704g for 3 min, and the
solvent/oil phase was decanted and saved. The procedure was repeated until the supernatant
was colorless. The mass of recovered TPH from oil/solvent-IL interface (TPHri) was then
obtained.
For both the cyclohexane extraction and IL-enhanced cyclohexane extraction, each was
performed in triplicate and the results were expressed as mean ± standard deviation. One way
analysis of variance (ANOVA) with Tukey’s analysis was used to determine the significant (P
< 0.05) differences between levels of treatment.
3.2.3. Optimization of solvent extraction and IL-enhanced solvent extraction
The optimal operational condition for each extraction process, including solvent extraction
and IL-enhanced solvent extraction, was examined using a L9 (34) orthogonal experimental
design (Table 3.3) (Taguchi et al., 2005). The factors and levels of each factor were chosen
depending on the results of single factor experiments. The extraction duration was set as a
constant (i.e. 10 min) for IL-enhanced solvent extraction, and all IL treatments were conducted
in a 2-mL centrifuge tube. Each experiment was performed in triplicate.

53

Table 3.3. Orthogonal array L9 (34) experimental design and factor levels

Extraction
process

Factors

Symbol

Levels
1

2

3

So/Sl ratio (mL/g)

A

4:1

8:1

12:1

Solvent
extraction

Shaking speed (rpm)

B

300

400

500

Extraction duration (min)

C

30

60

120

ILenhanced
solvent
extraction

So/Sl ratio (mL/g)

A

4:5

7:5

10:5

IL/Sl ratio (mL/g)

B

1:10

2:10

3:10

Shaking speed (rpm)

C

100

200

300

Experimental No.

Experimental factors and levels
A

B

C

E*

1

1

1

1

1

2

1

2

2

2

3

1

3

3

3

4

2

1

2

3

5

2

2

3

1

6

2

3

1

2

7

3

1

3

2

8

3

2

1

3

9

3

3

2

1

*

Empty column

The TPH recovery rate from each treatment was calculated by eq. (3.1). Additionally, the
recovered TPH was transferred into a 10 mL volumetric flask, and then n-hexane was added
to bring the final volume to 10 mL. The extract samples were stored at 4 °C for further analysis.
In order to calculate the mass balance of TPH for each orthogonal experiment, the TPH mass

54

in the residual sand/clay after solvent extraction or IL-enhanced solvent extraction was
determined. In this study, 10 mL of DCM was added into a 15-mL centrifuge tube containing
residual sand/clay at the bottom. The tube was shaken on a Talboy 3500 Orbital shaker at 300
rpm for 30 min and then centrifuged at 4704g for 10 min (Schwab et al., 1999). The extraction
solution was decanted and kept in a 40 mL glass vial. Another 10 mL clean DCM was added
to the same centrifugation tube, and the extraction procedure was repeated for two more times.
A total of 30 mL of extraction solution was then collected and used for silica gel cleanup. After
evaporation, the TPH extracted from residual sand/clay was brought to 10 mL with n-hexane
and stored at 4 °C until analysis.
Signal to noise (S/N) ratio which combines the mean and variance was used to evaluate the
experimental data. Since this study aims at maximizing the TPH recovery rate (%), the S/N
ratio (dB) was calculated by the following equation:
S
N

1

1

= − 10 log {n ∑ni=1 y 2}
i

(3.3)

where yi denotes the TPH recovery observed in the ith repetition of the experimental run; n is
the total number of repetitions under the same experimental condition (Hedayatipour et al.,
2017; Taguchi et al., 2005; Zolfaghari et al., 2011). In addition to S/N ratio, ANOVA was used
to test the significance of independent variables.
3.2.4. Laboratory analysis
3.2.4.1. Hydrocarbon analysis using gas chromatograph
Chromatographic characterization of samples with petroleum hydrocarbons (PHCs) was
carried out using an Agilent 6890 Gas Chromatograph with a flame ionization detector (GCFID). The hydrocarbon mixtures were grouped into four PHC fractions (CCME, 2001): F1
55

(nC6–nC10), F2 (nC10–nC16), F3 (nC16–nC34), and F4 (nC35–nC50). Decane (nC10),
hexadecane (nC16), and tetratriacontane (nC34) were used as the external standards to
determine the average response factor and then quantify the PHC fractions. The capillary
column used was a ZB-1HT INFERNO column (Phenomenex) with a length of 30 m, an inside
diameter of 0.32 mm and a film thickness of 0.25 µm. Helium was used as the carrier gas at a
rate of 2.1 mL/min. TPH extract (1 µL) was injected into the GC system and a split ratio of
10:1 was performed for each run. Both the injector and detector temperatures were kept at 360
°C during analysis. The initial temperature of the oven was held at 40 °C for 4 min, then
increased to 140 °C at 10.0 °C/min, and further increased to 340 °C at 20.0 °C/min and held at
340 °C for 11 min. The total run time was 35 min.
3.2.4.2. Characterization of recovered oil and residual sand/clay
The total carbon (C), nitrogen (N), hydrogen (H), and sulfur (S) of oil were determined
using a Costech ECS 4010 elemental analyzer according to ASTM D5291-16 (ASTM, 2016).
The calorific values of recovered oil and crude oil (donated by an oil refinery in western
Canada) were determined by using a bomb calorimeter (Model: 1341 Plain Jacket Calorimeter,
Parr Instrument Company) according to ASTM D2015-85 (ASTM, 1985). The elements (e.g.,
Fe, Cr, Pb, and Ni) in recovered oil and residual sand/clay were determined by digestion
(USEPA 3050B) (USEPA, 1996b) and then quantified by inductively coupled plasma optical
emission spectrometry (ICP-OES) (Agilent Technologies 5100) according to ASTM D5185
(ASTM, 2018). The residual sand/clay samples were ground and mixed with KBr to 0.1 wt%
and then pressed into pellets. Thin films of recovered oil were prepared by placing a DCM
solution containing the oil on a KBr window and allowing DCM to evaporate (Painter et al.,

56

2010a). The Fourier transform infrared (FTIR) spectra were then obtained using a Nexus 670
Thermo Nicolet Spectrometer.
3.3. Results and discussion
3.3.1. Single factor experiments
As seen from Figure 3.2a–c, the extraction conditions, including solvent/sludge ratio,
shaking speed, and extraction duration, had significant (P < 0.05) effects on the performance
of solvent extraction alone. The TPH recovery increased from 68.7 ± 0.8% to 84.4 ± 2.4%
when the solvent/sludge ratio increased from 2:1 to 8:1 mL/g. Such TPH recovery increase
could be owing to an increase in the amount of oil dissolved in solvent when increasing
solvent/sludge ratio (Elbashir et al., 2002). Other authors using solvent extraction to treat oily
sludge (Hu et al., 2015; Zubaidy and Abouelnasr, 2010) and used lubricant oil (Elbashir et al.,
2002; Rincón et al., 2005) found that extraction yield remained stable when the solvent/sludge
ratio was increased beyond a certain point. However, the stabilization was not observed in this
study because even at the highest solvent/sludge ratio (i.e. 8:1), the solvent was not sufficient
to dissolve all oil retained by solid matrices. According to Figure 3.2b, the TPH recovery went
up gradually from 68.3 ± 3.9% at 100 rpm to 75.5 ± 1.5% at 400 rpm. This might be because
the oil initially adsorbed on the sand/clay particles was stripped and then driven into the solvent
phase by increasing stirring strength (X. Li et al., 2012). As shown in Figure 3.2c, the TPH
recovery also experienced an upward trend with the extension of extraction duration (70.5 ±
1.3% and 76.1 ± 2.1% at 10 min and 120 min, respectively). With longer extraction time, the
interaction between oil and solvent was enhanced, thereby increasing the amount of recovered
TPH.

57

a
A

A

A

TPH recovery (%)

90

70

c
Solvent extraction
IL enhanced solvent extraction

60
50

90

a

ab

b

80

100

A

TPH recovery (%)

100

2

4

6

b

70

50

b

ab

Solvent Extraction
IL enhanced solvent extraction

100

200

300

400

Shaking speed (rpm)

c

100
A

90

TPH recovery (%)

A

a
b

60

8

A

80

So/Sl ratio (ml/g)

A

A

A

80

a
b

70

ab
b

Solvent extraction
IL enhanced solvent extraction

60
50

A

A

0

30

60

90

120

Extraction duration (min)

Figure 3.2. Influence of different operational parameters on TPH recovery: (a) So/Sl ratio (shaking
speed: 200 rpm, extraction duration: 30 min, IL/Sl ratio: 1:1 mL/g), (b) shaking speed (So/Sl ratio: 2:1
mL/g, extraction duration: 30 min, IL/Sl ratio: 1:1 mL/g), and (c) extraction duration (So/Sl ratio: 2:1
mL/g, shaking speed: 200 rpm, IL/Sl ratio: 1:1 mL/g). Significant (P < 0.05) differences are shown by
different letters: lowercase letter for the solvent extraction, and uppercase letter for the IL-enhanced
solvent extraction.

However, no significant differences (P > 0.05) among the variables (solvent/sludge ratio,
shaking time, and extraction duration) were observed on TPH recovery when IL was used. The
TPH recovery was in the range from 88.7 ± 2.8% to 93.8 ± 2.3%, indicating that an extra
58

recovery of 7.1‒21.5% was achieved as compared to that of solvent extraction alone. The TPH
recovery was substantially facilitated with the addition of IL probably due to two factors. First,
the adhesion forces between oil and silica were about an order of magnitude smaller in an IL
medium than in an aqueous solution, which may be attributed to the fact that IL can form
layered ion/charge structures on mineral surfaces (Hogshead et al., 2011). Second, the
separation was driven by the solubility characteristics of hydrophilic IL. The IL used in this
study was immiscible with hydrocarbons, and thus an intermediate phase separating minerals
from the hydrocarbon layer could be formed by IL (Painter et al., 2010a).
Considering the high cost of IL, a set of experiments with small IL/sludge ratios were
conducted. The observed TPH recovery is shown in Figure 3.3. The TPH recovery increased
from 68.7 ± 0.8% to 90.6 ± 1.3% as the IL/sludge ratio increased from 0 to 0.25 mL/g.
However, a further increase in IL/sludge ratio didn’t lead to significant enhancement of TPH
recovery. It was reported that even a low concentration of IL could reduce the viscosity of
crude oil by up to 35%, thus greatly facilitating solvent extraction efficiency (Subramanian et
al., 2015). Sakthivel et al. (2014) also found that only a low IL/sludge ratio was required for a
complete dissolution of oily sludge in solvents. The results suggest that the desirable IL/sludge
ratio sufficient to recover most oil from oily sludge could be very low, but requires further
investigation. In this regard, further optimization on oil recovery was carried out using
IL/sludge ratios in the range from 0.1 to 0.3 mL/g.

59

100

TPH recovery (%)

a

a

a

90

a

80

70

60

b

0.00

0.25

0.50

0.75

1.00

IL/Sl ratio (ml/g)

Figure 3.3. Influence of IL/sludge (IL/Sl) ratio on TPH recovery (So/Sl ratio: 2:1 mL/g, shaking speed:
200 rpm, extraction duration: 30 min). Different lowercase letters (a,b) indicates significant (P < 0.05)
differences.

3.3.2. Multifactor orthogonal experiments
Mass balance calculation of the F2, F3, and F4 fractions of TPH for orthogonal experiments
in solvent extraction alone and IL-enhanced solvent extraction are presented in Tables A1 and
A2 in the Appendix A, respectively. A good agreement between the TPHo (TPH in original
oily sludge) and TPHs (the sum of TPH in recovered oil and TPH in residual sand/clay)
indicated that the mass loss of TPH in each treatment was small and could be ignored in this
study.
3.3.2.1. Solvent extraction alone
Table 3.4 shows the results of orthogonal experiments using solvent extraction alone. The
TPH recovery was between 80.66 ± 2.42% and 95.49 ± 4.52%. The highest TPH recovery
(95.49 ± 4.52% in experiment # 9) was observed under the following conditions: 12:1 mL/g of

60

solvent/sludge ratio, 60 min of extraction duration, and shaking speed of 500 rpm. It is noted
that the F4 recovery rate for experiment # 7 and 9 was slightly over 100 %. This is because the
F4 recovery rate was calculated using the average of mass of F4 fraction in the original oily
sludge sample and recovered oil sample (F4o and F4t, as seen in Table A1 in the Appendix A)
without the consideration of variance.
Table 3.4. Results of the orthogonal experiments for solvent extraction alone

Exp.
No.

So/Sl ratio
(mL/g)

Shaking speed Extraction
(rpm)
duration (min)

Recovery rate (%)

1

4:1

300

30

80.66 90.01 78.09 65.21

2

4:1

400

60

84.37 95.41 79.06 74.78

3

4:1

500

120

82.87 86.44 80.82 81.05

4

8:1

300

60

88.65 96.46 83.53 86.99

5

8:1

400

120

89.23 93.65 84.72 94.39

6

8:1

500

30

86.60 87.92 83.74 93.85

7

12:1

300

120

93.88 94.43 91.33 102.05

8

12:1

400

30

89.98 91.95 86.65 97.28

9

12:1

500

60

95.49 99.62 90.61 102.86

61

TPH

F2

F3

F4

Figure 3.4a shows the main effects of three factors on the TPH recovery from solvent
extraction alone. It can be found that the S/N ratio increased rapidly with the increasing
solvent/sludge ratio (from 4:1 to 12:1 mL/g), peaking at 39.36 dB. The trend was in accordance
with the results for single factor experiments, suggesting that solvent/sludge ratio played a
significant role in extracting oil from oily sludge. Further increase of solvent/sludge ratio could
be beneficial to TPH recovery to some extent. However, the high solvent/sludge ratio means a
high cost of operation, thus 12:1 mL/g of solvent/sludge ratio was chosen as the optimal
condition in this study. According to the ANOVA analysis shown in Table A3 in the Appendix
A, there is no significant effect (F-value = 0.11) on TPH recovery by increasing the shaking
speed from 300 rpm to 500 rpm, indicating 300 rpm induced adequate turbulence for mixing
the oil and solvent in extraction. Moreover, the S/N ratio was maximized at 60 min of
extraction duration then followed by a moderate drop from 60 min to 120 min. By contrast,
there was a steady rise in TPH recovery from 60 min to 120 min in single factor experiment
(Figure 3.2c). This may be explained by the interaction effect between extraction duration with
other factors, especially with solvent/sludge ratio. Table A3 in the Appendix A shows the most
significant factor affecting TPH recovery was solvent/sludge ratio (F-value = 30.21) followed
by extraction duration (F-value = 4.28). The optimum condition for TPH recovery from oily
sludge using solvent extraction alone would be 12:1 mL/g of solvent/sludge ratio, 60 min of
extraction duration, and shaking speed of 500 rpm.

62

1

3

1

2

3

1

2

1

3

42.0

a

Mean of S/N ratios (dB)

Mean of S/N ratios (dB)

42.0

2

40.5

39.0

37.5

1

2

3

1

2

3

b

40.5

39.0

37.5

2

3

1

2

3

1

2

3

1
42.0

c

Mean of S/N ratios (dB)

1

So/Sl ratio

Shaking speed Extraction duration

So/Sl ratio

Mean of S/N ratios (dB)

3

36.0

36.0

42.0

2

40.5

39.0

37.5

36.0

2

Shaking speed Extraction duration

3

1

2

3

1

2

3

d

40.5

39.0

37.5

36.0

So/Sl ratio

So/Sl ratio

Shaking speed Extraction duration

Shaking speed Extraction duration

Figure 3.4. Effect of independent factors on (a) TPH recovery, (b) F2 recovery, (c) F3 recovery, and
(d) F4 recovery of the solvent extraction alone process.

As seen from Figure 3.4b–d and Table A3 in the Appendix A, the F4 recovery was
significantly affected by all variables, whereas the solvent/sludge ratio and extraction duration
were two significant factors impacting F3 recovery. However, no significant factor was
observed for F2 recovery. As compared with long-chain or aromatic hydrocarbons, short-chain
hydrocarbons can be more easily desorbed from sand/clay due to their lower intermolecular
forces (Feng and Aldrich, 2000). As a result, a high F2 recovery (i.e. 90.01 ± 2.27%) was
obtained when the variables were at level 1 and very few, if any, contribution could be seen in
F2 recovery by increasing the variables from level 1 to level 3. As seen in Figure 3.4b, the F2
recovery slightly increased with an increase in shaking speed up to 400 rpm, after which it
63

decreased with shaking speed. Similar changes with the extension of extraction duration
occurred. Taking into consideration the fact that the average F2 recovery at 400 rpm of shaking
speed (i.e., level 2) and 60 min of extraction duration (i.e., level 2) was as high as 93.67 ±
1.73% and 97.16 ± 2.19%, respectively, the re-adsorption of dissolved F2 onto the solid
particles surfaces could be responsible for the decline of F2 recovery with further increase in
shaking speed (above 400 rpm) or extraction duration (above 60 min) (Feng and Aldrich,
2000).
2.3.2.2. IL-enhanced solvent extraction
The TPH recovery from oily sludge using IL-enhanced solvent extraction was between
91.08 ± 5.34% and 98.33 ± 3.57% and the average TPH recovery was 95.80 ± 4.62% (Table
3.5). As shown in Figure 3.5, the S/N ratio for TPH recovery fluctuated between 39.16 dB and
39.84 dB. It is interesting to note that in IL-enhanced solvent extraction process, all variables
including solvent/sludge ratio, IL/sludge ratio, and shaking speed had an insignificant
influence on the TPH recovery (Table A4 in the Appendix A). The above results indicated that
with the addition of IL, most TPH could be extracted from oily sludge under very mild
conditions (e.g., 10 min extraction duration and 100 rpm of shaking speed). This might be due
to the IL addition reducing the viscosity of oily sludge, leading to the enhancement of oil
liberation from solid matrices during solvent extraction. By considering the operational cost
and time, the optimum condition for TPH recovery using IL-enhanced solvent extraction was
chosen in this study as 4:5 mL/g of solvent/sludge ratio, 1:10 mL/g of IL/sludge ratio, and 100
rpm of shaking speed.

64

Table 3.5. Results of the orthogonal experiments for IL-enhanced solvent extraction

Exp.
No.

So/Sl
ratio
(mL/g)

IL/Sl
ratio
(mL/g)

Shaking
speed
(rpm)

TPH Recovery rate (%)
Total

Top
oil/solvent
layer

Oil/solventIL interface

1

4:5

1:10

100

96.92

67.59

29.33

2

4:5

2:10

200

98.33

72.95

25.39

3

4:5

3:10

300

91.08

70.72

20.35

4

7:5

1:10

200

97.52

75.21

22.31

5

7:5

2:10

300

94.85

79.02

15.84

6

7:5

3:10

100

95.73

70.32

25.41

7

10:5

1:10

300

97.24

80.00

17.25

8

10:5

2:10

100

97.63

75.63

22.00

9

10:5

3:10

200

92.91

71.06

21.84

Total

Mean of S/N ratios (dB)

1

Top oil layer
2

3

1

Oil-IL interface
2

3

1

2

3

40
36
32
28
24

So/Sl ratio

IL/Sl ratio

Shaking speed

Figure 3.5. Effect of independent factors on TPH recovery of the IL-enhanced solvent extraction
process.
65

In general, a higher quantity of TPH was collected from the top oil/solvent layer when
compared with that from oil/solvent-IL interface (Table 3.5). Although the differences in TPH
recovery were not statistically significant, the influence of solvent/sludge ratio and shaking
speed on TPH recovery from the oil/solvent-IL interface was highly significant (P < 0.01),
which can be observed from Table A4 in the Appendix A. As seen from Figure 3.5, the increase
of solvent/sludge ratio reduced the S/N ratio from 27.76 to 25.82 dB for TPH recovery from
the oil/solvent-IL interface. Correspondingly, the S/N ratio for TPH recovery from the top
oil/solvent layer increased from 36.91 to 37.54 dB with the increase of solvent/sludge ratio.
Furthermore, increasing the shaking speed from 100 rpm to 300 rpm decreased the S/N ratio
from 27.98 to 24.71 dB for TPH recovery from the oil/solvent-IL interface. These results
implied that even though the break of incrustation and desorption of oil from solid matrices
were high at the low levels (i.e. level 1) of all the three experimental variables, the dissolution
of oil into the solvent was not fully complete, leaving the undissolved oil (containing 29.33%
TPH) in the oil/solvent-IL interface. However, higher solvent/sludge ratio and stronger
shearing force could enhance the dissolution of suspended oil into the solvent.
Table 3.5 shows that a high TPH recovery (> 95%) could be achieved even when the
IL/sludge ratio was as low as 0.1 mL/g. However, the high cost of IL can still be a barrier to
the application of IL-enhanced solvent extraction. For example, the cost of 1 kg of [Emim][BF4]
can range from $340 to $1300 USD depending on its purity. As a result, the reuse of IL is of
great importance. Plechkova and Seddon (2008) estimated that more than 10–20 times of IL
reuse would make the IL-enhanced solvent extraction less expensive than solvent extraction
alone. Even though the effectiveness of recycled [Emim][BF4] was not tested in this study, the
works of other authors concerning the recycle of IL showed promising results (Li et al., 2011;

66

Painter et al., 2010b). For example, no decrease in bitumen yield was observed when the IL
was reused five times (Painter et al., 2010b). Li et al. (2011) reused IL 10 times in extraction
with bitumen recovery ranging from 92 to 99%. The cost of IL-enhanced solvent extraction
should be compared with that of solvent extraction on the basis of extraction performance,
operation cost, materials cost, and solvent/IL reuse. For this reason, further research about the
reuse of IL and solvent need to be conducted in order to provide valuable information for the
application of IL-enhanced solvent extraction for oily sludge treatment.
2.3.3. Characteristics of the recovered oil and residual sand/clay
The recovered oil samples obtained under the optimum conditions of each extraction
process were further analyzed. Figure 3.6 shows the gas chromatography (GC) chromatogram
of samples from these treatments. No significant differences in composition among samples
were observed since the same solvent (i.e., cyclohexane) was used in both extraction processes.
Compared to F4 fraction in recovered oil from solvent extraction alone (Figure 3.6a), the F4
fraction in recovered oil from the top oil/solvent layer (Figure 3.6b) decreased and it
compensated for an increase of F4 fraction in oil/solvent-IL interface (Figure 3.6c). It is worth
noting that PHCs distribution in recovered oil using solvent extraction alone and recovered oil
(a combination of top oil/solvent layer and oil/solvent-IL interface) using IL-enhanced solvent
extraction were similar, indicating that IL addition did not promote alteration of the long-chain
and heavy PHCs.

67

a

b

c

Figure 3.6. GC chromatogram of petroleum hydrocarbons in (a) recovered oil using solvent extraction
alone, (b) recovered oil from top oil/solvent layer using IL-enhanced solvent extraction, and (c)
recovered oil from oil/solvent-IL interface using IL-enhanced solvent extraction.

As shown in Figure 3.7, the average F2, F3, and F4 fractions in the recovered oil using
solvent extraction alone were 37.5%, 48.1%, and 14.4%, respectively. The F2 and F3 fraction
in the oil recovered from top oil/solvent layer using IL-enhanced solvent extraction was higher
(i.e., 4.8% and 2.4%), whereas the F4 fraction was lower (i.e., 7.2%) than that using solvent
extraction alone. Approximately 92.7% of TPH distributed in F2 and F3 fractions in top
oil/solvent layer, compared to 97.4% for the crude oil. However, the highest percentage of F4
fraction (i.e., 26.1%) was found in recovered oil from oil/solvent-IL interface using ILenhanced extraction, indicating that a certain amount of heavy PHCs was gathered in
oil/solvent-IL interface instead of dissolving in solvent. GC analysis of the recovered oil also
68

showed that it contained less F2 fraction but more F3 fraction than crude oil, which is good
petrochemical industry feedstock.
80

Crude oil
RO (without IL)
RO in top oil layer (with IL)
RO in oil-IL interface (with IL)

70

Percentage (%)

60
50
40
30
20
10
0

F2

F3

F4

PHCs fractions
Figure 3.7. Distribution of PHCs fractions in crude oil, recovered oil (RO) using solvent extraction
alone, recovered oil in top oil/solvent layer using IL-enhanced solvent extraction, and recovered oil in
oil/solvent-IL interface using IL-enhanced solvent extraction.

The infrared spectra of the oily sludge, the residual sand/clay and the oil extracts obtained
by solvent extraction alone are shown in Figure 3.8. The functional groups in the recovered oil
included C‒H stretching of the saturate (2922 and 2852 cm‒1), C‒H deformation of the saturate
(1462 cm‒1), C‒H symmetric deformation of the saturate (1376 cm‒1), and aromatic C‒H
bending (724 cm‒1). These peaks were also present in the residual sand/clay (Figure 3.8c),
indicating organics remained in residual solids after extraction. Previous studies reported clay
bands (500 cm‒1 and 3800 cm‒1) were displayed in the spectrum of the solvent extract of oil
sands, indicating some fine solid in the recovered bitumen (Li et al., 2011; Painter et al., 2010a;
Sui et al., 2016). However, the recovered oil from solvent extraction of oily sludge was free of

69

fines since no clay bands was observed from Figure 3.8b. This might be because the solids in
the oil sands with strong hydrophobicity tend to enter the solvent phase (Yang et al., 2012).
The broadband between 3700 and 3000 cm‒1 corresponded to the stretching vibrations of
hydrogen bonded O‒H in the water, alcohols, phenols, and carboxylic acids. The peak near
1644 cm‒1 was assigned to vibration of the carboxylic acid group present in the oily sludge.
The bands at 1110 cm‒1 and 1021 cm‒1 could be attributed to Si‒O stretching from
montmorillonite (Wang et al., 2012). Bands due to carbonates, observed near 889 cm‒1
absorbed very strongly in the infrared of residual sand/clay.

a

b

c

Figure 3.8. FTIR spectra of the (a) oily sludge, (b) recovered oil, and (c) residual sand/clay for solvent
extraction alone.

70

The infrared spectra of the IL ([Emim][BF4]), residual sand/clay treated with water washing
at room temperature, and recovered oil (in top oil/solvent layer and oil/solvent-IL interface) by
IL-enhanced solvent extraction are shown in Figure 3.9. As mentioned before, some of the clay
fines were suspended in the middle IL phase after extraction (Figure 3.1). However, the
recovered oil from IL-enhanced solvent extraction was free of mineral fines as seen from the
absence of mineral bands (near 1100 cm‒1 and 890 cm‒1) in Figure 9a and b. Williams et al.
(2010) proposed that the much more favorable electrostatic interactions between the IL and
mineral fines than those between bitumen and mineral fines may contribute to the absence of
mineral fines in the extract. The IL ([Emim][BF4]) was completely immiscible with
hydrocarbons, which is evidenced by the absence of peaks in the range of 750–1200 cm‒1
because of [Emim][BF4] in the spectra of the extracts (Figure 3.9a,b). As a result, the recovered
oil from IL-enhanced solvent extraction is free of fine solids or residual IL. In this study, roomtemperature water was used to separate the entrained IL from residual sand/clay. It can be
found from Figure 3.9c that no IL exists in the residual sand/clay because the peaks for the
residual sand/clay did not overlap with that of [Emim][BF4]. As a result, the IL after extraction
can be removed from the residual sand/clay by water washing at room temperature, and this is
of significance for IL reuse.

71

a

b

c

d

Figure 3.9. FTIR spectra of the (a) recovered oil from top oil/solvent layer, (b) recovered oil from
oil/solvent-IL interface, (c) residual sand/clay, and (d) [Emim][BF4] in IL-enhanced solvent extraction.
Residual sand/clay was treated with water and washed 3 times.

The recovered oil from extraction had a calorific value of 43.92‒45.68 MJ/kg and was
similar to that of crude oil (43.02 MJ/kg), suggesting the recovered oil can be used as fuel
(Table 3.6). However, elemental analysis showed that the recovered oil from solvent extraction
contained small amounts of Fe, Ca, and Mg (Table 3.6). The addition of IL reduced the metal
content in recovered oil in top oil/solvent layer, but increased the metal content in recovered
oil in oil/solvent-IL interface (Table 3.6). There is no clear evidence that the addition of IL
improved the recovery of large molecular fractions of oily sludge (i.e., asphaltenes) since the
concentration of V and Ni (the main source of V and Ni are asphaltene (Jacobs and Filby,
72

1983)) in recovered oil from solvent extraction alone and IL-enhanced solvent extraction were
similar. Although extraction treatments recovered more than 90% TPH from oily sludge, the
TPH content in the residual sand/clay (7.9% and 9.5% for solvent extraction alone and ILenhanced solvent extraction, respectively) was still much higher than the criteria (<0.7% of
extractable PHC in the soil at industrial area) regulated by British Columbia Ministry of
Environment, Lands and Parks (CCME, 2008b). The concentrations of elements in the residual
sand/clay (such as Cr, Ni, and As) are of concern since they are above the regulatory limits in
soils set by CCME (i.e., 87 mg/kg for Cr, 89 mg/kg for Ni, and 12 mg/kg for As at industrial
area) (CCME, 1999). Therefore, the residual sand/clay after extraction may still require further
treatment.

73

Table 3.6. Properties of extracted oil and residual sand/clay after extraction in different systems

Recovered oil
System

Solvent
extraction
IL+solvent
extraction
a

1a
2b

Residual sand/clay

Calorific C
value
%
MJ/kg
43.92
83.6

H

N

S

Fe

Ca

Mg

V

Ni

C

N

TPH

Cr

Ni

%

%

%

mg mg
/kg /kg
2.3 <2

%

1.7

mg
/kg
0.6

%

0.1

mg
/kg
8.8

%

13.1

mg
/kg
30

17.8

0.2

7.9

mg/
kg
122

mg mg/
/kg kg
82 220

45.68
44.50

12.9
13.2

0.1
0.1

1.6
1.3

18
266

<8
21.1

<0.5
2.3

2.2
2.6

16.6

0.7

9.5

133

88

84.5
83.4

Recovered oil in top oil/solvent layer. b Recovered oil in oil/solvent-IL interface.

74

<2
<2

As

232

3.4. Summary
An IL-enhanced solvent extraction (using cyclohexane and [Emim][BF4]) was investigated
in this study to examine its efficiency of oil recovery from oily sludge. Compared to solvent
extraction alone (using cyclohexane), IL-enhanced solvent extraction achieved a similar or
higher TPH recovery (96.92 ± 4.79%) within a much shorter duration (i.e. 10 min) and a much
lower solvent/sludge ratio (i.e. 4:5 mL/g) under less energy consumption (i.e. 100 rpm)
condition. The proposed process required a remarkably low concentration of IL [Emim][BF4]
(i.e. 1:10 mL/g of IL/sludge ratio). When compared to crude oil, the recovered oil had a higher
level of F3 fraction and similar calorific value. This indicates that oil extracts can be utilized
as a potential energy source and petrochemical feedstock. After extraction, IL existed as a
separate phase that can be recycled for further utilization. There was no evidence of IL and
fine solids existing in the recovered oil after IL-enhanced solvent extraction. Residual IL was
completely separated from residual sand/clay after water washing, which is of significance for
recycling IL. However, further research about the reuse of IL is still needed. In summary, ILenhanced solvent extraction could present an economically competitive approach for the
management of hazardous oily solid waste. Further work also needs to be performed to
understand the extraction mechanism of the oil in the presence of IL, which will serve to the
selection/design of IL for oil recovery purpose.

75

Chapter 4
CO-PYROLYSIS OF METAL CONTAMINATED OILY WASTE FOR OIL RECOVERY
AND HEAVY METAL IMMOBILIZATION2
Abstract
The treatment of metal-contaminated oily waste was investigated by using an approach of
co-pyrolysis with hog fuel in a fixed bed reactor. Both oil recovery and heavy metal
immobilization were evaluated. Three experimental parameters, including pyrolysis
temperature, reaction time, and hog fuel addition in the feedstock, were examined to find their
effects on the treatment performance. Metal immobilization in the solid char obtained from copyrolysis was examined from its speciation characteristics. A higher oil recovery was observed
at a lower pyrolysis temperature, while metal speciation in the char residue showed great
variation with temperature, with more non-bioavailable species observed at a higher
temperature. This is also evident from the calculated risk index (RI) for eco-toxicity assessment,
with RI of 34.6 and 117.1 at 600 °C and 400 °C, respectively. Moreover, the addition of hog
fuel into oily waste for co-pyrolysis led to increased metal immobilization, but the overall oil
recovery was decreased. A low temperature co-pyrolysis (400 °C) by adding 20% of hog fuel
was found as the most effective oily waste disposal strategy in this study with satisfactory oil
recovery and an acceptable risk index of 54.1.
Keywords: co-pyrolysis, heavy metal, hog fuel, oil recovery, risk assessment

This work has been published as Tian, Y., Li, J., Yan, X., Whitcombe, T., Thring, R., 2019. Copyrolysis of metal contaminated oily waste for oil recovery and heavy metal immobilization. J.
Hazard. Mater. 373, 1–10. DOI: 10.1016/j.jhazmat.2019.03.061.
2

76

4.1. Introduction
Petroleum hydrocarbons (PHCs) and heavy metals are potential persistent environmental
toxins (Roy et al., 2005). They could exist at contaminated sites affected by accidental spills
and leaks from petroleum, transportation, and mining industries (Samaksaman et al., 2016).
The coexistence of PHCs and high levels of heavy metals has been detected in oily sludge,
which is one of the most abundant waste produced from petroleum refineries (da Silva et al.,
2012). For example, the heavy metal concentration in oily sludge from a refinery was reported
as 500, 480 and 480 mg/kg dry weight for copper, chromium and nickel, respectively (Admon
et al., 2001). The concentration of zinc (Zn) and lead (Pb) in a raw oily sludge sample was
reported to be 769 and 565 mg/kg, respectively (da Rocha et al., 2010). High concentration of
heavy metals have also been found in drill cuttings which are mixtures of drilling fluids with
small pieces of formation rock (Ball et al., 2012). For example, the results from an North Sea
oil based drill cuttings sample indicated a heavy metal concentration of 160, 374, 61, 489, and
2511 mg/kg for chromium (Cr), copper (Cu), nickel (Ni), lead (Pb), and zinc (Zn), respectively
(Breuer et al., 2004). It was reported that heavy metals in the environment may threaten plants,
animals, and humans due to their long-term persistence and possible bioaccumulation and biomagnification effects (Monroy et al., 2014). The existence of PHCs such as polycyclic
aromatic hydrocarbons (PAHs) would also pose adverse environmental effects, but the cooccurrence of PHCs and heavy metals may have more negative impacts. For example, a
stronger inhibition on soil microbial activity was observed in the case of co-existence of PAHs
with heavy metals than the individual existence of heavy metals or PAHs (MaliszewskaKordybach and Smreczak, 2003).

77

The remediation of soils co-contaminated with PHCs and heavy metals is a challenging
problem since each pollutant type requires a different remediation approach. Bioremediation
as a method to treat PHCs contaminated soils has the advantage of low cost while enhancing
the soil quality and functions. However, the long-term degradation process and the inhibition
of toxic heavy metals on the activity of microorganisms would limit its utilization, thus leading
to the requirement of other alternatives (AL-Saleh and Obuekwe, 2005). Among various
approaches, the pyrolysis of contaminated soil has received growing interests. It can effectively
recover oil from organic wastes, while immobilizing heavy metals in the solid residue, thus
leading to significantly reduced environmental risks due to decreased bioavailability of heavy
metal ions. A number of operating variables may affect the pyrolysis treatment, such as
pyrolysis temperature, reaction time, and chemical additives (Shie et al., 2003; Wang et al.,
2007). It was reported that the co-pyrolysis with other waste materials could even enhance the
treatment performance (Debela et al., 2012; Shi et al., 2013b). For example, Debela et al. (2012)
found that the pyrolysis of contaminated soil immobilized some heavy metals such as Pb and
Cd, while the Cd leaching was further reduced with the addition of 5, 10, and 15% biomass in
the soil. Shi et al. (2013b) also reported that the addition of rice husk during the hydrothermal
carbonization of sewage sludge resulted in a synergistic effect on reducing the risk of heavy
metals (Cr, Ni, Cu, and Cd) in the solid residues. However, many previous studies on the
thermal treatment of waste focused on the total heavy metal contents and the leaching behavior
of residues after treatment (Jin et al., 2014; Thurnau, 1996). Their results can hardly be used
for evaluating the eco-toxicity of treated waste, which depends on the specific chemical forms
or binding patterns of metals (He et al., 2010).

78

As an extension of our previous study on the co-pyrolysis of oily sludge for energy recovery
(Hu et al., 2017), this work was conducted to investigate the pyrolysis of oily waste containing
heavy metals for both oil recovery and heavy metal immobilization. Hog fuel (i.e. wood waste)
from lumber production was used as an additive, aiming to explore the reduced bioavailability
and toxicity of heavy metals in the pyrolysis char. The metal chemical speciation in the solid
char residue was examined to evaluate the ecological risk posed by oily waste treatment. The
effects of three selected experimental variables (pyrolysis temperature, reaction time, and
wood waste addition) on the process performance in terms of oil recovery and the ecological
risk of solid char residue, were studied using a response surface methodology (RSM). The
results would provide valuable information for developing effective technique to address oily
waste problem with added economic and environmental benefits.
4.2. Materials and methods
4.2.1. Materials
The clean soil was collected from a forest site located east of Prince George, BC, and it was
air-dried at room temperature for 3–4 days and then sieved to 0.84 mm (passed through a # 20
sieve). Crude oil sample was donated by an oil refinery in western Canada. The wood waste
(called hog fuel) used for co-pyrolysis was collected from a local saw mill in Prince George,
and it was ground to < 1.0 mm before use. To prepare the synthetic oily waste containing heavy
metals, 500 g of clean soil was first mixed well with the appropriate amount of dissolved
Cu(NO3)2, Cr(NO3)3, Ni(NO3)2, Pb(NO3)2, and ZnCl2 solution to obtain an initial soil
concentration of 1000 mg/kg for Cu2+, Cr3+, Ni2+, Pb2+, and Zn2+, respectively. The air-dried
heavy metal contaminated soil was then thoroughly mixed with 20 wt% of crude oil diluted in

79

dichloromethane (DCM), and the mixture was left in the fume hood to allow for solvent
evaporation. Table 4.1 lists the synthetic oily waste properties, including the concentrations of
total petroleum hydrocarbon (TPH) and heavy metals. Such synthetic heavy metal contained
oily waste was then used for pyrolysis study through the addition of hog fuel with different
mixing ratio of 0%, 10%, and 20% (w/w), respectively.
Table 4.1. Soil texture and synthetic oily waste properties

Synthetic oily waste
Parameter

Concentration (mg/kg)

Soil texture

TPH

17.5 ± 0.35% (by mass)

Sand

10.8%

Copper (Cu)

748 ± 3

Silt

73.9%

Chromium (Cr)

826 ± 3

Clay

15.3%

Nickel (Ni)

828 ± 4

Lead (Pb)

789 ± 1

Zinc (Zn)

865 ± 4

4.2.2. Experimental set up
Each pyrolysis experiment was carried out in a fixed-bed tube furnace pyrolysis reactor
(quartz tube length: 600 mm, Ø: 50 mm; MTI Corp.® GSL-1100X) under nitrogen atmospheric
pressure (Figure 4.1). The oily waste sample was put into the sample ark which was placed in
the tube reactor center. Before experiment, nitrogen gas was blown into the reactor at a rate of
1 L/min for 30 min to create an oxygen-free environment. The nitrogen flow rate was then set
to 100 mL/min and maintained throughout the pyrolysis reaction with a heating rate of
20 °C/min. During the reaction period, the pyrolysis temperature was kept at the set value that
was determined from a response surface experimental design method. The condensable volatile
80

matter was collected in the ice-water cooling traps as the recovered oil. The solid char residue
was collected after the furnace was cooled down to room temperature.
6

5
4

2

3

8

1

9

7

Figure 4.1. Schematic diagram of fixed-bed pyrolysis reactor (1 - N2 gas cylinder, 2 - flow meter, 3 -

thermal couple, 4 - quartz tube, 5 - sample ark, 6 - tube furnace, 7 - temperature controller, 8 - ice-water
bath, 9 - Tedlar gas bag).

4.2.3. Experimental design
In this study, three experimental factors were examined for their effects on the co-pyrolysis
performance, including pyrolysis temperature, reaction time, and the mixing ratio of hog fuel
with oily waste. The selection of these factors was based on previous pyrolysis studies of
organic wastes. Pyrolysis temperature not only affects oil yield but also the structure and
chemical characteristics of the solid char residue, including its chemical composition, pH,
surface area, and functional groups (H. Li et al., 2017; Qin et al., 2015). Reaction time could
affect the completeness of pyrolysis process (Xiu et al., 2010), while the addition of waste
biomass to pyrolysis process may generate synergistic effect on metal immobilization (Debela
81

et al., 2012; Shi et al., 2013b). In this study, a pyrolysis temperature range of 400–600 °C was
selected based on our previous study on oily sludge co-pyrolysis for energy recovery, and this
range was believed to cover the full decomposition temperature of PHCs (Hu et al., 2017). A
reaction time range of 30–120 min was selected based on the works of other authors concerning
oil production from biomass by liquefaction/pyrolysis and metal immobilization by thermal
treatment (Abnisa et al., 2011; Debela et al., 2012; L. Li et al., 2012; Yang et al., 2004). A hog
fuel addition range of 0–20% was selected based on the consideration that one purpose of the
study was to examine the synergistic effect of biomass addition but too much hog fuel addition
would increase the waste volume.
The relationship between the experimental factors and pyrolysis system performance was
examined using a response surface methodology (RSM) for experimental design (Bezerra et
al., 2008). When several independent factors and their interaction have influence on the
experimental outcome, the RSM is an effective tool for the optimization of these factors with
a minimum number of experimental runs (Ahmadi et al., 2005). The RSM design must include
three to five replicates of central point runs, but replicating the corner-point experimental run
is not mandatory (Montgomery, 2017). In this study, a total of 15 experiments with three
replicates in the central point (in order to estimate the pure error) were designed by using BoxBehnken design (BBD) method through the Design Expert software (Ver. 10.1). Such design
method was also reported in previous studies (Xu et al., 2018). Table 4.2 lists the arrangement
of experiments, with each experimental factor varying over three levels, including high level
(+1), low level (‒1), and central point (coded level 0). The experiments were carried out in a
random order to minimize the effects of uncontrolled factors. The co-pyrolysis process was
evaluated for its performance (i.e. response) by analyzing the TPH recovery rate (𝑌1 , %) and

82

the potential ecological risk index (RI) of solid char residue (𝑌2 ). A multiple regression analysis
was conducted by using Design Expert software (Ver. 10.1):
𝑌 = 𝛽0 + ∑𝑘𝑗=1 𝛽𝑗 𝑥𝑗 + ∑𝑘𝑗=1 𝛽𝑗𝑗 𝑥𝑗2 + ∑𝑖 ∑𝑘<𝑗=2 𝛽𝑖𝑗 𝑥𝑖 𝑥𝑗 + 𝑒𝑖

(4.1)

where 𝑌 is the response; 𝑘 is the number of independent variables; 𝑥𝑖 and 𝑥𝑗 are variables; 𝛽0
is the model intercept term; 𝛽𝑗 is the first order (linear) main effect; 𝛽𝑗𝑗 is the quadratic effect;
𝛽𝑖𝑗 is the interaction effect; and 𝑒𝑖 is the error (Bezerra et al., 2008). Analysis of variance
(ANOVA) was used to test the significance of independent variables and their interaction.

83

Table 4.2. Coded (𝑥𝑖 ) and un-coded (𝑋𝑖 ) Box-Behnken design matrix with the experimental results

Pyrolysis
temperature
No.

(°C)
𝑥1

𝑋1

Reaction time

Hog fuel addition

(min)

(%)

𝑥2

𝑋2

𝑥3

TPH
recovery

RI

(%)
𝑋3

𝑌1

𝑌2

(coded) (uncoded)

(coded) (uncoded)

(coded) (uncoded)

(observed)

(observed)

1

-1

400

-1

30

0

10

82.2

87.6

2

1

600

0

75

1

20

68.6

22.0

3

-1

400

0

75

1

20

68.7

53.3

4

0

500

-1

30

1

20

58.3

41.5

5

0

500

-1

30

-1

0

83.6

77.2

6

-1

400

1

120

0

10

65.4

84.3

7

1

600

0

75

-1

0

69.5

35.5

8

1

600

1

120

0

10

62.6

31.7

9

0

500

1

120

1

20

56.8

30.8

10

1

600

-1

30

0

10

68.3

33.2

11

-1

400

0

75

-1

0

82.9

120.2

12

0

500

1

120

-1

0

67.1

59.8

13

0

500

0

75

0

10

74.7

42.2

14

0

500

0

75

0

10

72.5

38.7

15

0

500

0

75

0

10

74.9

39.0

84

4.2.4. Chemical analysis and heavy metal species characterization
Analysis of petroleum hydrocarbons
A weighed mass (𝑀) of the recovered oil was collected into a 40 mL vial, followed by the
addition of 20 mL DCM solvent to the vial. The vial was then placed on a Talboy 3500 orbital
shaker for mechanical extraction at 150 rpm for 30 min. After shaking, the extraction solution
was transferred for silica gel column cleanup to remove moisture, particulate, and unwanted
polar organic compounds (CCME, 2001). A small amount of glass wool was placed into the
bottom of the glass column which has a length of 30 cm and an internal diameter of 16 mm.
The column was packed with approximately 6.5 cm of 70–230 mesh silica gel (activated at
110 °C for 12 h) and approximately 2.5 cm of 10–60 mesh anhydrous sodium sulfate (dried at
400 °C for 4 h). Before use, the column was rinsed with 20 mL of solvent (1:1 n-hexane/DCM).
The extraction solution was carefully transferred into the column and then followed by another
20 mL of solvent (1:1 n-hexane/DCM) to elute the column. A round flask weighed (𝑀0 ) was
placed under the bottom of the column to collect the cleaned extract. The solvent in the cleaned
extract was evaporated by using a Yamato RE400 rotary evaporator. The round flask
containing the PHCs after evaporation was weighed (𝑀𝑇 ). The concentration of TPH in the
recovered oil can then be calculated. Similarly, TPH concentration in the original oily waste
sample can also be obtained.
𝐶=

𝑀𝑇 −𝑀0
𝑀

× 106

(4.2)

where 𝑀 is the mass of recovered oil or the original oily waste sample (g); 𝑀0 is the mass of
the round flask (g); 𝑀𝑇 is the mass of round flask containing PHCs (g); and 𝐶 is the TPH
concentration in recovered oil or in the original oily waste sample (mg/kg).

85

The TPH recovery rate was calculated as the ratio of TPH mass in the recovered oil to the
mass of TPH in the original oily waste sample:
𝐶 ×𝑀

𝑅 = 𝐶𝑟 ×𝑀𝑟 × 100%
𝑠

𝑠

(4.3)

where 𝑅 is the TPH recovery rate (%); 𝐶𝑟 and 𝐶𝑠 (mg/kg) are TPH concentrations in the
recovered oil and oily waste sample, respectively; and 𝑀𝑟 and 𝑀𝑠 (g) are the mass of recovered
oil and oily waste sample, respectively.
The PHCs fractions (F2, F3, and F4) in the recovered oil and original crude oil were also
measured using GC analysis method on an Agilent 6890 GC-FID as described in Section
3.2.4.1.
Characterization of solid residue
The total carbon (C), nitrogen (N), and hydrogen (H) of the feed stocks (oily waste and hog
fuel) and solid residue samples were measured by combustion elemental analysis using a
Costch ECS 4010 elemental analyzer. The pH of the samples was measured in a supernatant
of aqueous solution (sample/water = 1:50, w/v) with a Mettler Toledo FG2 portable pH meter
(Bagreev et al., 2001). The ash content of sample was determined according to methods
recommended for forest soil and plant analysis (Kalra and Maynard, 1991). The samples were
also ground and mixed with KBr (0.1 wt%) and then pressed into pellets. The Fourier transform
infrared (FTIR) spectra were then obtained using a Nexus 670 Thermo Nicolet Spectrometer.
Sequential extraction of heavy metals
It was recognized that the ecological toxicity of heavy metal depends on its chemical
speciation, including water soluble fraction, exchangeable fraction, easily reduced fraction,
oxidizable fraction, and residual fraction. These five species can be determined using a five86

step sequential extraction procedure. The heavy metal species in the oily waste and the
pyrolysis char residue were analyzed using the following modified procedure (Devi and Saroha,
2014; Pöykiö et al., 2006):
Step 1 (Water soluble fraction): 0.5g of sample was extracted with 20 mL of acidified
deionized water (pH = 4.0) by shaking on a Talboy 3500 orbital shaker at 60 rpm for 16 h at
room temperature. The extract was then collected by centrifugation (Thermo Scientific Sorvall
Legend X1) at 5000 rpm for 10 min, filtered through a 0.45-µm nylon filter and stored at 4 °C
until further analysis. The remaining solid was washed by shaking at 160 rpm for 1 h with 10
mL of deionized water, followed by centrifugation at 5000 rpm for 10 min. The supernatant
was discarded.
Step 2 (Exchangeable fraction): the residue from step 1 was shaken on a Talboy 3500
orbital shaker at 60 rpm for 16 h at room temperature with 20 mL of 0.11M acetic acid (pH =
2.8). Similar to step 1, the extract was collected by centrifugation and then filtered. The residue
was washed, followed by centrifugal separation as in step 1.
Step 3 (Easily reduced fraction): 20 mL of 0.11 M hydroxylammonium chloride solution,
acidified to pH of 2 with HNO3, was added to the residue from the 2nd step, and the mixture
was shaken at 60 rpm at room temperature for 16 h. The extract was then subject to centrifugal
separation and nylon filter filtration, and the resulting residue was washed as in previous steps.
Step 4 (Oxidizable fraction): 5 mL of 30% H2O2 was carefully added to the residue from
step 3. The mixture was digested in a 50-mL centrifuge tube at room temperature for 1 h with
occasional manual shaking. The digestion was continued in a sand bath for 1 h at 85 ± 5 °C,
and the volume was reduced to < 1 mL by further heating of the uncapped tube. A second 5
mL of H2O2 was added to repeat the digestion procedure. 25 mL of 1 M ammonium acetate,

87

adjusted to pH of 2, was added to the cool moist residue, and the mixture was shaken at 60 rpm
for 16 h at room temperature. Similar to the previous steps, the extract was separated and
filtered, and the residue was washed.
Step 5 (Residual fraction): the residue from step 4 was digested in concentrated HNO3 and
concentrated HCl, following the microwave assisted digestion method (USEPA 3051A)
(USEPA, 2007a). The resulting solutions were used to determine the content of heavy metals.
The concentrations of Zn, Cu, Ni, Pb, and Cr in the extract from each step were analyzed
by inductively coupled plasma optical emission spectrometry (ICP-OES) (Agilent
Technologies 5100). Two independent replicates were performed for each sample.
4.2.5. Determination of potential ecological risk index (RI)
The potential risk posed by heavy metals in pyrolysis char residue was assessed using a
potential ecological risk index (𝑅𝐼) as follows (Devi and Saroha, 2014; Jin et al., 2016; Zhao
et al., 2012):
𝐶𝑓 = 𝐶𝑖 ⁄𝐶𝑛

(4.4)

𝐸𝑟 = 𝑇𝑟 𝐶𝑓

(4.5)

𝑅𝐼 = ∑ 𝐸𝑟

(4.6)

where 𝐶𝑓 is the contamination factor; 𝐶𝑖 is the sum of concentration of metal ions extracted in
the first four steps of the sequential extraction; 𝐶𝑛 is the concentration of metal ions in residual
fraction; 𝐸𝑟 is the potential ecological index for individual heavy metal ion; 𝑇𝑟 is toxicity
factor of the individual heavy metal ion; and 𝑅𝐼 is the potential ecological risk index. The 𝑇𝑟
values used in this study are 1, 5, 6, 5 and 2 for Zn, Cu, Ni, Pb, and Cr, respectively
(Chabukdhara and Nema, 2011; Devi and Saroha, 2014).

88

4.3. Results and discussion
4.3.1. Response surface model and statistical analysis
The following second-order polynomial equations were established by applying multiple
regression analysis of the experiment results:
TPH recovery (%) = 137.01 ‒ 0.12𝑋1 ‒ 0.11𝑋2 ‒ 2.43𝑋3+ 6.15 × 10‒4𝑋1 𝑋2 + 3.33 × 10-3𝑋2 𝑋3
+ 8.33 × 10‒3𝑋1 𝑋3 ‒ 2.60 × 10‒3𝑋22 ‒ 2.45 × 10‒2𝑋32

(4.7)

RI = 610.68 ‒ 1.65𝑋1 ‒ 0.69𝑋2 ‒ 9.86𝑋3+ 9.83 × 10‒5𝑋1 𝑋2 + 1.33× 10‒2𝑋1 𝑋3 + 3.68 × 10‒
3

𝑋2 𝑋3 + 1.23 × 10‒3𝑋12 + 3.41 × 10‒3𝑋22 + 5.44 × 10‒2𝑋32

(4.8)

where X1, X2, and X3 are pyrolysis temperature (°C), reaction time (minutes), and hog fuel
addition (%); if 400 °C ≤ X1 ≤ 600 °C, 30 min ≤ X2 ≤ 120 min, and 0 % ≤ X3 ≤ 20 %.
Fitness of the model was assessed using ANOVA (Tables 4.3 and 4.4). The model F-values
(10.659 and 49.498) indicated that these two models could well describe the TPH recovery rate
and RI, with P values (0.0049 and 0.0002) all less than 0.05. In terms of the model for TPH
recovery, the determination coefficient (R2) reached 0.9343, while all the variables (𝑥1 , 𝑥2 , and
𝑥3 ) showed significant effects. In terms of the model for ecological risk index (RI), the
determination coefficient (R2) reached 0.9889, while only pyrolysis temperature (𝑥1 ) and hog
fuel addition ratio (𝑥3 ) had significant effects.

89

Table 4.3. ANOVA of the regression model for TPH recovery

Source

Sum of
squares

DFa Mean square

F-value

Prob > F Significance

Model

891.605

8

111.451

10.659

0.0049

**

𝑥1

113.572

1

113.572

10.862

0.0165

*

𝑥2

205.885

1

205.885

19.691

0.0044

**

𝑥3

322.137

1

322.137

30.809

0.0014

**

𝑥1 𝑥2

30.597

1

30.597

2.926

0.1380

𝑥1 𝑥3

44.217

1

44.217

4.229

0.0855

𝑥2 𝑥3

56.227

1

56.227

5.378

0.0595

𝑥22

102.897

1

102.897

9.841

0.0201

𝑥32

22.311

1

22.311

2.134

0.1944

Residual

62.735

6

10.456

Lack of fit 58.945

4

14.736

7.777

0.1172

Pure error

3.790

2

1.895

Cor total

954.340

14

Std. Dev.

3.23

𝑅2

0.9343

mean

70.41

Adj 𝑅 2

0.8466

CV

4.59

Pred 𝑅 2

0.3472

PRESS

623.02

Adeq precision

10.5286

a

Degree of freedom; * Significant (P < 0.05); ** Highly significant (P < 0.01).

90

*

Table 4.4. ANOVA of the regression model for RI in solid residue

Source

Sum of
squares

DFa Mean square

F-value

Prob > F

Significance

Model

10473.981

9

1163.776

49.498

0.0002

**

𝑥1

6223.529

1

6223.529

264.703

<0.0001

**

𝑥2

135.454

1

135.454

5.761

0.0616

𝑥3

2633.774

1

2633.774

112.021

0.0001

𝑥1 𝑥2

0.783

1

0.783

0.033

0.8624

𝑥1 𝑥3

714.331

1

714.331

30.382

0.0027

𝑥2 𝑥3

11.051

1

11.051

0.470

0.5235

𝑥12

561.804

1

561.804

23.895

0.0045

**

𝑥22

175.656

1

175.656

7.471

0.0411

*

𝑥32

109.398

1

109.398

4.653

0.0835

Residual

117.557

5

23.511

Lack of fit

109.948

3

36.649

9.633

0.095505

Pure error

7.609

2

3.805

Cor total

10591.538

14

Std. Dev.

4.85

𝑅2

0.9889

mean

53.12

Adj 𝑅 2

0.9689

CV

9.13

Pred 𝑅 2

0.8323

PRESS

1776.28

Adeq precision

23.2560

a

**

**

Degree of freedom; * Significant (P < 0.05); ** Highly significant (P < 0.01).

4.3.2. TPH recovery rate
The co-pyrolysis of oily waste and wood waste produced an oil mixture with two phases
including a black immiscible liquid likely due to the oil derived from wood waste. Other studies

91

reported similar phase separation, with oils produced from oily sludge/wood waste blends
having two phases (an upper phase and a bottom phase) (Abnisa and Wan Daud, 2015; Hu et
al., 2017; Samanya et al., 2012). In this study, the bottom black immiscible liquid phase was
not observed in the oil produced from the pyrolysis of oily waste alone, indicating that this
phase was originated from the wood waste. Moreover, the yield of the bottom phase increased
(from visual observation) when the hog fuel mixing ratio in the feedstock increased. Given this
finding, the upper phase of oil mixture was considered as the recovered oil from oily waste in
this study.
As shown in Figure 4.2, there is no significant difference in the oil composition between
the original oily waste and the recovered oil, indicating that the distillation during pyrolysis is
the predominant mechanism for oil recovery. Figure 4.3 shows the variation of TPH recovery
against each experimental parameter while all other factors were maintained constant at the
coded level of 0. It was reported that the highest liquid yield was generally obtained in the
range of 500–600 °C (Akhtar and Amin, 2012; Lam et al., 2010; Liu et al., 2008). At
temperatures below this range, incomplete pyrolytic decomposition occurs, while higher
temperatures could enhance the production of condensable vapor from waste, leading to a
higher yield of liquid oil product. However, at pyrolysis temperatures above this range, oil
yield could decrease as a result of potential degradation of oils to gases by secondary cracking
(Lam et al., 2010). From Figure 4.3(a), a trend can be observed in which TPH recovery
decreased with increasing temperature in the experimental range of 400–600 °C. This trend
indicates that the desirable pyrolysis temperature could be less than or equal to 400 °C. Such
desirable temperature may vary because of different experimental conditions (e.g., heating rate,
type of reactor, and nitrogen flow rate). Sınağ et al. (2010) conducted similar experiments to

92

treat waste machinery oil and obtained the highest oil yield at 400 °C. Piskorz et al. (1986) also
reported that the maximum yield of organic liquid occurred at 400 °C. Our results are
consistent with these previous studies, suggesting that temperature higher than 400 °C is not
always needed for waste pyrolysis.

a

b

Figure 4.2. Gas chromatogram of (a) oil from untreated oily waste and (b) recovered oil after

pyrolysis treatment (run #11).

93

b

TPH recovery rate (%)

TPH recovery rate (%)

a

Pyrolysis temperature (ºC)

Reaction time (min)

TPH recovery rate (%)

c

Hog fuel addition (%)
Figure 4.3. The statistical evaluation of the effect of single factor on TPH recovery rate, (a) pyrolysis
temperature (reaction time: 75 min, hog fuel ratio: 10%), (b) reaction time (pyrolysis temperature: 500 °C, hog
fuel ratio: 10%), and (c) hog fuel addition (pyrolysis temperature: 500 °C, reaction time: 75 min).

As shown in Figure 4.3(b), the TPH recovery reached the maximum (75.8%) at a reaction
time of 54 min and then decreased beyond 54 min. An appropriate reaction time could improve
pyrolysis conversion, but excessive duration would cause decrease in the TPH recovery, owing
to the decomposition of some oil and/or the formation of solid residue by oil repolymerization
(Xiu et al., 2010; Yang et al., 2004). As seen in Figure 4.3(c), the TPH recovery decreased
94

with increasing hog fuel ratio in the feedstock. Abnisa and Wan Daud (2015) found lower
experimental yield of TPH than the calculated yield when studying the fuel recovery from copyrolysis of palm shell and scrap tire. Similar trends were reported in other co-pyrolysis studies
(Hu et al., 2017; Krerkkaiwan et al., 2013; Park et al., 2010; Song et al., 2014). Such effect of
hog fuel addition during co-pyrolysis of oily waste could be attributed to the role of alkali and
alkaline earth metallic species (AAEMs) present in wood waste (Song et al., 2014).
As shown in Figure 4.4, different combination of pyrolysis temperature and reaction time
could achieve similar TPH recovery rate, while an increase in reaction time could compensate
for the effect due to a decrease in temperature. However, the pyrolysis temperature did not
affect the TPH recovery at reaction time > 110 min, indicating that the secondary cracking of
oil was independent of the whole temperature range (400–600 °C). As seen in Figure 4.5(a),
the reaction time required for maximizing TPH recovery was dependent on the hog fuel ratio
in feedstock, with maximum TPH recovery for 0%, 10%, and 20% hog fuel addition occurred
at reaction time of 37.2 min, 53.6 min, and 68.1 min, respectively. This result implied that a
longer reaction time was required to allow a complete reaction for the co-pyrolysis of oily
waste/wood waste mixture as compared to the pyrolysis of oily waste alone. Figure 4.5(b)
shows that the higher TPH recovery rate occurred when both the pyrolysis temperature and
hog fuel addition were kept to a lower value, which is consistent with the trends observed from
Figure 4.3(a) and (c).

95

Reaction time (min)

Pyrolysis temperature (ºC)

Figure 4.4. Contour plot of TPH recovery (%) as function of pyrolysis temperature and reaction time

a

TPH recovery rate (%)

TPH recovery rate (%)

at hog fuel addition = 10%.

b

Figure 4.5. Response surface graph of TPH recovery as function of (a) reaction time and hog fuel
addition at pyrolysis temperature = 500 °C, and (b) pyrolysis temperature and hog fuel addition at
reaction time = 75 min.

4.3.3. Properties of solid residues
Table 4.5 lists the total C, H, N, and pH in the feedstock, the solid residue derived from
pyrolysis of oily waste alone, and the residue from the co-pyrolysis of oily waste and hog fuel
96

mixture. The aromaticity degree of solid residue can be identified by the H/C ratio (Xiao et al.,
2016). The H/C ratios of solid residues decreased with increasing pyrolysis temperature, but
they were all below 0.18 at different pyrolysis temperatures, indicating a high degree of
aromaticity under which the residues are resistant to decomposition (Yuan et al., 2013). The
N/C ratio was decreased from 0.034 for solid residue after pyrolysis at 400 °C (SRO400) to
0.022 for solid residue after pyrolysis at 600 °C (SRO600). This would suggest that the loss of
nitrogen-related function groups at higher temperature in the absence of hog fuel (T. Chen et
al., 2014). It was also observed that the pH value of solid residues increased with pyrolysis
temperature. This may be attributed to the formation of alkali carbonates within the pyrolyzed
material (T. Chen et al., 2014).
Table 4.5. Total C, H, N, and pH in the feedstock and the solid residue after pyrolysis

Sample

C (%)a

Hog Fuel

H (%)a

N (%)a

H/C

N/C

49.17 ± 1.69 6.17 ± 0.01

0.03 ± 0.04

0.125

0.001 ND

SRC

16.71 ± 1.80 2.50 ± 0.69

0.13 ± 0.03

0.150

0.008 3.93 ± 0.14 84.0 ± 0.14

SRO400

2.14 ± 0.06

0.39 ± 0.04

0.07 ± 0.01

0.180

0.034 6.07 ± 0.28 97.7 ± 0.01

SRO600

2.02 ± 0.09

0.10 ± 0.09

0.04 ± 0.01

0.050

0.022 7.20 ± 0.10 98.4 ± 0.04

SROH400

9.37 ± 0.97

0.65 ± 0.01

0.10 ± 0.03

0.069

0.010 6.32 ± 0.05 90.3 ± 0.09

SROH600

7.54 ± 0.48

0.30 ± 0.01

0.07 ± 0.01

0.040

0.009 7.14 ± 0.17 92.7 ± 0.16

pH

Ash (%)a
1.59 ± 0.09

Note: SRC: solid residue sample from control experiment without pyrolysis treatment; SROx: solid
residue from pyrolysis of oily waste alone at X (°C) temperature and reaction time = 75 min; SROHx:
solid residue from co-pyrolysis of oily waste and hog fuel at X (°C) temperature and reaction time = 75
min and hog fuel addition = 20%; ND: not detected.
a
On a dry basis.

The FTIR spectra of the feedstock and the solid residues after pyrolysis are shown in Figure
B1 in Appendix B. The peak occurring at around 3600 cm‒1 corresponds to the vibration of
OH groups in the mineralized matter (Hossain et al., 2011). The intensity of the broad band
97

between 3500 and 3300 cm‒1 decreased from feedstock to solid residues as a result of the
decomposition of organic OH groups at elevated temperatures (Jin et al., 2016). The pyrolysis
resulted in a disappearance in C‒H stretching at 2905 and 2850 cm‒1, indicating that the
carbonaceous materials were transformed into CH4, CO2 and other gases or to aromatic
structures (Lu et al., 2013). Jin et al. (2016) reported that the amide bond stretching (‒CO‒
NH‒) at 1633 cm‒1 was offset when the pyrolysis temperature increased, suggesting that the
amide functional groups were complexed with heavy metals during pyrolysis. However, the
FTIR spectra peak at 1633 cm‒1 was more likely to be resulted from C=O stretching vibration
of ketone, aldehyde, lacton, amide, and carboxyl in this study. The strong peaks of C‒O
stretching vibration at 1038 cm‒1 suggested that various forms of oxygen were directly bonded
to the adjacent carbon element during pyrolysis (T. Chen et al., 2014). The peak intensity
between 800–600 cm‒1 was likely arising from the aromatic and hetero-aromatic compounds
(Hossain et al., 2011), which is consistent with the results for H/C ratio shown in Table 4.5.
The peaks below 600 cm‒1 were assigned to the metal-ligand stretching vibrations in inorganic
compounds (Hossain et al., 2011).
4.3.4. Heavy metal speciation characteristics
Metal ions can be categorized into five species (fractions), including a water soluble fraction
(F1), exchangeable fraction (F2), reducible fraction (F3), oxidizable fraction (F4), and residual
fraction (F5). The bioavailability and mobilization of these metal species follow the order of
F1 > F2 > F3 > F4 > F5. The species in (F1 + F2) are very prone to leaching, and thus they are
included in the bioavailable category. Species in (F3 + F4) are identified as potential
bioavailable species which can undergo degradation and leaching under very rigorous
conditions (e.g., highly acidic conditions and an oxidizing atmosphere). Metals in F5 possess
98

no toxicity because they are intercalated into clay layers or ionic bound to the surface structures
of silicates (Huang and Yuan, 2016).
Figure 4.6 shows the ratio of the five species in each metal in the original oily waste and
its solid residue after pyrolysis. In the case of original oily waste, the non-bioavailable species
(F5) only accounted for less than 12% in each of the five studied metals, implying a relatively
high ecological risk associated with these metals in the oily waste. However, pyrolysis
treatment significantly reduced the potential risk by transforming weakly bound metal species
(F1 and F2) to more strongly bound complexes (F5) in the solid char residue. After pyrolysis
treatment, the ratios of (F1 + F2) in Cu, Ni, Pb, and Zn were significantly decreased, while the
(F3 + F4) and F5 ratios increased, except for Pb in solid residue after pyrolysis at 600 °C
(SRO600). In terms of Cr, the ratios of F1, F2, F3, and F4 decreased while the F5 ratio
increased. This indicates that the pyrolysis treatment enhanced the complexation of chromium
ions, generating more tightly bound complexes in F5 than (F3 + F4).

99

80

40

Fractions of Cr content (%)

60

Fractions of Cu content (%)

100

SRC
SRO400
SRO600
SROH400
SROH600

20
4
2

F1

Fractions of Ni content (%)

F1+F2

F3

F4

F3+F4

60
40
20
8
4
F1

100

F2

F1+F2

F3

F4

F3+F4

20
3.0
1.5
F1

60

F2

F1+F2

F3

F4

F3+F4

F5

F3

F4

F3+F4

F5

SRC
SRO400
SRO600
SROH400
SROH600

60
40
20
1.0
0.5
0.0

F5

SRC
SRO400
SRO600
SROH400
SROH600

80

F1

F2

F1+F2

SRC: solid residue from control experiment
without pyrolysis treatment;
SROx: solid residue from pyrolysis of oily
waste alone at X (°C) temperature at reaction
time = 75 min;

40
20

SROHx: solid residue from co-pyrolysis at X
(°C) temperature under hog fuel addition of
20% and reaction time of 75 min;

4

F1 - water soluble fraction, F2 - exchangeable
fraction, F3 - reducible fraction, F4 - oxidizable
fraction, F5 - residual fraction.

2
0

40

0.0

F5

SRC
SRO400
SRO600
SROH400
SROH600

80

0

60

80

100

Fractions of Zn content (%)

F2

Fractions of Pb content (%)

0

SRC
SRO400
SRO600
SROH400
SROH600

80

F1

F2

F1+F2

F3

F4

F3+F4

F5

Figure 4.6. Fraction distribution of metals in oily waste before and after pyrolysis treatment.

100

The concentrations of metal species in the bioavailable category (F1 + F2) were found to
decrease with pyrolysis temperature. For example, the percentage of the bioavailable (F1 + F2)
species in Cu decreased significantly from 67.32% for the original oily waste (SRC) to 32.07%
for solid residue after pyrolysis at 400 °C (SRO400) and further to 16.15% for solid residue
after pyrolysis at 600 °C (SRO600). Similar decrease trends were also observed for Cr, Ni, Pb,
and Zn (Figure 4.6). The (F1 + F2) ratios in Cr, Ni, Pb, and Zn decreased from 1.09%, 8.39%,
16.68%, and 11.83% for SRO400 to 0.13%, 3.65%, 5.10%, and 3.50% for SRO600,
respectively. In addition, the F5 ratios in Cu, Cr, Pb, and Zn were increased by 2.5, 2.9, 4.8,
and 2.0 times for SRO600, respectively, as compared to those for SRO400. Interestingly, the
immobilization of heavy metals was further enhanced by hog fuel addition. With 20% addition
of hog fuel, the ratios of (F1 + F2) and (F3 + F4) in Cu showed a decreasing trend in the solid
residue from the co-pyrolysis at 400 °C (SROH400) as compared to that for SRO400. The F5
ratio in Cu was increased to 22.75% for SROH400, which is 2.9 times more than that for
SRO400. A decreased ratio of bioavailable (F1 + F2) and potential bioavailable (F3 + F4)
species in metal was also found for Ni, Pb, and Zn after co-pyrolysis at 400 °C and 600 °C.
However, a slight increase of potential bioavailable species (F3 + F4) was detected for Cr after
hog fuel addition, but the ratio of bioavailable Cr species still decreased with co-pyrolysis
treatment.
It was found that the distribution pattern of chemical species (F1‒F5) was different for each
heavy metal after pyrolysis. In general, the bioavailable fractions (F1 and F2) of all the metal
ions decreased whereas the residual fraction (F5) increased after pyrolysis treatment, during
which the metal immobilization effect could be largely enhanced by increasing the pyrolysis

101

temperature. This might be caused by the precipitation and complexation between heavy
metals and the (crystal) lattice of the solid residue phase (Shi et al., 2013a). It might also be
due to the fact that pyrolysis contributes alkalinity to solid residue. The higher the temperature,
the more alkaline the solid residue (Hossain et al., 2011; Jin et al., 2016). The high pH could
then favor the formation of metal hydroxide, leading to the formation of insoluble metal oxides
which are less prone to leaching (Li et al., 2001). As a result, the pyrolysis had a positive effect
on the immobilization of heavy metals in solid residue.
Moreover, it is worth noting that compared with pyrolysis of oily waste alone, the addition
of hog fuel for co-pyrolysis further decreased the toxicity of solid residue by reducing the
bioavailability and potential bioavailability of Cu, Ni, Pb, and Zn. Uchimiya et al. (2010) found
similar results from the enhanced retention of Ni and Cd in soil by the addition of bio-char (i.e.
a carbon-rich, solid product from pyrolysis of biomass). The increase in soil pH by bio-char
amendment was believed to be responsible for the enhanced immobilization of Ni and Cd since
a pH increase could lead to the formation of metal (hydro) oxides, carbonates, or phosphate
precipitates (Uchimiya et al., 2010). X. Cao et al. (2009) stated that the sorptive interaction via
the coordination of metal d-electron to π-electron bonds within the bio-char (cation-π
interaction) could be a mechanism for heavy metal retention by bio-char. Some other
researchers argued that the complexation of heavy metals with the surface functional groups
of bio-chars (especially oxygen, phosphorus, sulfur, and nitrogen function groups) may
contribute to enhanced heavy metal immobilization (Jin et al., 2017; Shi et al., 2013b).
However, the detailed mechanism of heavy metal immobilization through co-pyrolysis with
biomass (e.g., wood waste) still needs further investigation.

102

4.3.5. Potential ecological risk assessment
Table B1 in Appendix B shows the concentrations of extractable fractions of heavy metals
in the oily waste (SRC) and its solid residues, which can be used to assess the potential
ecological risk assessment. The potential ecological risk index (RI) posed by heavy metals in
the original oily waste (SRC) was calculated to be 302.50, implying a potentially considerable
ecological risk (e.g., 300 < RI ≤ 600 indicating considerable risk). However, after any pyrolysis
treatment shown in Table 4.2, the risk was reduced from “considerable” to “low” (e.g., RI ≤
150 indicating low risk) (Huang and Yuan, 2016; Jin et al., 2017). Figure 4.7 presents the
relationships between RI of solid reside and the experimental variables. Among all the
experimental variables, pyrolysis temperature displayed the most significant effect on RI. The
higher the pyrolysis temperature, the lower the RI obtained. The effect of hog fuel addition on
RI was also important. The increasing hog fuel addition ratio in the feedstock caused a
gradually decreasing RI value. These observations are in agreement with the results for metal
species variation, in which the transformation from bioavailable category (F1+F2) to less
soluble metal forms was enhanced when the pyrolysis temperature and hog fuel addition
increased.
Additionally, due to the fact that the interaction between pyrolysis temperature and hog
fuel addition was significant (F-value = 30.382), the RI value varied with pyrolysis temperature
depending on hog fuel addition, and vice versa. As can be observed from Figure 4.7 (b), at 20%
hog fuel addition, the pyrolysis temperature had a slight effect on RI, while a low RI (< 55)
could always be achieved at different pyrolysis temperatures (400–600 °C). However, when
the hog fuel addition was zero, RI was decreased from 117.14 to 34.63 as the pyrolysis
temperature increased from 400 °C to 600 °C, indicating a much stronger effect of pyrolysis
103

temperature on RI. The experimental data illustrated that the co-pyrolysis of oily waste with
20% of hog fuel at 400 °C could greatly transform the potentially toxic metal fractions in the
waste into the least soluble fraction, while further increase in pyrolysis temperature showed
little beneficial effect.
It was found that the influence of reaction time on RI was not significant (F-value = 5.761),
which can be also observed from Figure 4.7 (a) and (c). Similarly, very little effect on the
reduction of metal leaching by increased heating time (from 1 h to 2 h) was reported by Debela
et al. (2012). However, L. Li et al. (2012) found that the effect of reaction time differed at
different reaction temperature during the hydrothermal gasification of sewage sludge, and it
also varied with metals. They found that at 375 °C the bioavailable metal fractions (F1+F2) of
Cu and Zn increased with increasing reaction time (from 30 min to 60 min), but no clear change
was observed for Cd, Cr, and Pb (L. Li et al., 2012). In contrast, the bioavailable fractions
(F1+F2) of Cu and Zn decreased at 400 °C as reaction time increased, while those for Cd, Cr
and Pb increased (L. Li et al., 2012). Based on results from this study and those reported in
literatures, the effect of reaction time on the distribution and fraction transformation of heavy
metals during pyrolysis treatment remains unclear and further investigations are needed.

104

RI

a

Re
ac

tio
nt

im

a

e(
tur

e(

mi

rol

n)

Py

ysi

er
mp
s te

ºC)

RI

b

RI

c

Figure 4.7. Response surface graph of risk index (RI) as function of (a) pyrolysis temperature and
reaction time (at hog fuel addition = 10%), (b) pyrolysis temperature and hog fuel addition (at
reaction time = 75 min), and (c) reaction time and hog fuel addition (at pyrolysis temperature =
500 °C).
105

4.4. Summary
In this study, a co-pyrolysis of oily waste with wood waste was examined for its oil recovery
and heavy metal immobilization. A maximum TPH recovery was obtained at a pyrolysis
temperature of 400 °C, but further temperature increase could reduce the potential ecological
risk of heavy metals. It was also observed that a longer reaction time could favor the pyrolysis
conversion, but had little effect on heavy metal immobilization. The addition of hog fuel for
co-pyrolysis reduced TPH recovery, but showed a positive effect on heavy metal
immobilization in solid residue. At 400 °C, the RI of solid residue decreased from 117.14 to
54.12 when the hog fuel addition ratio in the co-pyrolysis feedstock increased from 0% to 20%.
As a result, the co-pyrolysis of oily waste with other biomass waste is an effective approach
for both oil recovery and heavy metal immobilization in the obtained solid residue. However,
further studies may be required to explore the detailed mechanism of heavy metal
immobilization.

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Chapter 5
APPLICATION OF OILY SLUDGE-DERIVED CHAR FOR LEAD AND
CADMIUM REMOVAL FROM AQUEOUS SOLUTION
Abstract
The objective of this study was to investigate the physicochemical properties of oily sludgederived char produced from pyrolysis at 500 °C (OS500) and its sorption behavior towards
lead (Pb2+) and cadmium (Cd2+) in aqueous solution. The sorption kinetics and isotherm of
Pb2+ and Cd2+ sorption on OS500 were determined. The results were fitted with four kinetics
(pseudo first order, pseudo second order, Elovich equation, and intraparticle diffusion) and six
isotherm (Langmuir, Freundlich, Sips, Redlich-Peterson, Temkin, and Dubinin-Radushkevich)
models. The maximum sorption capacity (𝑄𝐿 ) of Pb2+ obtained from the Langmuir model was
373.2 mg/g, while 𝑄𝐿 of Cd2+ was 23.19 mg/g. The mechanisms of lead and cadmium sorption
on OS500 and their quantitative contributions were further studied by Fourier transform
infrared spectroscopy (FTIR), X-ray diffraction (XRD) and sequential extraction test. The
results indicated the chemical characteristics (e.g., alkaline minerals and effective cation
exchange capacity (CEC)) but not microstructure of OS500 controlled its Pb2+ and Cd2+
sorption. New mineral precipitates (i.e., hydrocerussite (Pb3(CO3)2(OH)2) and cerussite
(PbCO3)) formed during Pb2+ sorption. The Pb2+-π interaction and complexation of Cd2+ with
hydroxyl functional groups were evidenced by FTIR. Mineral precipitation with Pb2+ was the
predominant mechanism for Pb2+ sorption on OS500 (accounted for 93.79%), whereas the
complexation dominated Cd2+ sorption (accounted for 84.15%). The OS500 shows application
potential in removing heavy metal contaminants from solution, especially Pb2+.

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Keywords: lead, cadmium, sorption mechanism, oily sludge-derived char (OSDC), speciation
5.1. Introduction
It is estimated that more than 60 million tons of oily sludge is generated annually by oil
refineries worldwide (Hu et al., 2013). Such sludge contains high concentration of heavy
metals (da Rocha et al., 2010) and hazardous organic substances (Mazlova and Meshcheryakov,
1999), such as polycyclic aromatic hydrocarbon (PAHs), phenol, and benzene, posing risks to
human health and the environment (Kuang et al., 2011; Robertson et al., 2007). The effective
disposal of oily sludge is a difficult and challenging problem because of its hazardous and
recalcitrant nature (Hu et al., 2013). Compared to landfilling and incineration, pyrolysis can be
a potential promising method for oily sludge management due to its energy recovery (pyrolysis
oil and gas), solid waste reduction, and less secondary pollution (Qin et al., 2015). In addition
to the oil and gas recovery, solid residue char is generated from pyrolysis. In the past years,
growing interest has been shown in the application of oily sludge-derived char (OSDC) to
remove heavy metals from wastewater. For example, Mohammadi and Mirghaffari (2015)
reported that the cadmium (Cd2+) sorption capacity for the non-activated OSDC was 1.49 mg/g,
compared to that for activated OSDC was 19.23 mg/g. Seredych and Bandosz (2006) produced
the OSDC at 650 °C and found the maximum copper (Cu2+) sorption ability of OSDC (74.28
mg/g) was about five times higher than that of wood-based (15.88 mg/g) activated carbon. As
a result, oily sludge could be a potential material for producing effective sorbent.
Previous work showed that biochar (a carbon-rich solid obtained from pyrolysis of biomass)
can adsorb heavy metals through a range of mechanisms including (i) physical sorption; (ii)
ion exchange (such as K+, Na+, Ca2+, and Mg2+); (iii) surface complexation between metal
cation and oxygen-containing functional groups (e.g., ‒OH and ‒COOH); (iv) precipitations
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with minerals (e.g., CO32‒, SO42‒, and PO43‒); (v) electrostatic interactions between metal
cations and negatively charged biochar surface; and (vi) coordination of metal cation with π
electrons (Lu et al., 2012; Uchimiya et al., 2010). The organic functional groups are the
important components determining the sorption of heavy metals on biochar. For example,
Uchimiya et al. (2011b) showed the properties attributable to surface functional groups (i.e.,
volatile matter, oxygen content, and the point of zero charge) controlled the heavy metal
retention ability of cottonseed hull chars, while other properties such as fixed carbon and ash
content had minor effect on heavy metal retention ability. Tong et al. (2011) reported that the
major sorption mechanism of Cu2+ onto crop straw-derived biochar was surface complexation
between metal and organic functional group. In addition, the inorganic mineral in the biochar
contributes to the heavy metal removal through precipitation and ion exchange. For instance,
T. Zhang et al. (2017) indicated that the lead (Pb2+) sorption of celery-derived biochar involved
about 59.8% to 62.8% ion exchange, 36.6% to 34.9% precipitation, and 2.3% to 3.6%
complexation. Wang et al. (2015) found that the major sorption mechanism of Pb2+ on peanut
shells-derived biochars was precipitation, which accounts for more than 80% of the total
sorption. The above results showed that the sorption capacity and mechanism of metals on
biochar largely depend on the feedstock.
Different from most biomass, such as plant or manure, oily sludge generally contains
abundant minerals, including quartz, montmorillonite, calcite, and barite (Andrade et al., 2009;
Singh et al., 2010). The minerals may be enriched after the pyrolysis process and further
participate in metal sorption. As such, it is reasonable to hypothesize that the mineral-rich
OSDC could remove heavy metals through a complex mechanism, involving surface
adsorption, precipitation and ion exchange. However, there is a lack of information on the roles

109

of different sorption mechanisms for OSDC, which is critical for its field application.
Investigation and even quantification of metal sorption mechanisms of OSDC will aid the
design of future field-scale remediation projects, and estimation of the long-term performance
of OSDC in field conditions.
In this study, OSDC was prepared by pyrolysis of oily sludge at 500 °C in N2 atmosphere.
The microstructure and chemical properties of OSDC were characterized. The objectives of
this study were (i) evaluating the application potential of OSDC as a low-cost sorbent for
removing Pb2+ and Cd2+ in aqueous solution; (ii) investigating the mechanism of Pb2+ and Cd2+
sorption by OSDC using batch sorption studies, Fourier transform infrared spectroscopy
(FTIR), and X-ray diffraction (XRD); and (iii) quantifying the contribution of different
mechanisms to the overall sorption of Pb2+ and Cd2+ by OSDC using a modified sequential
extraction test.
5.2. Materials and methods
5.2.1. Materials and reagents
The oily sludge used in this study was collected from an oil tank truck bottom in western
Canada. The sludge was oven dried at 80 °C for 3-4 days and stored in a sealed glass jar at
24 °C. The properties of oily sludge are listed in Table 5.1. The extractable petroleum
hydrocarbons (i.e., nC10–nC19 and nC19–nC32) was determined by USEPA Method 8015C
(USEPA, 2007b). The metal elements were measured using inductively coupled plasma (ICP)
analysis based on the ASTM D5185 (ASTM, 2018). Sodium nitrate (≥99%) and cadmium
nitrate tetrahydrate (≥98%) were purchased from Sigma-Aldrich. Lead nitrate (≥99%) was

110

purchased from Fisher Scientific. Ultra high purity (UHQ) water with a resistivity of 18.2
MΩ·cm (ELGA Purelab) was used for all procedures.
Table 5.1. Properties of oily sludge

Parameter
Hydrocarbons

Metals

a
b

Concentration (mg/kg)
nC10–nC19a 137,000
nC19–nC32b 99,000
Al

3,491

Ba

15,698

Ca

29,755

Fe

10,379

K

14,747

Mg

13,946

Na

12,835

P

759

Si

7,953

Hydrocarbons that elute between nC10 and nC19.
Hydrocarbons that elute between nC19 and nC32.

5.2.2. Char preparation
About 30 g of oily sludge was put into the sample ark and pyrolyzed in a fixed-bed tube
furnace pyrolysis reactor (quartz tube length: 600 mm, Ø: 50 mm; MTI Corp.® GSL-1100X)
under N2 atmosphere. The furnace temperature was gradually increased to 500 °C at a rate of
10 °C/min, and then maintained for 30 min with a 100 mL/min nitrogen flow rate. Our previous
study on oily sludge co-pyrolysis for energy recovery found the maximum oil yield occurred
in the range of 500–550 °C (Hu et al., 2017). In addition, the preliminary sorption results
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indicated the solid char produced at 500 °C exhibited good performance in removal of heavy
metals from solution. Therefore, the pyrolysis temperature of 500 °C was chosen for this study.
The solid product was then cooled to room temperature under nitrogen flow (100 mL/min) and
then washed with UHQ water in a Soxhlet apparatus for 48 h. Subsequently, the sample was
oven dried at 105 °C for 2-3 days and ground using a ball mill (Retsch MM200). 2 g of the
sample and agate ball (12mm diameter) were mixed in a 10 mL milling cup at 25 Hz for 40
seconds. The resulting sample was referred to as OS500.
5.2.3. Char characterization
The OS500 was analyzed by a Costech ECS 4010 elemental analyzer to obtain the elemental
concentrations of carbon (C), nitrogen (N), hydrogen (H), and sulfur (S). The elements (K, Na,
Ca, Mg, Cd, and Pb) in OS500 were determined by inductively coupled plasma optical
emission spectrometry (ICP-OES) (Agilent Technologies 5100) after digestion in a
nitric/hydrochloric acid mixture (USEPA, 1996b). Ash content was determined after burning
the char to a constant weight in a muffle furnace at 600 °C for 16 h. The pH value of char was
measured in UHQ water at the ratio of 1:50 w/v after being shaken for 24h at 200 rpm on
Talboy 3500 Orbital shaker (Seredych and Bandosz, 2006).
The point of zero charge (pHPZC) of char was determined by the pH drift method (Uchimiya
et al., 2011a). A solution of 5mM CaCl2 was boiled to remove CO2 and cooled to room
temperature. The pH of CaCl2 was adjusted to pH 4, 6, 8, 10, and 12 using 0.5 M HCl or 0.5
M NaOH. OS500 (0.06g) was mixed with 20mL of the pH adjusted CaCl2 solution in 50-mL
centrifuge tube. The mixture was shaken for 24 h at 200 rpm and the final pH was measured.
The pHPZC is defined as the point where the curve of final pH versus initial pH crosses the

112

pHinitial = pHfinal line. The determination of pHPZC for OS500 is shown in Figure C1, Appendix
C.
Total surface acidity (TSA) was determined using the modified Boehm titration method
(Goertzen et al., 2010). Briefly, 1.2 g char was added to 40mL of 0.05 M NaOH solution. After
shaking for 24 h, the char suspension was filtered by a 0.45-µm nylon filter. The resulting
solution (10 mL) was transferred into a 40-mL glass vial, and then 20 mL of 0.05 M HCl
solution was added. The N2 was then bubbled (2–3 bubbles/s) into the vial through a needle
submerged in the solution for 2 h to remove CO2. After degasification, the solution was
transferred into a 125 mL-Erlenmeyer flask which had been purged with the N2 gas and
covered with Parafilm, and then back-titrated with 0.05 M NaOH. The degasification was
continued during back-titration. Total surface acidity was calculated as the consumption of
HCl. Similarly, total surface basicity (TSB) can be determined as the consumption of NaOH.
Phenolphthalein was used in acidic samples titrated with NaOH, and methyl orange was used
in basic samples titrated with HCl.
The effective cation exchange capacity (CEC) of char was measured using the method of
ammonium chloride solution (Schulz and Glaser, 2012). 1 g of char was added to 20 mL 0.5M
NH4Cl. The mixture was rotated end-over-end on a Glass-Col rotator at 37 rpm for 16 h. The
extract was collected by centrifugation at 5000 rpm for 10 min. The extraction procedure was
repeated and the extracts were combined. A total of 40 mL of extracts was collected and then
filtrated through a 0.45-µm nylon filter. The concentration of exchangeable cations (K+, Na+,
Ca2+, Mg2+, Ba2+, Fe2+, Mn2+, and Al3+) in the filtrate were measured by ICP-OES (Agilent
Technologies 5100). The effective CEC was calculated by the sum of exchangeable cations.

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The Brunauer-Emmett-Teller (BET) surface area (SBET) and pore structure of char were
obtained from N2 adsorption-desorption isotherms at –196 °C using Quantachrome Autosorb1 analyzer. Before test, 0.1 g of sample was outgassed at 300 °C for 24h. SBET was calculated
based on BET equation. Total pore volume (PV) was determined by the nitrogen adsorption at
a relative pressure (P/P0) of 0.99. Average pore diameter (DP) and pore size distribution were
calculated by the desorption data using Barrett-Joyner-Halenda (BJH) method (Barrett et al.,
1951).
5.2.4. Batch sorption studies
Pb2+ and Cd2+ were selected as target heavy metals since they are among the most seriously
concerned water pollutants due to their high degree of toxicity, non-biodegradability,
bioaccumulation, and “known” or “probable” carcinogenicity (IARC WGECRH , 2012, 2006;
Tchounwou et al., 2012). The Pb2+ (or Cd2+) solution was prepared by dissolving Pb(NO3)2 (or
Cd(NO3)2·4H2O) in cooled boiled UHQ water. All the tested Pb2+ (or Cd2+) solution contained
0.01 M NaNO3 as the background electrolyte. Before sorption, the pH level was adjusted to
5.0 ± 0.05 using either 0.1 M HNO3 or NaOH solution. Based on previous results, the sorption
of Pb2+ (or Cd2+) on biochar increased with solution pH from 2 to 5 and became steady when
the initial pH exceeded 5 (Cui et al., 2016; Lu et al., 2012; Tan et al., 2018; Wang et al., 2018).
At lower pH, the protonation and competition of H+ for sorption sites inhibited the Pb2+ (or
Cd2+) sorption (Cui et al., 2016; Tan et al., 2018). Therefore, the initial pH of 5 was chosen for
sorption studies.
In order to study the kinetics of Pb2+ (or Cd2+) sorption onto OS500, 0.1 g of OS500 was
placed into a 50-mL centrifuge tubes and mixed with 30 mL of 1600 mg/L Pb2+ (or 1000 mg/L
Cd2+) solution. The mixture was shaken at 200 rpm and room temperature (24 °C) for different
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time (0.125 h, 0.25 h, 0.5 h, 1 h, 2 h, 4 h, 8 h, 16 h, 32 h, 64 h, 128 h, 256 h, and 536 h). After
shaking, all mixtures were filtrated through a 0.45-µm nylon filter and the concentration of
Pb2+ (or Cd2+) in the filtrate was determined by ICP-OES. The final pH values of the filtrate
solutions were recorded using a Mettler Toledo FG2 portable pH meter. The sorption isotherms
were studied using the same procedures as stated before but the concentration of Pb2+ varied
from 50 to 9600 mg/L (or Cd2+ concentration varied from 31 to 1000 mg/L) and shaking time
was 128 h. According to the kinetics sorption results, the sorption amount of Pb2+ (or Cd2+) on
OS500 varied little with time after the contact time exceeded 128 h. Therefore, the shaking
time of 128 h was used in sorption isotherm studies.
All the sorption experiments were performed in triplicate, with appropriate blanks without
OS500 or without heavy metals. The Pb2+ (or Cd2+) sorption capacity of OS500 was calculated
as follows:
𝑄𝑡 =

(𝐶0 −𝐶𝑡 )∗𝑉
𝑊

(5.1)

where 𝑄𝑡 is amount of Pb2+ (or Cd2+) sorbed per unit mass of char at time 𝑡 (mg/g); 𝐶0 is the
initial concentration of Pb2+ (or Cd2+) in aqueous solution (mg/L), 𝐶𝑡 is the concentration of
Pb2+ (or Cd2+) in solution at time 𝑡 (mg/L), 𝑉 is the volume of solution (L), and 𝑊 is the mass
of OS500 (g).
5.2.5. Post-sorption analysis
The equilibrium studies indicated that OS500 reached its maximum sorption capacities of
Cd2+ at initial Cd2+ concentration of 1000 mg/L. The Pb2+ sorption ability of OS500 at initial
Pb2+ concentration of 1600 mg/L was as high as 338.3 ± 7.5 mg/g. Therefore, the OS500 after
Pb2+ (or Cd2+) sorption at initial Pb2+ concentration of 1600 mg/L (or initial Cd2+ concentration
115

of 1000 mg/L) was chosen for the post-sorption analysis to identify sorption mechanisms. The
OS500 after Pb2+ (or Cd2+) sorption was quickly washed by 20 mL UHQ water, followed by
centrifugation (Thermo Scientific Sorvall Legend X1) at 5000 rpm for 10 min. The supernatant
was discarded. The remaining solid was oven dried at 60 °C for 48 h and was then identified
as post-sorption OS500 (OS500-Pb or OS500-Cd). The same procedure was applied to control
sample without the presence of Pb2+ (or Cd2+). The resulting sample was identified as OS500
before sorption (OS500-B).
To investigate the sorption mechanisms qualitatively, XRD and FTIR analysis were
conducted on OS500-Pb, OS500-Cd, and OS500-B. The FTIR spectra was obtained using
Bruker Alpha II spectrometer equipped with an attenuated total reflection (ATR) accessory.
24 scans were taken from 4000 to 370 cm‒1 with a resolution of 4 cm‒1. XRD analysis was
conducted to identify possible crystalline phases. The sample was mounted on a flat sample
holder and examined by Rigaku Miniflex 600 diffractometer with Cu-Kα at 40 kV and 15 mA.
Diffraction patterns were recorded from 3 to 90° (2θ) at a rate of 10 °C/min and a resolution
of 0.02°/step.
The amount of Pb2+ (or Cd2+) sorption resulted from the release of cation (K+, Ca2+, Mg2+,
and Ba2+) was determined by the difference of concentration of cations (K+, Ca2+, Mg2+, and
Ba2+) in solution before and after sorption. The contribution of different mechanisms for Pb2+
(or Cd2+) sorption on OS500 was calculated using the following 6-step sequential extraction
test (X. Li et al., 1995; Tessier et al., 1979):
Step 1 (Water soluble fraction, WS): The solid sample (0.1g OS500 + sorbed Pb2+/or Cd2+)
was added to 20 mL UHQ water. The mixture was shaken at 200 rpm for 24 h at room
temperature. The extract was collected by centrifugation at 5000 rpm for 10 min, filtered

116

through a 0.45-µm nylon filter and stored at 4 °C until further analysis. The remaining solid
was washed by shaking at 200 rpm for 15 min with 20 mL of UHQ water, followed by
centrifugation at 5000 rpm for 10 min. The supernatant was decanted.
Step 2 (Exchangeable fraction, EX): The residue from step 1 was extracted with 8 mL of
0.5 M magnesium chloride (MgCl2) at pH of 7 by shaking at 200 rpm for 20 min at room
temperature. Similar to step 1, the extract was collected by centrifugation and then filtered.
The residue was washed, followed by centrifugal separation as in step 1.
Step 3 (Fraction bound to carbonate, CB): The residue from step 2 was extracted with 8
mL of 1 M sodium acetate (NaOAc; adjusted to pH 5 with acetic acid, HOAc) by shaking at
200 rpm for 5 h at room temperature. The extract was then subject to centrifugal separation
and nylon filter filtration, and the resulting residue was washed as in previous steps.
Step 4 (Fraction bound to Fe-Mn oxides, OX): The residue from step 3 was extracted with
20 mL of 0.04 M hydroxylamine hydrochloride (NH2·OH·HCl) in 25% (v/v) HOAc for 6 h at
96 °C with occasional agitation. The extract solution was diluted to 20 mL with UHQ water
and shaken at 200 rpm for 10 min. Similar to the previous steps, the extract was separated and
filtered, and the residue was washed.
Step 5 (Fraction bound to organic matter, OM): The residue from step 4 was extracted with
3 mL of 0.02 M HNO3 and 5 mL of 30% H2O2 (adjusted to pH 2 with HNO3). The mixture
was heated at 85 °C for 2 h with occasional agitation. A second 3 mL of 30% H2O2 (adjusted
to pH 2 with HNO3) was added and the mixture was heated at 85 °C for 3 h with occasional
agitation. After cooling, 5 mL of 3.2 M ammonium acetate (NH4OAc) in 20% (v/v) HNO3 was
added, followed by dilution to a final volume of 20 mL using UHQ water. The sample was

117

then shaken at 200 rpm for 30 min. The extract was separated, then filtered, and the residue
was washed as in previous steps.
Step 6 (Residual fraction, RS): The residue from step 5 was digested in concentrated HNO3
and concentrated HCl, following the US EPA 3050B (USEPA, 1996b). The resulting solutions
were used to determine the concentration of Pb2+ (or Cd2+).
The concentration of Pb2+ (or Cd2+) in the extract from each step were analyzed by ICPOES (Agilent Technologies 5100). Three independent replicates were performed for each
sample.
5.3. Results and discussion
5.3.1. Characteristics of char
The physicochemical properties of OS500 are summarized in Table 5.2. The nitrogen
adsorption-desorption isotherms and pore size distribution of OS500 are shown in Figure 5.1.
According to international union of pure and applied chemistry (IUPAC), the isotherm of
OS500 could be classified as type II (Sing et al., 1985). The type II represents non-porous or
macroporous sorbent. The hysteresis loop (type H4) between the adsorption and the desorption
isotherm suggests the presence of mesopores (pores in the range 2–50 nm), which is consistent
with the measured average pore diameter (3.63 nm, as listed in Table 5.2). The adsorption
isotherm increased rapidly at higher P/P0 could be due to the capillary condensation in the
mesopores (Ren et al., 2012). The Type H4 loop is usually associated with aggregates of plantlike particles giving rise to silt-shaped pores (Andrade et al., 2009). As shown in Figure 5.1(b),
the pore volume of OS500 was mainly distributed in the mesopore range, with a peak around
3.6 nm. The BET surface area of OS500 was 17.4 ± 0.6 m2/g, which was much smaller than

118

that (112–642 m2/g) of plant-based biochar such as wheat residue, oak wood, or pine needle
(H. Li et al., 2017). This might be because mineral components in oily sludge reduced porosity
when pyrolyzed (Cao and Harris, 2010). The high pH of OS500 (10.75 ± 0.02) could be
attributed to the release of alkali salts from the pyrolytic structure at the temperature of 500 °C
(Shinogi and Kanri, 2003). OS500 had the high pHPZC of 11.57, which was in agreement with
its high ash content since the mineral oxide in ash could increase pHPZC (Godinho et al., 2017).
It can be found from Table 5.2 that OS500 contained a large amount of mineral elements such
as Ca, Mg, K, and Na, with Ca being the highest at 7.259 ± 0.270%. The abundant mineral
elements can be released and replaced by other cations, thus OS500 had a CEC as high as 86.95
± 0.35 cmol/kg.

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Table 5.2. Physicochemical properties of OS500a

OS500
pH

10.75 ± 0.02

pHPZC

11.57

Ash (%)b

90.44 ± 0.03

Ultimate analysis

Metal content

a

6.11 ± 0.04

H (%)b

0.28 ± 0.00

N (%)b

0.08 ± 0.00

K (%)

0.372 ± 0.023

Ca (%)

7.259 ± 0.270

Na (%)

0.044 ± 0.002

Mg (%)

1.347 ± 0.071

Cd (mg/kg)

2.5 ± 0.14

Pb (mg/kg)

38 ± 2.8

Cr (mg/kg)

37.2 ± 1.9

Cu (mg/kg)

66.0 ± 2.8

Ni (mg/kg)

42.0 ± 2.8

Zn (mg/kg)

230.5 ± 17.7

CEC (cmol/kg)

86.95 ± 0.35

TSA (mmol/g)

0.07 ± 0.01

TSB (mmol/g)

4.52 ± 0.10

Microstructures

b

C (%)b

SBET (m2/g)

17.4 ± 0.6

PV (cm3/g)

0.0643

DP (nm)

3.63

Values are given as mean ± standard deviation for duplicate measurement.
On a dry basis.

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a

Volume (cm3/g STP)

40

Adsorption
Desorption

30
20
10
0
0.0

0.2

0.4

0.6

0.8

1.0

Relative pressure (P/P0)
0.007

b

dv/dD (cm3/nm/g)

0.006
0.005
0.004
0.003
0.002
0.001
0.000

1

10

100

1000

Pore diameter (nm)
Figure 5.1. (a) Nitrogen adsorption/desorption isotherms and (b) pore size distribution of the OS500.

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In addition, OS500 contained heavy metals, such as cadmium (2.5 ± 0.1 mg/kg), lead (38 ±
2.8 mg/kg), chromium (37.2 ± 1.9 mg/kg), copper (66.0 ± 2.8 mg/kg), nickel (42.0 ± 2.8 mg/kg),
and zinc (230.5 ± 17.7 mg/kg). As described in our previous study (Tian et al., 2019), the heavy
metals (Cu2+, Cr3+, Ni2+, Pb2+, and Zn2+) in synthetic oily waste were effectively immobilized
after pyrolysis at 400–600 °C. To evaluate the environmental safety of the OS500 when it is
used to sorb heavy metals in solution, the mobility and speciation of heavy metals in OS500B was investigated by the sequential extraction test (the procedures were described in 5.2.5).
Table C1 in Appendix C shows the concentrations of extractable fractions of Cd, Pb, Cr, Cu,
Ni, and Zn in the OS500. The negligible water soluble and exchangeable fraction of these
metals in OS500 indicates OS500 can be used as sorbent in wastewater treatment with no direct
risks to plants and humans. In fact, no Cd (or Pb), Cr, Cu, Ni, and Zn was detected in solution
after Pb2+ (or Cd2+) sorption (initial concentration: Pb2+ 1600mg/L; or Cd2+ 1000 mg/L).
Therefore, the effect of Pb and Cd originating from OS500 could be ignored in the sorption
studies.
5.3.2. Sorption kinetics
The influence of contact time on the Pb2+ or Cd2+ sorption by OS500 is shown in Figure 5.2.
The initial Pb2+ (or Cd2+) sorption rate was fast, with approximately 61% (or 65%) of Pb2+ (or
Cd2+) sorption occurring in the first 4 h, followed by a slow approaching to equilibrium. After
128 h, the rate of sorption became very low and the total sorption amount of Pb2+ (or Cd2+)
varied little with time. There was no distinct “endpoint” to the sorption, even at much longer
times. 128 h was therefore deemed sufficient to establish equilibrium in this study. The
experimental data were further fitted by kinetics models including pseudo first order kinetic
model (PF-order, eq. (5.2)), pseudo second order kinetic model (PS-order, eq. (5.3)), Elovich
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equation (eq. (5.4)), and intraparticle diffusion model (IPD, eq. (5.5)), which can be presented
as follows (Chien and Clayton, 1980; Ho and McKay, 1998; Lagergren, 1898; Weber and
Morris, 1963):
𝑄𝑡 = 𝑄𝑒 (1 − 𝑒 −𝑘1 𝑡 )
𝑘 𝑄2 𝑡

𝑄𝑡 = 1+𝑘2 𝑒𝑄2 𝑡
2 𝑒

1

(5.2)
(5.3)

𝑄𝑡 = 𝛽 ln(𝛼𝛽𝑡 + 1)

(5.4)

𝑄𝑡 = 𝑘3 𝑡 0.5 + 𝑏

(5.5)

where 𝑄𝑡 is amount of Pb2+ (or Cd2+) sorbed at time 𝑡 (mg/g), 𝑄𝑒 is amount of Pb2+ (or Cd2+)
sorbed at equilibrium (mg/g); 𝑘1 is the pseudo first order rate constant (h–1); 𝑘2 is the pseudo
second order rate constant (g/(mg·h)); 𝛼 is the initial sorption rate (mg/(g·h)); 𝛽 is the
desorption constant, related to activation energies for chemisorption (g/mg); 𝑘3 is the
intraparticle diffusion rate constant (mg/(g·h0.5)); 𝑏 is the intercept (mg/g).
The sorption kinetics curves are shown in Figure 5.2 and Figure C2 (Appendix C), and the
fitted model parameters are listed in Table 5.3. The correlation coefficients (R2) of the PSorder model was higher than those of the PF-order model. In addition, the calculated 𝑄𝑒 (339.4
mg/g for Pb2+ and 21.29 mg/g for Cd2+) using PS-order model well matched the experimental
𝑄𝑒 (355.0 ± 8.5 mg/g for Pb2+ and 22.04 ± 2.2 mg/g for Cd2+), indicating the PS-order model
could preferably describe the Pb2+ (or Cd2+) sorption. The PF-order kinetic process can be
considered as a reversible process with an equilibrium being established between liquid and
solid phases while PS-order model assumes that the rate-limiting step is more likely a chemical

123

sorption involving bonding forces through sharing or exchange of ions/electrons between
sorbent and sorbate (Ho and McKay, 1999). The better fitting of the data to the PS-order model
suggests the chemical process (e.g., precipitation, ion exchange, and complexation) may be the
rate-limiting step in the sorption process of Pb2+ and Cd2+. The Elovich equation provided the
best fit to experimental data since the correlation coefficients (R2) of Elovich equation was
higher than other models (Table 5.3). The initial sorption rate (Elovich 𝛼) had value of 3866
mg/(g·h) for Pb2+ sorption on OS500, which is 2 times greater than that for Cd2+. The Elovich
equation is often used to describe predominantly chemisorption on highly heterogeneous
sorbents. It is based on the assumption that the sorption sites increase exponentially with
sorption, which implies a multilayer sorption (Alberti et al., 2012). The Elovich equation fitted
the experimental data well, which is consistent with chemical interaction between active sites
on the OS500 surface and Pb2+ (or Cd2+) in solution. As seen from Figure C2 in Appendix C,
the intraparticle diffusion plot exhibited multi-linear nature. The first linear plot corresponds
to the film diffusion (transport of Pb2+ (or Cd2+) from solution to the exterior surface of OS500)
whereas the second linear plot attributes to the intraparticle diffusion. The IDP model was well
fitted with the data of Pb2+ sorption within 4 h (R2 = 0.955) and Cd2+ sorption within 16 h (R2
=0.887), confirming diffusion was important for Pb2+ and Cd2+ sorption. Both lines did not
pass through the origin and therefore the intraparticle diffusion was not the only control step
of sorption process.

124

500

a

Qt (mg/g)

400
300
Pseudo first order
Pseudo second order
Elovich

200
100
0
0

100

200

300

400

500

600

Time (h)
30

b

25

Qt (mg/g)

20
15

Pseudo first order
Pseudo second order
Elovich

10
5
0
0

100

200

300

400

500

600

Time (h)
Figure 5.2. The influence of contact time on (a) Pb2+ and (b) Cd2+ sorption onto OS500 and the fit of
the data by pseudo first order, pseudo second order, and Elovich equation (0.1g char in 30 mL solution
(0.01 M NaNO3), initial Pb2+ (Cd2+) concentration 1600 (1000) mg/L; room temperature 24 °C; initial
solution pH 5).

125

Table 5.3. The fitting results of Pb2+ and Cd2+ sorption on OS500 using the pseudo first order kinetic
(PF-order), pseudo second order kinetic (PS-order), Elovich, and intraparticle diffusion (IPD) models.

Pb2+

Parameter 1

Parameter 2

R2

PF-order

𝑘1 = 0.8811 h–1

𝑄𝑒 = 321.9 mg/g

0.775

PS-order

𝑘2 = 0.00357 g/(mg·h)

𝑄𝑒 = 339.4 mg/g

0.869

Elovich

𝛼 = 3866 mg/(g·h)

𝛽 = 0.02629 g/mg

0.975

IPDa

𝑘3 = 106.8 mg/(g·h0.5)

𝑏 = 60.11 mg/g

0.955

Experimental 𝑄𝑒 = 355.0 mg/g
Cd2+

PF-order

𝑘1 = 1.700 h–1

𝑄𝑒 = 20.31 mg/g

0.505

PS-order

𝑘2 = 0.1161 g/(mg·h)

𝑄𝑒 = 21.29 mg/g

0.711

Elovich

𝛼 = 1933 mg/(g·h)

𝛽 = 0.5130 g/mg

0.964

IPDb

𝑘3 = 2.394 mg/(g·h0.5)

𝑏 = 10.10 mg/g

0.887

Experimental 𝑄𝑒 = 22.04 mg/g
IPD model was fitted the data of Pb2+ sorption within 4 h.
b
IPD model was fitted the data of Cd2+ sorption within 16 h.
a

5.3.3. Sorption isotherms
The sorption isotherm of Pb2+ and Cd2+ onto the OS500 were simulated using the Langmuir
(L), Freundlich (F), Sips (S), Redlich-Peterson (RP), Temkin (T), and Dubinin-Radushkevich
(DR) models. The mathematical expression of these models are presented as follows (Dubinin
and Radushkevich, 1947; Freundlich, 1906; Langmuir, 1918; Redlich and Peterson, 1959; Sips,
1948; Tempkin and Pyzhev, 1940):
𝑄 𝑏 𝐶

𝐿 𝐿 𝑒
𝑄𝑒 = 1+𝑏𝐶
Langmuir,

(5.6)

𝑄𝑒 = 𝐾𝐹 𝐶𝑒 1⁄𝑛 Freundlich,

(5.7)

𝑒

126

𝑄 𝐾 𝐶 𝑛𝑆

𝑆 𝑆 𝑒
𝑄𝑒 = 1+𝐾
Sips,
𝐶 𝑛𝑆

(5.8)

𝐾𝑅𝑃 𝐶𝑒

Redlich-Peterson,

(5.9)

𝑄𝑒 = 𝑏 ln(𝐾𝑇 𝐶𝑒 ) Temkin,

(5.10)

2
1
))
𝐶𝑒
−2𝐸 2

(5.11)

𝑆 𝑒

𝑄𝑒 =

1+𝛼𝑅𝑃 𝐶𝑒 𝛽𝑅𝑃

𝑅𝑇
𝑇

𝑄𝑒 = 𝑄𝐷𝑅 𝐸𝑥𝑝 [

(𝑅𝑇 ln(1+

] Dubinin-Radushkevich,

where 𝑄𝑒 is amount of Pb2+ (or Cd2+) sorbed at equilibrium (mg/g), 𝐶𝑒 is the equilibrium
concentration of Pb2+ (or Cd2+) (mg/L), 𝑄𝐿 is maximum sorption capacity based on Langmuir
model (mg/g), 𝑏𝐿 is the Langmuir isotherm constant (L/mg), 𝐾𝐹 is the Freundlich isotherm
constant ((mg/g)(L/mg)1/n) related to sorption capacity, 𝑛 is the Freundlich isotherm constant
representing the surface site heterogeneity, 𝑄𝑆 is the maximum sorption capacity based on Sips
model (mg/g), 𝐾𝑆 is the Sips isotherm constant ((L/mg) nS ), 𝑛𝑆 is the Sips heterogeneity
exponent, 𝐾𝑅𝑃 (L/g), 𝛼𝑅𝑃 ((L/mg)1/βRP), and 𝛽𝑅𝑃 are the Redlich-Peterson isotherm constants,
R is the ideal gas constant (8.314 J/mol K), T is the temperature (K), 𝐾𝑇 is the Tempkin
equilibrium binding constant (L/g), 𝑏𝑇 is the Tempkin isotherm constant (kJ/mol), 𝑄𝐷𝑅 is
maximum sorption capacity based on Dubinin-Radushkevich model (mg/g), and 𝐸 is the mean
free energy (kJ/mol).
The sorption isotherms of Pb2+ and Cd2+ onto OS500 are presented in Figure 5.3 and Figure
C3 (Appendix C), and the fitting parameters are listed in Table 5.4. The equilibrium
concentration of Pb2+ was below the ICP detection limit (0.09 mg/L for Pb2+) when the initial
Pb2+ concentrations ranged from 0–400 mg/L, suggesting an irreversible chemisorption
behavior. Langmuir isotherm is based on the assumption that monolayer sorption takes place
127

at specific homogeneous sites, while the Freundlich isotherm assumes sorption occurs on
heterogeneous surfaces. The Sips isotherm is a combination of Langmuir and Freundlich
models, which incorporate three parameters into an empirical equation. At low sorbate
concentrations, it reduces to a Freundlich isotherm; while at high sorbate concentration, it
predicts a monolayer sorption capacity characteristic of the Langmuir isotherm. RedlichPeterson is widely used as a compromise between Langmuir and Freundlich system. It
incorporates three parameters where the exponent 𝛽𝑅𝑃 lies between 0 and 1. It becomes the
Henry’s equation when 𝛽𝑅𝑃 = 0, and becomes the Langmuir isotherm when 𝛽𝑅𝑃 = 1. From the
correlation coefficient, the sorption of Pb2+ was better fitted to Freundlich model (R2 = 0.897)
than Langmuir model (R2 = 0.839), while the sorption of Cd2+ was better fitted to Langmuir
model (R2 = 0.977) than Freundlich model (R2 = 0.825). In general, three parameter isotherms
(Sips and Redlich-Peterson) had higher correlation coefficients (R2 = 0.898–0.986) than two
parameter isotherms (Langmuir and Freundlich) (R2 = 0.825–0.977). This suggests the
heterogeneous surface condition may coexist with monolayer sorption under the experimental
conditions. The maximum sorption capacity (𝑄𝐿 ) of Pb2+ on OS500 based on Langmuir model
was 373.2 mg/g (Table 5.4), which was larger than that of other sorbents, such as plant-based
biochar (8.1–113.6 mg/g), manure-based biochar (51.4–230.7 mg/g), and sludge-based char
(18.2–82.5 mg/g) (Table 5.5). The maximum sorption capacity (𝑄𝐿 ) of Cd2+ on OS500 based
on Langmuir model was 23.19 mg/g (Table 5.4), which is comparable to that of other sorbents
(7.4–76.3 mg/g) (Table 5.5). The results clearly demonstrated the potential of OS500 to be
used as an effective sorbent for Pb2+ and Cd2+ sorption.
Temkin isotherm contains a factor that takes into account the sorbent-sorbate interactions
(Tempkin and Pyzhev, 1940). It assumes the decline of the heat of sorption as a function of

128

temperature is linear rather than logarithmic (Tempkin and Pyzhev, 1940). A slightly better fit
of Temkin isotherm was found for Pb2+ sorption (R2 = 0.899) compared to that for Cd2+ sorption
(R2 = 0.847), implying that Pb2+ sorption process was more likely to be affected by the
sorbate/sorbent interaction. Dubinin-Radushkevich isotherm is usually applied to distinguish
the physical and chemical sorption of metal ions by mean free energy (𝐸) (Saltalı et al., 2007).
The 𝐸 values of Pb2+ and Cd2+ are calculated as 6.551 and 6.733 kJ/mol, respectively, which
are lower than 8 kJ/mol, showing the physical sorption (Saltalı et al., 2007). However, in the
case of sorption of Pb2+ on OS500, a relatively lower R2 was found with the DubininRadushkevich model as compared to those of other models (Table 5.4), indicating the
application of this model in Pb2+ sorption is limited.

129

400

a

Qe (mg/g)

360
320

Langmuir
Freundlich
Sips
Redlich-Peterson

280
240
0

2000

4000

6000

8000

10000

Ce (mg/L)
24

Model
Equation
Plot
a
b
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

LangmuirT (User)
y=a*b*x/(1+b*x)
Qe
373.19033
19.77826
614.93275
0.83945
0.79931

Model
Equation
Plot
Kf
n
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

Freundlich (User)
Kf*x^(1/n)
Qe
273.81433
25.08163
393.80285
0.89718
0.87148

Model
Equation
Plot
Klf
nlf
alf
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

LFsips (User)
(Klf*x^nlf)/(1+(alf*x)^nlf)
Qe
472.32798
0.07654
0.01177
513.06631
0.89953
0.83256

Model
Equation
Plot
K
a
b
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

RedlichPeterson (User)
(K*x)/(1+a*x^b)
Qe
62493.79342
221.91343
0.96368
521.75065
0.89783
0.82972

Model
Equation
Plot
A
b
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

Temkin (User)
(8.314*297.15)/(b)*(Ln(A*x))
Qe
4.32493E9
198.01794
386.58714
0.89907
0.87383

b

Qe (mg/g)

22
20
Langmuir
Freundlich
Sips
Redlich-Peterson

18
16
14
0

200

400

600

800

1000

Ce (mg/L)
Figure 5.3. Isotherms of (a) Pb2+ and (b) Cd2+ sorption onto OS500 (0.1g char in 30 mL solution (0.01
M NaNO3); contact time 128 h; room temperature 24 °C; initial solution pH 5).

130

Table 5.4. Langmuir (L), Freundlich (F), Sips (S), Redlich-Peterson (RP), Temkin (T), and DubininRadushkevich (DR) isotherm parameters of Pb2+ and Cd2+ sorption on OS500.

Pb2+

Parameter 2

L

𝑄𝐿 = 373.2 mg/g

𝑏𝐿 = 19.78 L/mg

F

𝐾𝐹 = 273.8 (mg/g)(L/mg)1/n 𝑛 = 25.08

S

𝑄𝑆 = 663.5 mg/g

𝐾𝑆 = 0.7119 (L/mg)ns

𝑛𝑆 = 0.07655 0.900

RP

𝐾𝑅𝑃 = 6.249 × 104 L/g

𝛼𝑅𝑃 = 221.9 (L/mg)1/βRP

𝛽𝑅𝑃 = 0.9637 0.898

T

𝐾𝑇 = 4.325 × 109 L/g

𝑏𝑇 = 0.1980 kJ/mol

0.899

𝐸 = 6.551 kJ/mol

0.839

DR 𝑄𝐷𝑅 = 373.2 mg/g

Parameter 3

R2

Parameter 1

0.839
0.897

𝑄𝑚𝑎𝑥,𝑒𝑥𝑝 a = 373.1 mg/g
Cd2+ L

𝑄𝐿 = 23.19 mg/g

0.977

𝑏𝐿 = 21.86 L/mg

F

𝐾𝐹 = 18.35 (mg/g)(L/mg)1/n 𝑛 = 23.64

S

𝑄𝑆 = 23.28 mg/g

𝐾𝑆 = 14.88(L/mg)ns

𝑛𝑆 = 0.8450

RP

𝐾𝑅𝑃 = 607.0 L/g

𝛼𝑅𝑃 = 27.49 (L/mg)1/βRP

𝛽𝑅𝑃 = 0.9905 0.986

T

𝐾𝑇 = 1.532 × 109 L/g

𝑏𝑇 = 2.832 kJ/mol

0.847

𝐸 = 6.733 kJ/mol

0.977

DR 𝑄𝐷𝑅 = 23.19 mg/g

0.825
0.981

𝑄𝑚𝑎𝑥,𝑒𝑥𝑝 = 23.24 mg/g
a

𝑄𝑚𝑎𝑥,𝑒𝑥𝑝 is experimental maximum sorption capacity, mg/g.

131

Table 5.5. Summary of the literature data on sorption of Pb2+ and Cd2+ to different biochars or other sorbents

Sorbents

Metal
ions

Pyrolysis
conditions
Tpa tpb
(°C) (min)

Salisbury biochar

Pb2+

600

810

Wheat Straw pellets biochar

Pb2+

700

Rice husk biochar

Pb2+

Soft wood pellets biochar

Experiment conditions
QLc

C0d
(mg/L)

pH

Te
(K)

Reference

47.6

Sorbent
dosage
(g/L)
5

20–1000

5

293

(Shen et al., 2017)

NAf

113.6

5

20–1000

5

293

(Shen et al., 2017)

700

NA

35.0

5

20–1000

5

293

(Shen et al., 2017)

Pb2+

550

NA

8.1

5

20–1000

5

293

(Shen et al., 2017)

Magnetic oak bark biochar

Pb2+

450

30.2

3

1–100

5

298

(Mohan et al., 2014)

Phyllostachys pubscens
biochar
Dairy manure biochar

Pb2+

700

fast
pyrolysis
180

62.1

2

100–800

5

298

(C. Zhang et al., 2017)

Pb2+

350

240

88.2

5

0–1037

NA

298

(X. Cao et al., 2009)

Pig manure biochar

Pb2+

600

NA

230.7

10

207–5175

5

293

(Kołodyńska et al., 2012)

Anarobically digested animal
waste biochar
Sewage sludge biochar with
KOH activation
Municipal wastewater sludge
char
Chinese medicine material
residues biochar

Pb2+

600

120

51.4

2

5–600

NA

295

(Inyang et al., 2012)

Pb2+

600

20

49.7

2

10–160

5

293

(Lin et al., 2012)

Pb2+

400

120

18.2

2

10–150

5

298

(Zhang et al., 2013)

Pb2+

400

120

82.5

5-7.5

100–1000

5

298

(Wang et al., 2015)

(mg/g)

132

Sorbents

Metal
ions

Pyrolysis
conditions
Tpa tpb
(°C) (min)

Miscanthus biochar

Cd2+

500

60

Phyllostachys pubescens
biochar
Magnetic oak bark biochar

Cd2+

700

Cd2+

450

Corncob biochar

Cd2+

Rape straw biochar

Experiment conditions
QLc

C0d
(mg/L)

pH

Te
(K)

Reference

13.2

Sorbent
dosage
(g/L)
1

1–50

7

298

(W.-K. Kim et al., 2013)

180

14.7

2

50–600

5

298

(C. Zhang et al., 2017)

7.4

3

1–100

5

298

(Mohan et al., 2014)

550

fast
pyrolysis
120

33.4

2

20–200

6.8

298

(Luo et al., 2018)

Cd2+

600

60

32.7

1.25

2–200

5.5

298

(B. Li et al., 2017)

Dairy manure biochar

Cd2+

350

240

54.6

5

0–562

NA

298

(Xu et al., 2013)

Pig manure biochar

Cd2+

600

NA

117.0

10

112–2810

5

293

(Kołodyńska et al., 2012)

Sewage sludge biochar

Cd2+

500

60

15.1

10

0.5–252

NA

303

Sewage sludge biochar with
ZnCl2 activation
Biogas production residue char

Cd2+

850

NA

16.7

5

0–100

5.8

298

(Gutiérrez-Segura et al.,
2012)
(Zhai et al., 2004)

Cd2+

600

300

76.3

4

5–500

5.5

295

(Bogusz et al., 2017)

(mg/g)

Tp is the pyrolysis temperature, °C.
tp is the pyrolysis reaction time, min.
c
QL is the maximum sorption capacity fitted by Langmuir model, mg/g.
d
C0 is the initial concentration of Pb2+ (Cd2+) solution, mg/L.
e
T is the temperature of sorption experiment, K.
f
NA represents the data was not obtained.
a

b

133

5.3.4. Post-sorption characteristics and sorption mechanisms
5.3.4.1. XRD and FTIR results
Figure 5.4 presents the XRD patterns of OS500-B, OS500-Pb and OS500-Cd. The peak
centred at 2θ value of 26.640° in OS500-B was assigned to quartz (SiO2), while that at 25.842°
was attributed to barite (BaSO4). A small peak at 8.854° indicated the presence of muscovite
2M2 (K0.77Al1.93(Al0.5Si3.5)O10(OH)2). The peak at 29.400° and 31.007° showed the existence
of calcite (CaCO3) and dolomite (CaMg(CO3)2), respectively. The presence of the carbonate
minerals is consistent with the alkalinity of the OS500 (Table 5.2). Compared with the OS500B, XRD evidenced the disappearance of peaks for CaCO3 (2θ = 29.400°) in OS500-Pb. The
new peaks of hydrocerussite (Pb3(CO3)2(OH)2) at 27.080° and 34.155° was observed in
OS600-Pb. In addition, the characteristic peaks belonging to cerussite (PbCO3) at 24.792°,
29.057°, 34.611°, 43.469°, and 47.016° were appeared in OS500-Pb. These precipitates formed
in OS500-Pb may be attributed to the reactions between Pb2+ and the dissolved carbonate from
the minerals in OS500. There was no change in XRD patterns before and after Cd2+ sorption,
providing no evidence of precipitation during Cd2+ sorption process.

134

Figure 5.4. XRD patterns of OS500 before (OS500-B) and after Pb2+/Cd2+ sorption (OS500-Pb and
OS500-Cd). Minerals with peaks labeled: M, Muscovite 2M2, K0.77Al1.93(Al0.5Si3.5)O10(OH)2; B, Barite,
BaSO4; Q, Quartz, SiO2; C, Calcite, CaCO3; D, Dolomite, CaMg(CO3)2; Ce, Cerussite, PbCO3; H,
hydrocerrusite, Pb3(CO3)2(OH)2.

Figure 5.5 shows the changes in solution pH during Pb2+/Cd2+ sorption by OS500. When
the initial Pb2+ concentration was 832 mg/L, solution pH increased from about 5 to about 7 due
to the alkalinity of OS500. When the initial Pb2+ concentration was greater than 3543 mg/L,
the solution pH stayed unchanged or became lower. XRD results confirmed that Pb2+ can
precipitate with the negatively charged ion (i.e., carbonate) released by OS500. Even though
Pb2+ precipitation might release H+ (transformation of H2O into OH–), the carbonate released
by OS500 could still increase solution pH due to its strong buffer capacities (Inyang et al.,
2012). However, when the initial Pb2+ concentration was high, all the available carbonate was
consumed and therefore the solution pH decreased. In terms of Cd2+ sorption system, such

135

decreasing solution pH was not observed, suggesting the precipitating may not be the dominant
sorption mechanism.
8

a

6

6

5

5

4
3

3
2

1

1

832

1637

2666

3543

5363

Initial pH
Final pH

4

2

0

b

7

Solution pH

Solution pH

7

8

Initial pH
Final pH

0

10597

47

66

133

256

511

1017

2+

2+

Initial Cd concentration (mg/L)

Initial Pb concentration (mg/L)

Figure 5.5. Comparisons between solution pH before and after (a) Pb2+ sorption and (b) Cd2+ sorption
in equilibrium studies.

The FTIR spectra of OS500-B, OS500-Pb, and OS500-Cd are shown in Figure 5.6. For
OS500-B, the weak peak at 3385 cm‒1 represented the stretching vibration of O–H groups. The
C–O vibration bands at 1417 cm‒1 and 694 cm‒1 were assigned to CO32‒ groups (T. Zhang et
al., 2017). The strong band at 1002 cm‒1 may be attributed to either Si–O–Si or Si–O–C groups
(Pan et al., 2011), which is associated with the silicon content in oily sludge (Table 5.1). The
peaks appearing at 874 cm‒1 and 777 cm‒1 correspond to the aromatic C–H groups (Uchimiya
et al., 2011b). After Pb2+ sorption, the peak of CO32‒ at 1417 cm‒1 and 694 cm‒1 shifted to 1397
cm‒1 and 679 cm‒1, respectively. Meanwhile, the intensity of the CO32‒ group increased after
Pb2+ sorption. These results were attributed to the formation of lead carbonates, which is agreed
with the observation of XRD analysis. The aromatic C–H stretching vibration, located at 874
cm‒1 was found shifting to 837 cm‒1 after Pb2+ sorption, suggesting the Pb2+-π interaction might
also be responsible for Pb2+ sorption. FTIR analysis evidenced no precipitation formation
136

during Cd2+ sorption since no changes in any peak positions except the peak of O–H groups at
3385 cm‒1. The peak at 3385 cm‒1 corresponding to O–H groups shifted to 3366 cm‒1 in
OS500-Cd, indicating a possible complexation of Cd2+ with O–H group on the OS500 surface.

Figure 5.6. FTIR spectra of OS500 before (OS500-B) and after Pb2+/Cd2+ sorption (OS500-Pb and
OS500-Cd).

5.3.4.2. Sequential extraction results
The speciation of Pb2+ (Cd2+) on OS500-Pb (OS500-Cd) are shown in Table C2 (Appendix
C) and Figure 5.7. The total recovery of Pb2+ (or Cd2+) from sequential extraction can be
obtained by dividing the total extracted Pb2+ (or Cd2+) in sequential extraction by the total
sorbed amount of Pb2+ (or Cd2+) and are listed in Table C3 (Appendix C). The recovery for
Pb2+ and Cd2+ were 88.03% and 80.92%, respectively, which were comparable to the recovery
of heavy metals using sequential extraction methods in previous studies. A recovery range of
137

82.4–98.7% was observed for sequential extraction of Cd2+ from two Cd-loaded woody
biochars (Frišták et al., 2015). The sequential extraction recovery of Pb2+ from three biochars
after Pb2+ sorption was within 78.41–85.52% (Shen et al., 2017).

80

100

a

Cd percentage in each fraction (%)

Pb percentage in each fraction (%)

100

OS500-Pb

60
40
20
0

WS

EX

CB

OX

OM

80

b

OS500-Cd

60
40
20

RS

0

WS

EX

CB

OX

OM

RS

Figure 5.7. Sequential extraction results for (a) OS500-Pb, and (b) OS500-Cd (WS: water soluble
fraction, EX: exchangeable fraction, CB: fraction bound to carbonate, OX: fraction bound to Fe-Mn
oxides, OM: fraction bound to organic matter, RS: residual fraction).

The net amount of released cations (K+, Ca2+, Mg2+, and Ba2+) during Pb2+ and Cd2+ sorption
by OS500 were calculated and shown in Table 5.6. The sum of net amount of released cations
into Pb2+ sorption system was 281.45 ± 1.35 meq/100g, equivalent to 291.58 ± 1.40 mg Pb2+/g,
which was accounted for 83.0% of the total sorbed Pb2+. Whereas for Cd2+ sorption system,
the sum of net amount of released cations was only 9.98 ± 0.36 mg Cd2+/g, accounting for 41.7%
of the total sorbed Cd2+. However, as seen from Figure 5.7, the exchangeable fraction for
OS500-Pb and OS500-Cd was low (0.45–15.12%) and their water soluble fraction was
negligible (0–0.11%). This means that most of the sorbed Pb2+ or Cd2+ were non-exchangeable
with Mg2+. The Pb2+ or Cd2+ substituted with the cations was mainly attributed to coprecipitation and complexation on mineral surface.
138

Table 5.6. The release of K+, Ca2+, Mg2+, and Ba2+ during Pb2+ and Cd2+ sorption by OS500

The net amount of released cations (meq/100g)a

Sum

K+

Ca2+

Mg2+

Ba2+

(meq/100g)

(mg/g)

Pb2+
sorption

0.27 ±
0.01

229.64 ±
0.79

48.75 ±
0.46

2.78 ±
0.11

281.45 ±
1.35

291.58 ±
1.40

Cd2+
sorption

0.02 ±
0.02

13.19 ±
0.54

4.32 ±
0.11

0.23 ±
0.02

17.75 ±
0.65

9.98 ±
0.36

System

The net amount of released cations were the difference values between the concentration of K+, Ca2+,
Mg2+, and Ba2+ released into Pb2+ (or Cd2+) sorption system at 1600 mg/L Pb2+ (or 1000 mg/L Cd2+)
solution with 0.01M NaNO3 at pH 5.0 and the corresponding release from OS500 in 0.01M NaNO3 at
pH 5.0.
a

The Pb2+ or Cd2+ sorbed on char through physical sorption represent the water soluble
fraction, whereas those sorbed through cation exchange with a weak binding strength (no metal
exchange reaction) represent the exchangeable fraction (Cao et al., 2019). The fraction bound
to carbonate may come from the Pb2+ or Cd2+ sorbed via cation-π interaction or the formation
of precipitates and complex (Morera et al., 2001; Shen et al., 2017). The fractions from step 4,
5, 6 in sequential extraction may result from the formation of stable complexes with organic
matter or precipitates that cannot be dissolved in the NaOAc/HOAc solution (Shen et al., 2017).
For both OS500-Pb and OS500-Cd, the water soluble fraction was negligible (0–0.11%) and
the total fraction from step 4, 5, 6 was low (0.63–1.50%). This result indicated that physical
sorption and complexation with organic matter had a limited impact on Pb2+ or Cd2+ sorption,
which is in accordance with the fact that OS500 has a small BET surface area (i.e., 17.4 ± 0.6
m2/g) and a small amount of acidic surface function group (i.e., 0.07 ± 0.01 mmol/g) (Table
5.2). The fraction bound to carbonate (98.06%) dominated the existence of Pb2+ in OS500-Pb
while Cd2+ was mainly bound to carbonate (84.15%) and exchangeable fraction (15.12%)
(Table C2). Among the amount of Pb2+ existed as carbonate fraction (303.42 mg/g, as listed in
139

Table C2), 290.20 mg/g (= sum of net amount of released cations in Pb2+ sorption (291.58
mg/g, as listed in Table 5.6) ‒ the amount of Pb2+ existed as exchangeable fraction (1.38 mg/g,
as listed in Table C2)) were due to precipitation and the rest (13.22 mg/g) could be attributed
to Pb2+-π interaction. Therefore, the amount of Pb2+ sorption resulting from precipitation was
estimated to be 290.20 mg/g, accounted for 93.79% of total sorption of Pb2+ on OS500. The
contribution of Pb2+-π interaction was estimated to be 13.22 mg/g, accounted for 4.27% of total
sorption of Pb2+ on OS500. Since no precipitation was formed during Cd2+ sorption according
to XRD and FTIR results, Cd2+ sorption primarily involved the surface complexation
(accounted for 84.15%, representing carbonate fraction) and cation exchange with a weak
binding strength (accounted for 15.12%, representing exchangeable fraction).
5.4. Summary
The oily sludge-derived char from pyrolysis at 500 ˚C (OS500) sorbed up to 373.2 mg/g of
Pb2+ from aqueous solution, which is superior to most biochars reported previously. The
sorption capacity of Pb2+ by OS500 was higher than that of Cd2+, which is mainly due to
mineral precipitation with Pb2+. The kinetics of sorption of Pb2+ and Cd2+ can be well described
by the Elovich equation, indicating the chemical interaction between active sites on the OS500
surface and Pb2+ (or Cd2+) in solution. The Redlich-Peterson isotherm model best fitted the
experimental data with highest correlation coefficients (R2). The OS500 was rich in alkaline
minerals and had a high CEC, which contributed to its high sorption capacity. The substitution
of Pb2+ with alkali or alkaline earth metals led to precipitation in the forms of hydrocerussite
(Pb3(CO3)2(OH)2) and cerussite (PbCO3) was mainly responsible for Pb2+ sorption, with less
contribution from Pb2+-π interaction. The sorption of Cd2+ on OS500 was mainly attributed to
complexation and cation exchange. Microstructure of OS500 played a less important role in
140

the removal of Pb2+ and Cd2+ since only 0–0.11% of total sorbed Pb2+ or Cd2+ existed as
physically bonded fraction. Overall, converting oily sludge into sorbents can be considered a
promising solution to solve disposal problem with the beneficial use of mineral constituents in
the sludge. However, the majority of Pb2+ (98.06%) and Cd2+ (84.15%) was sorbed on OS500
as the carbonate fraction, which may bring challenges for the metal recovery and reuse of
OS500. Further study to facilitate the practical application of OS500 is needed in future.
Moreover, very limited research tried to establish the relationship among properties of OSDC,
metal sorption behavior, and pyrolysis conditions. Therefore, investigating the exact
mechanisms for Pb2+ and Cd2+ sorption on OSDC prepared at different pyrolysis temperature
and activation methods is also necessary.

141

Chapter 6
SORPTION OF LEAD AND CADMIUM ON OILY SLUDGE-DERIVED CHAR
PREPARED AT DIFFERENT PYROLYSIS TEMPERATURES AND CHEMICAL
ACTIVATION
Abstract
The objective of this study was to investigate the lead (Pb2+) or cadmium (Cd2+) sorption
from aqueous solution by oily sludge-derived char (OSDC) prepared at different pyrolysis
temperatures and chemical activation. The maximum Pb2+ sorption capacity for OSDC at
pyrolysis temperature of 500 ˚C (OS500) was found as 351.48 mg/g, which was greater than
that of chars produced at other temperatures. Further investigations of Pb2+ sorption onto
OSDC by low pyrolysis temperature (≤500 °C) indicated that the heavy metal removal was
mainly through a precipitation mechanism, which was confirmed by qualitative (i.e., Fourier
transform infrared spectroscopy (FTIR), X-ray diffraction (XRD)) and quantitative (i.e.,
sequential extraction test) characterization of Pb2+-loaded chars. With increasing pyrolysis
temperature from 500 to 900 °C, the contribution of precipitation to Pb2+ sorption decreased
from 93.79% to 34.63%, while the contribution of complexation increased from 0% to 44.68%.
Sorption results showed that OSDC prepared at pyrolysis temperature of 700 ˚C (OS700) had
the best sorption capacity for Cd2+ (92.14 mg/g). The high sorption capacity of OS700 was
mainly attributed to the carboxyl/hydroxyl functional groups and complexation with mineral
oxides. The complete conversion of barite (BaSO4) to witherite (BaCO3) after chemical
activation led to the mineral precipitation with Cd2+. The contribution of precipitation to the
total Cd2+ removal was 0% for OS500, but 76.12% for the activated OS500.

142

Keywords: heavy metals, oily sludge-derived char (OSDC), pyrolysis temperature, chemical
activation, sorption mechanism
6.1. Introduction
One of the consequences of rapid industrialization and urbanization is the widespread
contamination of water and soil with heavy metals (Gong et al., 2018). Heavy metals pose a
risk to environment and human health due to their toxicity, non-biodegradability, and
bioaccumulation in the food chain. Most of the point sources of heavy metal pollutants are
industrial wastewater from mining, smelting, battery manufacturing, pesticide application, and
others (Inyang et al., 2012). Lead is the most systemic toxicant that affects several organs in
the body such as the kidneys, liver, and central nervous system. Cadmium is one of the most
toxic heavy metals and can cause serious damage to kidneys. The US Environmental Protection
Agency (USEPA) established a maximum permissible limit of 0.015 mg/L for lead and 0.005
mg/L for cadmium in drinking water (USEPA, 2015). Some methods have been used for the
removal of heavy metals from wastewater, including chemical precipitation (Djedidi et al.,
2009), ion exchange (Cavaco et al., 2007), and membrane filtration (Doke and Yadav, 2014).
However, these methods could be associated with high operational cost, disposal of resulting
sludge, and/or possible production of secondary toxic compounds (Fu and Wang, 2011).
Sorption has become an attractive alternative for the removal of heavy metals from
wastewater due to its simplicity, ease of operation and handling, and regeneration capacity
(Saha and Sarkar, 2012). In recent years, growing interest has been shown in the production of
low-cost and efficient sludge-based sorbents (Xu et al., 2015). Sewage sludge-derived char, a
solid by-product of pyrolysis, was reported to have a high capacity for adsorbing copper (Cu2+)
from wastewater due to the complexity of surface chemistry and high specific porosity
143

(Seredych and Bandosz, 2006). Lu et al. (2012) reported the sewage sludge-derived char
effectively removed lead (Pb2+) with the capacity of 30.88 mg/g at initial pH 5. It was found
that the Pb2+ sorption primarily involved the precipitation and complexation on mineral
surfaces, which was 57.7–61.8% of the total sorbed Pb2+ (Lu et al., 2012). Apart from sewage
sludge, industrial oily sludge was also used as material for developing efficient sorbent. For
example, Mohammadi and Mirghaffari (2015) produced oily sludge-derived char (OSDC)
from pyrolysis of oily sludge with the cadmium (Cd2+) sorption capacity of 1.49 mg/g. To
further improve the sorption capability of OSDC, thermochemical treatment was used to
prepare the activated OSDC, with cadmium (Cd2+) sorption capacity reaching 19.23 mg/g. My
study in Chapter 5 also indicated that OSDC rich in alkaline minerals had higher Pb2+ sorption
capacity (373.2 mg/g) than most biochars reported previously. However, limited information
is available on how pyrolysis temperature could affect the heavy metal sorption by OSDC. In
fact, pyrolysis temperature not only affects product distribution but also the structures and
chemical characteristics of the biochar, including the chemical composition, pH, surface area,
and acidic functional group (Chen et al., 2011). Kloss et al. (2012) observed an accumulation
of carbon and increased surface area, but decreased cation exchange capacity (CEC) of biochar
with increase in pyrolysis temperature. Ahmad et al. (2012) reported better sorbent properties
of biochar produced at higher temperature as a consequence of the higher surface area and
lower polarity.
In addition to pyrolysis temperature, a modification of biochar with chemical reagents could
also affect the porous structure development as well as the functional group, and consequently
change its sorption ability (Liu et al., 2012; Trakal et al., 2014). A few studies were reported
to prepare activated OSDC using chemical activation. For example, Guritno et al. (2016)

144

synthesized porous carbon from oily sludge using mesoporous silica template and found the
surface area of the produced activated carbon was 250.6 m2/g. Wang et al. (2017b) found the
potassium hydroxide (KOH)-activated OSDC had a surface area of 1103 m2/g. After a solvent
de-oiling step before activation, the surface area of activated OSDC increased to 3292 m2/g.
However, these studies were more focused on microstructure of activated OSDC, and the
utilization of activated OSDC in heavy metal removal has not been paid much attention. The
objective of this study is then to examine how the preparation condition (i.e., pyrolysis
temperature and chemical activation) and resulting OSDC properties translate into the heavy
metal sorption capacities. The sorption mechanisms of heavy metal on activated and nonactivated OSDC were investigated by qualitative methods (i.e., Fourier transform infrared
spectroscopy (FTIR) and X-ray diffraction (XRD)) and quantitative method (i.e., sequential
extraction). The results will serve to facilitate the industrial application of the developed OSDC.
6.2. Materials and methods
6.2.1. Materials and reagents
The oily sludge was obtained from the bottom of an oil tank truck in western Canada. The
sludge was oven dried at 80 °C for 3-4 days and kept in darkness at 24 °C in a sealed glass jar.
The sludge was sticky, and its characteristics are listed in Table 5.1. The test methods for
extractable petroleum hydrocarbons (i.e., nC10–nC19 and nC19–nC32) and metal elements in
oily sludge can be found in Section 5.2.1. Sodium nitrate (≥99%) and cadmium nitrate
tetrahydrate (≥98%) were purchased from Sigma-Aldrich. Lead nitrate (≥99%) was purchased
from Fisher Scientific. Sodium hydroxide (≥99%) and ferric sulfate (>99%) were purchased

145

from VWR International. Ultra high purity (UHQ) water with a resistivity of 18.2 MΩ·cm
(ELGA Purelab) was used for all procedures.
6.2.2. Chars preparation and characterization
About 30 g of oily sludge were pyrolyzed at 300, 500, 700, and 900 °C for 30 min under 100
mL/min nitrogen flow rate using a fixed-bed tube furnace (quartz tube length: 600 mm, Ø: 50
mm; MTI Corp.® GSL-1100X). The pyrolysis heating rate was 10 °C/min. The solid product
was cooled to room temperature, then washed with UHQ water in a Soxhlet apparatus for 48h,
dried at 105 °C for 2-3 days, and ground using a ball mill (Retsch MM200). The resulting chars
were referred to as OS300, OS500, OS700, and OS900 respectively, according to the pyrolysis
temperature.
The chemical activated char (OS500N900) was prepared according to modified method from
Wang et al. (2017b). The oily sludge was impregnated with sodium hydroxide (NaOH) solution
in a beaker with water:NaOH:oily sludge equal to 1:1:1 by mass. The mixture was dehydrated
in an oven overnight at 105 °C and then pyrolyzed at 500 °C for 30 min under nitrogen
atmosphere. The product was heated directly from 500 °C to 900 °C at 10 °C/min and held for
80 min for the activation treatment. The activated char was then washed, dried, and ground
similar to non-activated sample.
Hydrothermal treatment of char with NaOH solution was performed based on modified
method from Visa et al. (2012). About 10 g of OS500 (without wash and grinding) was added
to 100 mL of 2M or 8M NaOH solution. The slurries were then refluxed with stirring at
atmospheric pressure and 100 °C for 24 h. At the end of activation period, the product was
collected, washed, dried, and ground. The resulting char samples were hereby denoted as
OS500N100-2M and OS500N100-8M respectively, according to the NaOH solution
146

concentration. For comparison, the activation (with 2M NaOH) was conducted using the same
procedures as stated above but the reaction was conducted at room temperature (24 °C). The
activated sample obtained was denoted as OS500N24-2M.
The preparation of ferric-activated char (FSOS500) followed the method of Yang et al.
(2016). About 20 g of oily sludge was mixed with 10 g of ferric sulfate (FeSO4). The mixture
was then pyrolyzed in the same conditions as those used in the OS500 preparation. After
cooling, the solid product was washed, dried, and ground before use.
The characteristics of chars were determined and the analytical methods can be found in
Section 5.2.3.
6.2.3. Preliminary test
Sorption experiment was conducted in 50-mL centrifuge tubes by mixing 0.1 g of nonactivated char or activated char with 30 mL of 1000 mg/L Cd2+, Cu2+, Pb2+, or Zn2+ solution.
The mixture was shaken at 200 rpm and room temperature (24 °C) for 48 h. After shaking, all
mixtures were subject to nylon filter (0.45-µm) filtration and the concentration of metal (i.e.,
Cd2+, Cu2+, Pb2+, or Zn2+) in the filtrate was determined by ICP-OES. The initial and final pH
values of solutions were recorded by a Mettler Toledo FG2 portable pH meter. The Cd2+, Cu2+,
Pb2+, or Zn2+ sorption capacity of char or activated char was calculated by eq. (5.1).
6.2.4. Post-sorption analysis
To keep consistency with Chapter 5, the OSDC after Pb2+ (or Cd2+) sorption at initial Pb2+
concentration of 1600 mg/L (or initial Cd2+ concentration of 1000 mg/L) was chosen for the
post-sorption analysis. The Pb2+ or Cd2+ solution contained 0.01 M NaNO3 as the background
electrolyte and the pH of solution was pre-adjusted to 5.0 ± 0.05. The shaking time of 128 h
147

was used to achieve equilibrium in sorption experiments. The OSDC after Pb2+ (or Cd2+)
sorption was quickly washed by 20 mL UHQ water and the water was discarded after
centrifugation. The remaining solid was oven dried at 60 °C for 48 h and was then identified
as post-sorption OSDC (OSDC-Pb or OSDC-Cd). The same procedure was employed to
control sample without the presence of Pb2+ (or Cd2+). The resulting sample was identified as
OSDC before sorption (OSDC-B). The net amount of released cation (K+, Ca2+, Mg2+, and
Ba2+) during Pb2+ and Cd2+ sorption was determined by the difference of concentration of
cations in solution before and after sorption. The OSDC-Pb, OSDC-Cd, and OSDC-B were
examined using the XRD, FTIR, and sequential extraction test. The analytical methods and
extraction procedures are given in Section 5.2.5.
6.3. Results and discussion
6.3.1. Characteristics of chars
In order to identify the most promising activated char for the subsequent sorption
mechanism studies, a preliminary sorption of heavy metal using chars prepared at different
pyrolysis temperatures and activation methods was carried out. Preliminary sorption results
are shown in Table 6.1. The activated chars showed similar Pb2+ removal performance, with
the sorption capacities in the range of 251.8–304.6 mg/g. However, among the five activated
chars, OS500N100-8M exhibited the best performance in removal of Cu2+, Cd2+, and Zn2+
from solution. The Cd2+ sorption capacity was increased by 4.7-fold for OS500N100-8M as
compared to that for OS500. Therefore, OS500N100-8M was chosen in this study to further
investigate its sorption mechanisms for heavy metals.

148

The physicochemical properties of chars are summarized in Table 6.2. Non-activated char
was not rich in C (3.73–7.33%) and its C concentration decreased as the pyrolysis temperature
increased from 300 °C to 700 °C. The decrease of C could be attributed to the transformation
of carbonaceous materials into CH4, CO2 and other gases or aromatic structures. During
pyrolysis, the release of volatile matters led to the loss of surface functional groups elements
(H and N). As a result, the content of H decreased from 0.50% at 300 °C to 0.09% at 900 °C,
and the content of N decreased from 0.10% at 300 °C to 0.02% at 900 °C (Table 6.2). The ash
content of chars increased from 88.8% to 96.3% with increasing pyrolysis temperature from
300 °C to 900 °C, mainly due to decomposition of volatile matters and enrichment of minerals.
The pH values of chars at 300–700 °C ranged from 10.7 to 11.3, and then decreased to 9.5 at
900 °C, which is in agreement with the previous result (Zhang et al., 2015). The lower pH
value of the char pyrolyzed at 900 °C may be related to its relatively lower concentration of
Ca, as compared to that of chars pyrolyzed at 300–700 °C. The Ca and Mg concentration
decreased when the pyrolysis temperature increased from 700 °C to 900 °C. On the one hand,
the concentrations of elements in chars were determined by acid digestion method (i.e., method
3050 B). It was reported that the elemental recovery using USEPA method 3050 B was affected
by a number of factors such as hot plate temperature, refluxing times, and acid additions (Chen
and Ma, 1998). Therefore, incomplete recovery might happen for method 3050 B. On the other
hand, the chloride (Cl) in the char sample could be incorporated with Ca and Mg to form MgCl2
or CaCl2, which volatilized at 700–900 °C (Zhang et al., 2015). The total basicity of chars
increased with increasing pyrolysis temperature because of increasing contents of ash. The
BET surface area increased gradually from 10.8 m2/g in char at 300 °C to 23.8 m2/g in char at
700 °C. This might be because volatilization was more subtle at higher temperature leading to

149

more porous char (Zhang et al., 2015). As can be observed from Table 6.2, the increase of
pyrolysis temperature from 700 °C to 900 °C caused a reduction of surface area and total pore
volume. Similar results has been reported in other studies (Chun et al., 2004; Yang et al., 2016).
It was reported that high temperature (above 650 °C) could destroy the pore walls between
adjacent pores, leading to decreased surface area (Zou et al., 2013). The activation process
improved the development of pores and created new pores. The BET surface area of
OS500N100-8M was 32.9 m2/g, which is approximately 1.9 times that of OS500. The total
pore volume (PV) of OS500N100-8M was 0.231 cm3/g, which is approximately 3.6 times that
of OS500.
Table 6.1. Preliminary sorption results of chars

Cu2+

Cd2+

Pb2+

Zn2+

Chars

Q48ha
mg/g

pHf b

Q48h
mg/g

pHf

Q48h
mg/g

pHf

Q48h
mg/g

pHf

OS300

134.5

5.06

44.90

5.98

260.9

5.48

68.35

6.54

OS500

140.3

4.94

17.56

6.43

297.0

5.67

75.22

6.51

OS700

173.2

6.07

83.00

5.02

298.4

6.54

75.75

6.18

OS900

114.3

4.94

15.73

6.14

81.82

5.17

9.153

6.26

OS500N900

241.5

5.40

62.03

6.11

300.0

ND

72.18

6.50

OS500N100-2M

184.6

5.21

54.68

5.98

273.9

5.40

55.02

6.45

OS500N100-8M

256.8

6.01

99.81

6.35

297.5

5.69

87.57

6.53

OS500N24-2M

154.2

4.93

16.85

6.11

304.6

6.86

81.31

6.54

FSOS500

19.29

4.64

18.21

6.43

251.8

5.19

14.18

6.43

Note: initial Cu2+ solution pH = 4.52; initial Cd2+ solution pH = 6.22; initial Pb2+ solution pH = 4.53;
initial Zn2+ solution pH = 5.98; ND: not detected.
a
The amount of heavy metal sorbed per unit mass of char at 48 h (mg/g).
b
Final pH values of solution.

150

Table 6.2. Physicochemical properties of charsa

Elemental content (%)b

a
b

Ca

Na

Mg

pH

pHPZC

TSA
mmol
/g

0.10 0.29

7.34

0.02

1.32

11.3

11.67

0.09

4.37

88.8

0.28

0.08 0.37

7.26

0.04

1.35

10.7

11.57

0.07

4.52

3.73

0.42

0.03 0.45

8.55

0.16

1.66

11.2

11.66

0.29

3.92

0.09

0.02 0.54

6.55

0.25

0.93

9.5

9.96

OS500N 4.54
100-8M

0.75

0.06 0.20

9.77

2.40

1.43

10.5

10.84

Chars

C

H

N

OS300

7.33

0.50

OS500

6.11

OS700
OS900

K

Values are given as the average of duplicate measurements.
On a dry basis.

151

TSB
mmol
/g

Ash
%b

CEC
cmol
/kg

Microstructures
SBET
m2/g

PV
cm3/g

DP
nm

117.1

10.8

0.069

3.62

90.4

86.95

17.4

0.064

3.63

6.10

94.3

255.9

23.8

0.203

3.87

0.26

6.23

96.3

104.7

6.50

0.034

3.89

0.14

8.60

89.6

163.3

32.4

0.231

3.85

The nitrogen adsorption-desorption isotherms and pore size distribution of chars are shown
in Figures 6.1 and 6.2. According to international union of pure and applied chemistry
(IUPAC), the N2 adsorption/desorption isotherms for all the chars corresponded to type II
adsorbents (Sing et al., 1985). As seen from Figure 6.1a and Figure 6.2a, the adsorbed volume
of N2 increased slowly with increasing P/P0 from 0.0 to 0.8. When the relative pressure P/P0
approached to 1, the capillary condensation could take place and the adsorbed volume
increased rapidly. The average pore diameter (Dp) of chars between 3.62 and 3.89 nm denoted
mesopores characteristics. It is interesting that new pores (in the range 3.5–4.5 nm) formed at
the expense of the small pores (1.0–3.0 nm) for OS700 (Figure 6.1b). As shown in Figure 6.2b,
the total pore volume for activated char was significantly larger than that of non-activated char,
which is mainly attributed to the appearance of new pores in the range 10.0–200.0 nm.
The mobility and speciation of heavy metals in OS300-B, OS500-B, OS700-B, OS900-B,
and OS500N100-8M were shown in Table D1 in Appendix D. Their water soluble fraction and
exchangeable fraction were negligible, suggesting there was no direct risks to plants and
humans when OSDCs were applied in water treatment. In fact, there was no Cd (or Pb), Cr,
Cu, Ni, and Zn was detected in solution after Pb2+ (or Cd2+) sorption (initial concentration:
Pb2+ 1600mg/L; Cd2+ 1000 mg/L) except a low concentration of Zn (0.6–0.7 mg/L) was
leached out from OS900 after the Pb2+ sorption. In this regard, the effect of Pb and Cd
originating from OSDCs could be ignored in the sorption studies.

152

Volume (cm3/g STP)

150

a

OS300
OS500
OS700
OS900

120
90
60
30
0
0.0

0.2

0.4

0.6

0.8

1.0

Relative pressure (P/P0)
0.018
0.015

dv/dD (cm3/nm/g)

OS300
OS500
OS700
OS900

b

0.012
0.009
0.006
0.003
0.000

1

10

100

1000

Pore diameter (nm)
Figure 6.1. (a) Nitrogen adsorption/desorption isotherms and (b) pore size distribution of the OSDCs
prepared at different pyrolysis temperatures.

153

a

Volume (cm3/g STP)

150

OS500
OS500N100-8M

120
90
60
30
0
0.0

0.2

0.4

0.6

0.8

1.0

Relative pressure (P/P0)
0.007

dv/dD (cm3/nm/g)

OS500
OS500N100-8M

b

0.006
0.005
0.004
0.003
0.002
0.001
0.000

1

10

100

1000

Pore diameter (nm)
Figure 6.2. (a) Nitrogen adsorption/desorption isotherms and (b) pore size distribution of the nonactivated OSDC and activated OSDC obtained at 500 °C.

154

6.3.2. XRD and FTIR results
The XRD spectra of the OSDCs before Pb2+/Cd2+ sorption is shown in Figure D1 (Appendix
D). The OS300-B presents mainly muscovite 2M2 (K0.77Al1.93(Al0.5Si3.5)O10(OH)2), barite
(BaSO4), quartz (SiO2), calcite (CaCO3), and dolomite (CaMg(CO3)2). Calcite and dolomite
started to decompose to calcium oxide (CaO) after heating above 500 °C. The peaks of
muscovite 2M2 and barite started to disappear above 700 °C, leading to the formation of
celsian (BaAl2Si2O8) and barium oxide hydrate (BaO2 (H2O)8). The hydrothermal treatment of
OS500 with NaOH solution was performed in order to synthesize zeolite (crystalline
aluminum-silicates) which could be an effective sorbent. However, the zeolite was not
identified in OS500N100-8M (Figure D1 in Appendix D). This might be because the
concentration of aluminum (Al) which is a primary source for forming zeolites was relatively
low in oily sludge (Table 5.1). As shown in Figure D1 in Appendix D, the complete conversion
of barite to witherite (BaCO3) was achieved under hydrothermal condition, probably resulting
from dissolution-recrystallization (Rendón-Angeles et al., 2008).
The higher pyrolysis temperature and activation led to the disappearance and the formation
of certain minerals, which could affect the heavy metal sorption capacities. To further illustrate,
the char samples loaded with Pb2+ or Cd2+ were also characterized using XRD (Figure 6.3).
New peaks at 27.080° and 34.155° showed hydrocerussite (Pb3(CO3)2(OH)2) formed on
OS300-Pb, OS500-Pb, OS700-Pb, and OS500N100-8M-Pb. Also, a precipitation of cerussite
(PbCO3) at 2θ = 24.792° and 47.016° were observed in OS300-Pb and OS500-Pb. These
precipitates formed in Pb2+-loaded OSDCs may be the product of the precipitation of Pb2+ and
the dissolved anions (i.e., OH‒ and CO32‒) from minerals. No precipitation of Pb-carbonates
was detected in OS900-Pb, probably because most of calcite and other carbonate underwent
155

decomposition at 900 °C. XRD results (Figure 6.3) showed that no precipitate formed on the
surface of OSDCs after Cd2+ sorption except for OS500N100-8M. Formation of otavite
(CdCO3, at 2θ = 30.275° and 23.485°) precipitates was probably related to the existence of
witherite (BaCO3) in OS500N100-8M.
Figure D2 in Appendix D presents the FTIR spectra of OSDC before Pb2+ or Cd2+ sorption.
The peaks at 3385–3425 cm‒1 corresponded to hydroxyl stretching (O–H) and disappeared at
900 °C. The band for aliphatic CH2 (2923 cm‒1) disappeared upon heating to 500 °C. At 300–
500 °C, the bands at 777 and 875 cm‒1 was assigned to aromatic C–H. The carboxyl C=O
(1633 cm‒1) became more apparent at 700 °C. The peak at 1007 cm‒1, which shifted towards
lower wavenumbers as a result of increasing pyrolysis temperature, was assigned to Si–O
stretching.
The organic functional groups on the surface of OSDCs before and after Pb2+ or Cd2+
sorption were also characterized using FTIR (Figure 6.4). The peaks at 1431 and 694 cm‒1
(assigned to CO32‒) of the OS300, were intensified obviously after Pb2+ sorption, suggesting
formation of lead carbonate which was consistent with that observed in XRD. Similar trends
were also observed for OS500, OS700, OS500N100-8M (Figure 6.4). The peaks at 1429 cm‒1
corresponding to CO32‒ in the OS500N100-8M was shifted to 1391 cm‒1 after it sorbed Cd2+,
indicating a change in the vibration of C–O bond which was probably due to formation of
cadmium carbonate precipitate. Compared with the FTIR spectra of the OS900-B, a new band
at 1370 and 1413 cm‒1 was observed in the OS900-Pb and OS900-Cd, respectively, indicating
a weak formation of Pb- or Cd-carbonate precipitation (Figure 6.4d). But no precipitates of Pb
or Cd-carbonate were detected in OS900-Pb and OS900-Cd by XRD (Figure 6.3d), which may
be due to the detection limit. The absence of peak associated with CO32‒ for OS900-B (Figure
156

6.4d) may be due to its low concentration. After Pb2+ sorption, the peak at 874 cm‒1 (assigned
to aromatic C–H) of OS500, and the peak at 860 cm‒1 (assigned to aromatic C–H) of
OS500N100-8M weakened obviously, confirming the effect of Pb2+-π interaction on the Pb2+
sorption process. After Cd2+ sorption, the peaks at 3385–3425 cm‒1 corresponding to O–H
vibrations of hydroxyl groups in OS300, OS500, OS700, and OS500N100-8M shifted, likely
from the coordination between Cd2+ and hydroxyl (–OH) functional groups. However, there
was no peak shifting from FTIR results suggesting the coordination between the –OH and Pb2+
in OSDCs, except for OS700 and OS500N100-8M.

157

b
Intensity (counts)

Intensity (counts)

a
OS300-Cd

OS300-Pb

H

Ce H

H
H

H H
OS300-B
M

Q

10

20

H
H Ce

40

H

Ce
OS500-Pb

H

H
Ce
H
H
Ce H Ce

Q

OS500-B
M

Q

Q

60

10

20

Q

50

H

30

40

Q

Q

Q

50

60

d
Intensity (counts)

Intensity (counts)

H

2θ (degree)

c
OS700-Cd

OS700-Pb

H

H

H

H

H

Q
OS700-B
M

Q

10

20

B

MB
OB B Q Q Q & B

30

40

Q

e
OS500N100-8M-Cd

OS900-B
Cl

60

10

Q

50

H

HH

20

F

30

Cl Q

Q

40

50

Q

Q

60

2M2, K0.77Al1.93(Al0.5Si3.5)O10(OH)2; B, Barite,

H H

Q
OS500N100-8M-B W
W
M
M
Q M

20

O & Cl & F

2θ (degree)

Ot

BaSO4; Q, Quartz, SiO2; C, Calcite, CaCO3; D,

Ot Ot

Ot

Dolomite, CaMg(CO3)2; O, Calcium oxide,

H

H

Q Cl Cl

Minerals with peaks labeled: M, Muscovite
Ot

OS500N100-8M-Pb

OS900-Pb

Q

Ot

Ot

OS900-Cd

Q

2θ (degree)

Intensity (counts)

Ce
H
CeCe

B BC
Q&B
M DBB Q Q

2θ (degree)

10

H

Q

Q
B B
Q&B
CD
M BB Q Q

30

H

OS500-Cd

H HH

H

H

H

CaO; Cl, Celsian, BaAl2Si2O8; F, Ferritungstite,

H

(W, Fe)(O,OH)3; W, Witherite, BaCO3; Ce,
Cerussite, PbCO3; H, hydrocerrusite,

M

30

W

Q W QW W Q

40

50

Q

Pb3(CO3)2(OH)2; Ot, Otavite, CdCO3.

Q

60

2θ (degree)

Figure 6.3. XRD patterns of (a) OS300, (b) OS500, (c) OS700, (d) OS900, and (e) OS500N100-8M
before and after Pb2+/Cd2+ sorption.
158

a

b

1431

OS500-Cd 3366

875 694
1007

OS300-Pb 3394

1404
682

OS300-B

4000 3500 3000 2500 2000 1500 1000

1397

OS500-B

500

1

Wavenumbers (cm )

Wavenumbers (cm )

c

837679

4000 3500 3000 2500 2000 1500 1000

500

694

1002

OS500-Pb 3385

1

d

1644 1417

OS700-Cd 3361

1417
874

Transmittance

Transmittance

2923
OS300-Cd 3255

1413

OS900-Cd

1633

OS700-Pb

1396

682

OS700-B 3421

4000 3500 3000 2500 2000 1500 1000

Transmittance

Transmittance

692
962

500

935
1370

OS900-Pb

OS900-B

4000 3500 3000 2500 2000 1500 1000

1

500

1

Wavenumbers (cm )

Wavenumbers (cm )

e

Transmittance

OS500N100-8M-Cd
3373

1632

OS500N100-8M-Pb
3535

OS500N100-8M-B
3425

1391

860
681
956

1402

1429

4000 3500 3000 2500 2000 1500 1000

500

1

Wavenumbers (cm )

Figure 6.4. FTIR spectra of (a) OS300, (b) OS500, (c) OS700, (d) OS900, and (e) OS500N100-8M
before and after Pb2+/Cd2+ sorption.

159

6.3.3. Sequential extraction results
The speciation of Pb2+ or Cd2+ in OSDCs after Pb2+ or Cd2+ sorption are shown in Table D2
(Appendix D) and Figure 6.5. The recovery percentages from sequential extraction tests are
shown in Table D3 (Appendix D). The recoveries of Cd2+ from OS500, OS700, OS900, and
OS500N100-8M were 80.92–93.52%, while that for OS300 was 65.22%. The recoveries of
Pb2+ from OSDCs were within 74.15–87.26%. Shen et al. (2016) conducted sequential
extraction to determine the speciation of Pb2+ on biochar treated soil. The recoveries of the
Pb2+ were within 61.50–97.28%, which were comparable to the recovery of Pb2+ and Cd2+ in
this study. Figure 6.5a-d demonstrated that physical sorption can be ignored in most of the
cases. Only 0–1.34 % of total Pb2+ or Cd2+ could be extracted from water in all OSDCs except
for OS900. The contribution of physical sorption in the OS900 was also very little. The
concentration of water-extractable Pb2+ accounted for 5.46 % of the total sorbed Pb2+ in OS900,
while the water-extractable Cd2+ accounted for 3.29 % of the total sorbed Cd2+. The carbonate
fraction dominated the existence of Pb2+ although to a lesser extent in OS900 (79.30%)
compared with OS300, OS500, OS700, and OS500N100-8M (92.99–98.06%). Conversely, the
exchangeable fraction was greater in OS900 (7.73%) than the other OSDCs (0.28–0.81%).
Cd2+ was mainly bound to the carbonate fraction (61.35–92.96%) and the exchangeable
fraction (5.59–29.78%). The proportion bound to carbonate was highest (92.96%) in
OS500N100-8M while the exchangeable fraction (5.59%) was lowest in OS500N100-8M.
Significant concentrations of K+, Ca2+, Mg2+, and Ba2+ were released during Pb2+/Cd2+
sorption (Table 6.3). These cations were mainly released from the minerals (e.g., CaCO3,
CaMg(CO3)2, and BaCO3) on the surface of OSDCs (Lu et al., 2012). Alternatively, it may
result from substitution between Pb2+ or Cd2+ and metal ions (e.g., Ca2+, Mg2+, and Ba2+)
160

associated with inner-sphere complexation. The formed precipitates and complexes containing
Pb2+ or Cd2+ could be desorbed at step 3 (fraction bound to carbonate) due to acidic
environment (Lu et al., 2012; Shen et al., 2017). Hence, the sorption (QR ) attributed to the
release of cations can be described as:
QR = Qe + Qp + Qcm

(6.1)

Where QR is the amount of Pb2+ or Cd2+ sorption attributed to the release of cations (mg/g),
Qe is the amount of Pb2+ or Cd2+ existed as exchangeable fraction (mg/g), Qp is the amount of
Pb2+ or Cd2+ attributed to the precipitation with minerals (mg/g), and Qcm is the amount of Pb2+
or Cd2+ substituted with metal ions to form inner-sphere complexation (mg/g).
According to the sequential extraction results, the net amount of released cations and eq.
(6.1), the contribution of different mechanisms to Pb2+ or Cd2+ sorption on OSDCs could be
estimated as shown in Table 6.4.

161

400

a

WS
CB

100

EX
OX+OM+RS

Q (mg Cd2+/g)

Q (mg Pb2+/g)

240
160

0

OS300

OS500

OS700

OS900

60
40

0

OS500
N100-8M

100

c
Fractions of Cd content (%)

Fractions of Pb content (%)

EX
OX+OM+RS

20

80

80
60
40
20
0

WS
CB

80

320

100

b

OS300

OS500

OS700

OS900

OS500

OS700

OS900

OS500
N100-8M

OS300

OS500

OS700

OS900

OS500
N100-8M

d

80
60
40
20
0

OS500
N100-8M

OS300

Figure 6.5. Sequentially extracted (a) Pb2+ and (b) Cd2+ amounts in the OSDCs after Pb2+ or Cd2+
sorption. Fractions of (c) Pb2+ and (d) Cd2+ in each step of sequential extraction (WS: water soluble
fraction, EX: exchangeable fraction, CB: fraction bound to carbonate, OX: fraction bound to Fe-Mn
oxides, OM: fraction bound to organic matter, RS: residual fraction).

162

Table 6.3. The release of K+, Ca2+, Mg2+, and Ba2+ during Pb2+ and Cd2+ sorption by OSDCs

The net amount of released cations (meq/100g)a

Pb2+

Ba2+

Sum of net amount Maximum
of released cations metal sorption
capacity
(mg/g)
(mg/g)

Rr
(%)b

K+

Ca2+

Mg2+

OS300

0.31 ± 0.00

177.15 ± 12.55

22.74 ± 0.38 ‒0.12 ± 0.08

207.28 ± 13.31

282.96 ± 3.22

73.3

OS500

0.27 ± 0.01

229.64 ± 0.79

48.75 ± 0.46 2.78 ± 0.11

291.58 ± 1.40

351.48 ± 2.08

83.0

OS700

‒0.09 ± 0.01

85.56 ± 2.93

51.27 ± 0.76 7.35 ± 0.10

149.28 ± 3.79

261.14 ± 2.02

57.1

OS900

‒0.31 ± 0.02

8.66 ± 0.59

2.30 ± 0.05

16.26 ± 0.99

49.68 ± 2.31

32.7

OS500N100- 0.11 ± 0.00
8M

69.01 ± 2.45

54.51 ± 0.64 115.37 ± 2.00

247.60 ± 0.97

310.00 ± 2.19

79.9

5.04 ± 0.31

Cd2+ OS300

‒0.01 ± 0.05

24.12 ± 9.22

0.61 ± 0.01

0.23 ± 0.06

14.02 ± 5.18

46.67 ± 0.69

30.0

OS500

0.02 ± 0.02

13.19 ± 0.54

4.32 ± 0.11

0.23 ± 0.02

9.98 ± 0.36

23.95 ± 3.79

41.7

OS700

‒0.15 ± 0.03

46.81 ± 3.12

21.18 ± 0.68 3.86 ± 0.03

40.30 ± 2.06

92.14 ± 3.11

43.7

OS900

‒0.34 ± 0.02

1.61 ± 0.24

1.41 ± 0.06

2.73 ± 0.37

3.04 ± 0.36

11.05 ± 1.50

27.5

OS500N100- 0.07 ± 0.07
8M

5.42 ± 0.23

7.33 ± 0.19

109.61 ± 1.93

68.82 ± 1.14

90.06 ± 3.17

76.4

The net amount of released cations were the difference values between the concentration of K +, Ca2+, Mg2+, and Ba2+ released into Pb2+ (Cd2+)
sorption system at 1600 mg/L Pb2+ (1000 mg/L Cd2+) solution with 0.01M NaNO3 at pH 5.0 and the corresponding release from OSDC in 0.01M
NaNO3 at pH 5.0.
b
Rr is calculated by dividing the sum of net amount of released cations by the total sorbed Pb2+ or Cd2+.
a

163

Table 6.4. Quantitative analysis of sorption mechanisms of Pb2+/Cd2+ onto OSDCs

Pb

Cd

Physical
sorption

Cation
exchange
with weak
binding

Precipitation

Complexation
with mineral
oxides of char

Complexation
with free
functional
groups

Cation-π
interaction

Others

Qwa
Rmk
mg/g %

Qeb
Rm
mg/g %

Qpc
mg/g

Rm
%

Qcmd
mg/g

Rm
%

Qcfe
mg/g

Rm
%

Qπf
mg/g

Rm
%

Qog
mg/g

Rm
%

OS300

0

0

1.71

0.81

205.57

97.97

/h

0

–i

0

–

0

2.90

1.38

OS500

0

0

1.38

0.45

290.20

93.79

/

0

–

0

13.22

4.27

4.62

1.49

OS700

0.29

0.13

0.44

0.20

148.84

68.13

/

0

54.32

24.86

–

0

14.59

6.68

OS900

2.10

5.47

2.97

7.74

13.29

34.63

/

0

17.15

44.68

–

0

2.88

7.50

OS500N
100-8Mj

0

0

0.76

0.28

246.84

91.25

/

0

<11.06 <4.09

<11.06 <4.09

11.86

4.38

OS300

0.41

1.35

6.29

20.66

–

0

7.73

25.39

15.83

52.00

–

0

0.18

0.59

OS500

0.02

0.10

2.93

15.12

–

0

7.05

36.38

9.26

47.78

–

0

0.12

0.62

OS700

0.02

0.02

8.79

10.66

–

0

31.51

38.21

41.22

49.99

–

0

0.92

1.12

OS900

0.33

3.28

3.00

29.82

–

0

0.04

0.40

6.13

60.93

–

0

0.57

5.67

OS500N
100-8M

0

0

4.71

5.59

64.11

76.12

/

0

14.18

16.84

–

0

1.22

1.45

Qw is the amount of Pb2+/Cd2+ existed as water soluble fraction.
b
Qe is the amount of Pb2+/Cd2+ existed as exchangeable fraction.
c
Qp is the sum of net amount of released cations in Pb2+/Cd2+ sorption ‒ the amount of Pb2+ existed exchangeable fraction (if the precipitation is
evidenced by XRD/FTIR)
d
Qcm is the sum of net amount of released cations in Cd2+ sorption ‒ the amount of Cd2+ existed exchangeable fraction (if the precipitation is not
evidenced by XRD/FTIR).
a

164

Qcf is the amount of Pb2+/Cd2+ existed as carbonate fraction ‒ Qcm or Qp (if the cation-π interaction is not evidenced by FTIR).
Qπ is the amount of Pb2+ existed as carbonate fraction ‒ Qp (if the cation-π interaction is verified but complexation with free functional groups is
not evidenced by FTIR).
g
Qo is the amount of Pb2+/Cd2+ existed as Fe-Mn oxides, organic matter, or residual fraction.
h
Not applicable.
i
Not evidenced by XRD and FTIR.
j
Qcf + Qπ = 11.06 mg/g, since both complexation with free functional groups and cation-π interaction were evidenced by FTIR.
k
Rm is calculated by dividing the Q resulted from each mechanism by the total extracted Pb2+ (Cd2+) in sequential extraction (Qt).
e
f

165

The sum of net amount of released cations (i.e., K+, Ca2+, Mg2+, and Ba2+) into Pb2+ sorption
system was accounted for 73.3%, 83.0%, 57.1%, and 32.7% of the total sorbed Pb2+ on OS300,
OS500, OS700, and OS900, respectively (Table 6.3). The released K+, Ca2+, Mg2+, and Ba2+
dramatically decreased from 291.58 to 16.26 mg Pb2+/g with the pyrolysis temperature
increasing from 500 to 900 °C. Accordingly, the Pb2+ sorption capacity decreased from 351.48
to 49.68 mg/g as the pyrolysis increased from 500 to 900 °C. As shown in the XRD patterns
(Figure D1 in Appendix D), CaCO3 and CaMg(CO3)2 decomposed at high temperature
(≥700 °C), thereby the amount of CO32‒ released from OS700 and OS900 was much smaller
than that from OS500, which could explain the sorption capacity difference among chars
prepared at different pyrolysis temperature. These results confirmed that the precipitation with
minerals dominated the Pb2+ sorption on the low-pyrolysis-temperature chars (≤500 °C). The
contribution of precipitation to Pb2+ sorption decreased with increasing pyrolysis temperature.
For example, the sorption capacity attributed to precipitation (Qp) of OS500 was 290.20 mg/g,
accounting for 93.97% of total extracted Pb2+ in sequential extraction (Qt), while Qp and Qp/Qt
of OS900 were only 13.39 mg/g and 13.29%, respectively (Table 6.4). The contribution of
complexation with free carboxyl/hydroxyl functional groups (Qcf) by OS700 and OS900 were
higher than OS300 and OS500. The Qcf in OS700 and OS900 reached 54.32 mg/g and 17.15
mg/g, respectively, while the corresponding proportions were 24.86% and 44.68%,
respectively. This may be attributed to the higher total surface acidity (TSA) value in OS700
and OS900 (Table 6.2). In the case of OS500N100-8M sample, the Qp accounted for 91.25%
of Qt value, indicating the precipitation is the dominant mechanism for Pb2+ sorption. The
released Ba2+ of OS500N100-8M was 115.37 mg/g, which is approximately 42 times greater

166

than that of OS500 (2.78 mg/g). This phenomenon may be due to the fact that BaCO3 was the
major carbonate mineral formed in the OS500N100-8M (Figure 6.3e).
For Cd2+ sorption, the Qcf/Qt value of non-activated OSDCs were 47.78–60.93%,
suggesting the complexation with free carboxyl/hydroxyl functional groups played a dominant
role in Cd2+ sorption. The sorption capacity of Cd2+ for OS700 (92.14 mg/g) was greater than
that of other three non-activated chars (11.05–46.67 mg/g), implying that 700 °C was the
optimized pyrolysis temperature for Cd2+ sorption. Due to the higher TSA value in OS700 than
in other non-activated chars, the Qcf of OS700 increased to 41.22 mg/g compared to that of
OS300 and OS500 (9.26–15.83 mg/g). Besides, the amount of Cd2+ substituted with metal ions
to form complexes (Qcm) in OS700 was obviously larger than that of other non-activated chars,
which may be related to its higher CEC value (Table 6.2). After chemical activation, the effect
of carboxyl/hydroxyl functional group on Cd2+ sorption (Qcf/Qt) decreased, while the
contribution of the precipitation (Qp/Qt) to sorption significantly increased. For OS500N1008M, the Qp value was 64.11 mg Cd2+/g, accounting for 76.12% of the Qt value, indicating the
precipitation between Cd2+ and the minerals could be the dominant mechanism for Cd2+
sorption.
Pearson correlation analysis (Table D4 in Appendix D) was conducted to explore the
relationship between the OSDCs properties and Pb2+/Cd2+ sorption mechanism. No
correlations were observed between the ash content and Qp,Pb, which was consistent with
previous studies (Li et al., 2018; Wang et al., 2015). The minerals that could release dissolved
anions (e.g., CO32‒, OH‒, and PO43‒) instead of Al2O3 and SiO2 in ash led to the formation of
precipitates (Li et al., 2018). Therefore, an increase in ash content could not increase the Pb2+
sorption performance in this study. The CEC of OSDCs showed positive correlations with

167

Qcm,Cd, Qcf,Cd, and Qt,Cd, implying the CEC values in OSDCs is a good predictor of Cd2+
sorption capacities.
6.4. Summary
The Pb2+ sorption capacity of OSDCs at different temperatures followed the order of OS500
(351.48 mg/g) > OS300 (282.96 mg/g) > OS700 (261.14 mg/g) > OS900 (49.68 mg/g). For
Cd2+, the order of sorption capacity was OS700 (92.14 mg/g) > OS300 (46.67 mg/g) > OS500
(23.95 mg/g) > OS900 (11.05 mg/g). Precipitation with minerals (e.g., CO32‒ and OH‒)
dominated the Pb2+ sorption on OSDCs except for OS900. Pb2+ was more effectively sorbed
onto low-pyrolysis-temperature OSDCs (≤500 °C), which was mainly due to their abundant
CO32‒ that can precipitate with Pb2+. With increasing pyrolysis temperature, the contribution
of precipitation to metal sorption reduced due to the decomposition of minerals which could
release dissolved anions (e.g., CO32‒, and OH‒). The sorption capacity of Cd2+ was less than
that of Pb2+ and no precipitation was formed during Cd2+ sorption by OSDCs. Complexation
with free carboxyl/hydroxyl functional groups dominated Cd2+ sorption on the OSDCs, while
the contribution of interaction with minerals (complexation with mineral oxides and metal ion
exchange) was second to functional groups. OS700 had a greater sorption capacity for Cd2+
than other OSDCs probably due to its higher effective cation exchange capacity (CEC) (255.9
cmol/kg) and total surface acidity (0.29 mmol/g). Among various activation methods,
hydrothermal treatment with NaOH solutions (8M) at 100 °C was determined to be the
optimum activation method considering the performance of activated OSDCs in removal Pb2+,
Cu2+, Cd2+, and Zn2+ from solution. The resultant char (OS500N100-8M) had significantly
increased CEC, surface area, and total pore volume compared to the unmodified char (OS500).
Moreover, the complete conversion of barite (BaSO4) to witherite (BaCO3) after hydrothermal
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treatment contributed to the mineral precipitation of Cd2+. As a result, the Cd2+ sorption
capacity of OS500N100-8M (90.06 mg/g) was over 3 times higher than that of OS500 (23.95
mg/g). Overall, the oily sludge can be converted into value-added char as an effective sorbent
for Pb2+ and Cd2+. Pyrolysis temperature and activation method affected the chemical
properties (e.g., H/C, CEC, and ash content), microstructure (e.g., surface area and total pore
volume), as well as the mineral components (e.g., CO32‒, OH‒) in the OSDC, consequently
changed the Pb2+ and Cd2+ sorption capacity and the contribution of different mechanism.
However, further research is necessary to evaluate and minimize the environmental risk of
preparation and application of OSDCs.

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Chapter 7
CONCLUSION AND RECOMMENDATIONS
7.1. Synthesis and conclusion
The effective treatment of oily sludge is a worldwide problem faced by many refineries,
pipelines, and oil producing companies. Most common ways to treat oily sludge such as
landfills and incinerations are challenged by their unsatisfactory efficiency, high operational
cost, and/or possible production of secondary toxic compounds. In fact, oily sludge contains a
large amount of combustibles with high heating values. The proper solution to solve the
problem of oily sludge requires consideration of the economic benefit and in compliance with
environmental standards. This study proposed three effective oily sludge management
techniques: (i) IL-enhanced solvent extraction; (ii) co-pyrolysis with wood waste; and (iii)
converting oily sludge into sorbent to remove Pb2+ and Cd2+ from solution.
Solvent extraction is a commonly used approach to recycle petroleum hydrocarbons from
oily sludge. However, the organic solvents introduce problems associated with handling
volatile, flammable, and toxic material, limiting the extension of this technology to field-scale
oily sludge treatment. Chapter 3 proposed an ionic liquid (IL) enhanced solvent extraction
method for oily recovery from oily sludge. Compared with conventional solvent extraction,
the advantages of IL-enhanced solvent extraction are as follows: (i) the issue associated with
solvent loss by volatilization was addressed since the solvent volume was greatly reduced; (ii)
the residual IL was completely separated from residual sand/clay after water washing, which
is of significance for recycling IL; (iii) there was no evidence of IL and mineral fines existing
in the recovered oil after IL-enhanced solvent extraction; and (iv) the process can be
economically competitive since IL yielded high oily recovery at low energy consumption and
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short extraction duration. This work expanded the existing limited knowledge regarding the
use of IL in oily sludge treatment. Chromatographic characterization of recovered oil indicated
that the IL addition did not promote alteration of the long-chain and heavy petroleum
hydrocarbons (PHCs). The break of incrustation and desorption of oil from solid matrices were
high at short extraction duration (10 min), low solvent/sludge ratio (4:5 mL/g), and low energy
consumption (100 rpm). However, due to the constraint of the small solvent/sludge ratio (4:5
mL/g), not all desorbed oil could be dissolved in a solvent, leaving 29.33% of TPH in the oilIL interface. The higher solvent/sludge ratio and stronger shearing force could enhance the
dissolution of suspended oil into the solvent.
As one of the thermal methods to treat oily sludge, pyrolysis has the advantages of energy
recovery (pyrolysis oil and gas), solid waste reduction, and less secondary pollution. More
importantly, pyrolysis can immobilize heavy metals in the solid residue. Chapter 4 concluded
that co-pyrolysis of oily waste with other biomass waste is an effective approach for both oil
recovery and heavy metal immobilization in the obtained solid residue. The co-pyrolysis
treatment could achieve more than 50% of total petroleum hydrocarbon (TPH) recovery. The
degree of potential ecological risk posed by heavy metals in the solid residue from co-pyrolysis
was also reduced from “considerable risk” (RI of 302.50 for untreated oily waste) to “low risk”
(RI of 21.96–120.2 for solid residue). This work adds to the current understanding of factors
that contribute the oil recovery and heavy metal immobilization during co-pyrolysis treatment.
The individual factors of pyrolysis temperature, reaction time, and hog fuel addition on TPH
recovery and ecological risk of solid char residue, were studied using a response surface
methodology (RSM). In addition, the interaction effects between any two of the
aforementioned factors were also studied. Oil recovery and heavy metal immobilization

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responded very differently to experimental factors. Specifically, a maximum TPH recovery
was obtained at a low temperature (400 °C), but the increase in temperature could reduce the
potential ecological risk of heavy metals. The addition of hog fuel into oily waste for copyrolysis led to an increased metal immobilization, but the overall oil recovery was decreased.
A longer reaction time could favor the pyrolysis conversion, but had little effect on heavy metal
immobilization. In this study, a low temperature co-pyrolysis (400 °C) by adding 20% of hog
fuel was found as the most effective oily waste disposal strategy with satisfactory oil recovery
and an acceptable risk index of 54.1. However, it is important to consider the site specific
factors for determination of optimum condition in co-pyrolysis treatment. Environmental
testing, including the determination of heavy metal concentration and oil, water, and solid
composition, in oily sludge is essential for co-pyrolysis treatment decision making.
Converting oily sludge into sorbent could be regarded as a “win-win” strategy since this
conversion not only provides safe disposal of oily sludge, but also allows its reuse in water
treatment applications. Chapter 5 investigated the Pb2+ and Cd2+ sorption capacity and
mechanisms by char (OS500) prepared from pyrolysis of oily sludge at 500 °C. The sorption
of Pb2+ and Cd2+ on OS500 was a chemical rather than physical process. Sorption of Pb2+ by
OS500 was mainly attributed to its precipitation with CO32‒ originating in OS500, while
complexation dominated Cd2+ sorption on OS500. Chapter 6 investigated the influence of
pyrolysis temperature and chemical activation on properties and heavy metal sorption
performance of oily sludge-derived chars (OSDCs). With the temperature increasing from
300 °C to 700 °C, the ash content and microstructure development of OSDCs promoted. The
best Cd2+ sorption performance of OSDC prepared at pyrolysis temperature of 700 °C (OS700)
was probably due to its high effective cation exchange capacity (CEC) and total surface acidity.

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However, the Pb2+ sorption capacity dramatically decreased as the pyrolysis increased from
500 to 900 °C. This is because the minerals such as CaCO3 and CaMg(CO3)2 decomposed at
high temperature (≥700 °C), thereby the mineral precipitation with Pb2+ was inhibited. The
hydrothermal treatment (8 M NaOH solution) activated OS500 significantly with respect to
surface area, CEC, and total pore volume (PV). The activated OSDC showed a higher sorption
capacity for Cd2+ than OS500 because conversion of barite (BaSO4) to witherite (BaCO3) after
hydrothermal treatment favored the precipitation of Cd-carbonate. This work contributed to
understand the mechanisms of Pb2+ and Cd2+ sorption by OSDCs, which is of significant
importance to the practical application of OSDCs in water treatment. Pyrolysis temperature
and chemical activation were crucial for both heavy metals sorption capacities of OSDCs and
the contribution of different mechanisms. Chemical activation significantly contributed to the
precipitation of Cd2+. With increasing pyrolysis temperature, the contribution of mineral
precipitation with Pb2+ reduced, while the contribution of functional groups complexation
increased.
In summary, the proposed methods could represent environmentally friendly and
economically competitive alternatives for the effective treatment of oily sludge. The selection
of alternative management methods should be based on a solid understanding of the
characteristics of oily sludge. IL-enhanced solvent extraction is a very promising solution to
treat oily sludge containing relatively high oil content. This method could not only maximize
the benefit of heating values originating from oily sludge, but also reduce the waste volume.
Meanwhile, the greater amount of PHCs was recovered as oil, the less amount of unrecovered
PHCs was left in residual sand/clay after extraction, indicating the further remediation of
residual sand/clay would be easier. Co-pyrolysis with wood waste is recommended to treat oily

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sludge with a high concentration of heavy metals and a certain amount of TPH. After copyrolysis treatment, TPH was effectively recovered from oily sludge. Pyrolysis temperature
and wood waste addition had a positive effect on heavy metal immobilization in solid char.
Converting oily sludge into sorbent could be a desirable method to treat oily sludge with
relatively low TPH content. It was found that the minerals which could release dissolved anions
(i.e., CO32‒ and OH‒) played a very important role in Pb2+ and Cd2+ sorption on OSDCs. In
order to evaluate the suitability of converting oily sludge into sorbent, it is essential to identify
the crystalline phases in OSDC using XRD.
7.2. Limitations and future research
Although the methods proposed in this study demonstrated promising results, they are all
at a lab scale. Before large-scale application, the linkage between treatment performance and
laboratory test results needs careful verification.
In this study, cyclohexane was selected as the solvent in IL-enhanced solvent extraction. In
fact, any organic solvent that dissolves bitumen but immiscible with the IL will accomplish
the separation of oil from oily sludge. The extraction performance is very sensitive to the
solvent properties, such as solubility parameter, the distribution of the polar force, dispersion
force, and hydrogen-bonding force diameters, and the structure of the solvent. It is also possible
that the mixed solvent could achieve higher oil recovery then the pure solvents. Unlike
laboratory solvents, the industrial solvents are generally of lower purity and consist of a
mixture of related compounds. Therefore, the effect of other solvents or solvent mixtures on
the extraction performance in IL-enhance solvent extraction needs to be further investigated.
Since ILs are expensive materials, the future challenge of IL for oil recovery from oily
sludge is to customize low-cost but effective ILs. It is recommended to continue more research
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works to understand the extraction mechanism of the oil in the presence of IL. In this respect,
zeta potential distribution measurements and atomic force microscopy (AFM) could be used
to study the surface interactions between the components of oily sludge (e.g., oil and silica).
In addition, the near infrared spectroscopy (NIR) could be used to study the effects of ILs on
extraction kinetics. To minimize the concerns about the high cost of IL, further research about
the reuse of IL need to be conducted.
In co-pyrolysis treatment, there were more than one reactions occurred. Oil recovery took
place due to distillation rather than pyrolysis considering the similarities between the
composition of the recovered oil and of the oil in the original oily waste. However, the charred
biomass formed from pyrolysis did favor the immobilization of heavy metal ions. More work
should be done to further study pathway of the formation of co-pyrolysis products and the
related mechanism using real oily sludge sample. In doing so, the compositions of noncondensable gas and condensable oil from co-pyrolysis of oily sludge should be analyzed by
gas chromatography-mass spectrometry (GC-MS) and gas chromatography-thermal
conductivity detector (GC-TCD). The kinetic and thermal conversion behavior of oily sludge
and wood waste during co-pyrolysis should be investigated through thermogravimetric
analysis (TGA). To better understand the mechanism involved in the immobilization of the
metals during co-pyrolysis, the characterization of solid char using X-ray absorption near-edge
spectra (XANES), TGA, and XRD is recommended.
The addition of catalysis could not only increase the conversion of oily sludge to desirable
products, but also improve the quality of pyrolysis oil. Further study should be focused on the
addition of catalysis in the co-pyrolysis process of oily sludge. From an economic point of

175

view, cheap and easily available catalysis or additives such as fly ash and oily sludge ash,
should be considered when the raw materials used are wastes.
Once Pb2+ and Cd2+ were sorbed on OSDC, they are difficult to be released since the
majority of them would not leach out into water or MgCl2 solution. Future studies on the metal
recovery and reuse of OSDC is needed in future. Conversely, the metal-loaded OSDC could
be solidified to an environmentally safe form. A thermal treatment could be conducted for
further stabilization of the sorbed metal. The mineral constituents (e.g., silicon, aluminum, iron,
and calcium) of OSDC may potentially act as precursors to incorporate the sorbed metals into
the crystal structure through a thermal treatment. The role of mineral constituents in
stabilization of sorbed metals would be another venue for further study.
The original contamination of oily sludge might migrate, transform, and volatilize during
the preparation process of OSDC, which might become an environmental issue. Further
research will include the off-gas treatment in the preparation process and migration and
transformation of element in sludge by pyrolysis. To assess the environmental risk of
application of OSDCs to water treatment, long-term heavy metal leaching experiment are
needed.
The majority of Pb2+ was sorbed on OS500 through precipitation, which represented a
bound to carbonate fraction. When OS500 is applied to Pb2+ contaminated soil, it is expected
the OS500 could reduce the bioavailability of Pb2+, and consequently reduced the
environmental risks. The stability of the bound to carbonate fraction could be affected by the
solubility of formed precipitates and the site conditions such as soil pH. Further sorption
studies are necessary to investigate the long-term effectiveness of OS500 as a heavy metal
sorbent in soil. Heavy metals such as Pb, Cu, Zn, and Cd may coexist in contaminated waters,

176

thereby there is a competition for sorption sites on OSDC surface between metals. Future study
should explore OSDC potential to treat multi-metal contaminated water. To synthesize zeolites
from OSDC, the modification of aluminum/silicon (Al/Si) ratio by adding aluminum oxide
(Al2O3) during hydrothermal treatment is recommended.
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APPENDIX A
Table A1. Mass balance calculation of the F2, F3, and F4 fractions of TPH for orthogonal experiments
of solvent extraction alone

Exp. No.

F2o
F2t
F2b
F3o
F3t
(µg)
(µg)
(µg)
(µg)
(µg)
1
Rep1
105,869
97,816
4,657
148,833
120,486
Rep2
106,609
93,683
5,066
149,873
113,599
Rep3
106,102
95,226
3,621
149,160
115,632
2
Rep1
106,905
104,586
3,911
150,289
120,657
Rep2
106,863
10,0243
3,399
150,230
114,618
Rep3
106,102
100,362
1,985
149,160
120,210
3
Rep1
106,440
91,596
4,182
149,635
125,413
Rep2
105,172
91,204
4,418
147,852
116,729
Rep3
105,743
91,517
4,957
148,655
118,453
4
Rep1
105,700
101,445
902
148,595
129,972
Rep2
105,996
99,683
745
149,011
120,081
Rep3
107,138
106,462
1,641
150,616
124,304
5
Rep1
106,525
102,535
2,081
149,754
128,291
Rep2
105,827
99,858
1,436
148,773
121,563
Rep3
107,053
96,722
2,660
150,497
130,621
6
Rep1
106,673
81,048
3,396
149,962
124,989
Rep2
106,948
96,484
3,197
150,349
123,480
Rep3
105,256
102,705
2,078
147,971
126,867
7
Rep1
105,362
79,100
2,002
148,120
137,957
Rep2
104,728
111,885
1,163
147,228
131,963
Rep3
106,271
107,745
1,515
149,398
136,297
8
Rep1
104,051
83,927
1,435
146,277
125,923
Rep2
106,102
107,317
2,889
149,160
127,029
Rep3
106,123
99,805
1,825
149,190
132,324
9
Rep1
106,652
95,827
2,862
149,932
131,858
Rep2
106,673
103,002
2,296
149,962
136,641
Rep3
106,800
120,085
1,411
150,141
139,262
F(2/3/4)o: mass of F2/F3/F4 fractions of TPH in the original oily sludge sample;
F(2/3/4)t: mass of F2/F3/F4 fractions of TPH in the recovered oil;
F(2/3/4)b: mass of F2/F3/F4 fractions of TPH in the residual sand/clay;
F(2/3/4)s: sum of F(2/3/4)t and F(2/3/4)b;
TPHs: sum of F2t, F2b, F3t, F3b, F4t, and F4b;
TPHo: sum of F2o, F3o, and F4o.

208

F3b
(µg)
16,551
18,709
18,784
16,237
17,441
17,865
15,440
15,004
17,743
7,357
10,748
14,271
11,655
10,583
13,291
13,008
11,663
14,095
8,616
9,674
11,864
11,001
12,996
9,029
9,157
10,080
8,620

Table A1. Mass balance calculation of the F2, F3, and F4 fractions of TPH for orthogonal experiments
of solvent extraction alone (continued)

Exp. No.

F4o
(µg)
39,267
39,542
39,354
39,652
39,636
39,354
39,479
39,009
39,220
39,205
39,314
39,738
39,510
39,252
39,706
39,565
39,667
39,040
39,079
38,844
39,416
38,593
39,354
39,361
39,557
39,565
39,612

F4t
(µg)
26,898
26,118
24,042
29,657
29,439
29,627
31,291
33,667
30,430
36,083
33,437
33,333
37,674
38,278
35,857
40,064
36,636
34,328
44,443
36,452
38,858
39,451
37,254
37,371
42,315
42,057
37,753

F4b
F2s/
F3s/
F4s/
(µg)
F2o
F3o
F4o
1
Rep1
11,192
0.97
0.92
0.97
Rep2
12,025
0.93
0.88
0.96
Rep3
11,095
0.93
0.90
0.89
2
Rep1
11,152
1.01
0.91
1.03
Rep2
10,760
0.97
0.88
1.01
Rep3
9,550
0.96
0.93
1.00
3
Rep1
9,978
0.90
0.94
1.05
Rep2
10,579
0.91
0.89
1.13
Rep3
8,800
0.91
0.92
1.00
4
Rep1
4,241
0.97
0.92
1.03
Rep2
8,908
0.95
0.88
1.08
Rep3
6,588
1.01
0.92
1.00
5
Rep1
6,964
0.98
0.93
1.13
Rep2
6,480
0.96
0.89
1.14
Rep3
5,849
0.93
0.96
1.05
6
Rep1
5,496
0.79
0.92
1.15
Rep2
6,740
0.93
0.90
1.09
Rep3
6,127
1.00
0.95
1.04
7
Rep1
5,182
0.77
0.99
1.27
Rep2
5,863
1.08
0.96
1.09
Rep3
6,522
1.03
0.99
1.15
8
Rep1
5,564
0.82
0.94
1.17
Rep2
6,215
1.04
0.94
1.10
Rep3
5,345
0.96
0.95
1.09
9
Rep1
4,181
0.93
0.94
1.18
Rep2
4,924
0.99
0.98
1.19
Rep3
5,669
1.14
0.98
1.10
Average 0.95
0.93
1.08
F(2/3/4)o: mass of F2/F3/F4 fractions of TPH in the original oily sludge sample;
F(2/3/4)t: mass of F2/F3/F4 fractions of TPH in the recovered oil;
F(2/3/4)b: mass of F2/F3/F4 fractions of TPH in the residual sand/clay;
F(2/3/4)s: sum of F(2/3/4)t and F(2/3/4)b;
TPHs: sum of F2t, F2b, F3t, F3b, F4t, and F4b;
TPHo: sum of F2o, F3o, and F4o.

209

TPHS/
TPHo
0.94
0.91
0.91
0.96
0.93
0.95
0.94
0.93
0.93
0.95
0.93
0.96
0.98
0.95
0.96
0.90
0.94
0.98
0.95
1.02
1.03
0.93
1.00
0.97
0.97
1.01
1.05
0.96

Table A2. Mass balance calculation of the F2, F3, and F4 fractions of TPH for orthogonal experiments
of IL-enhanced solvent extraction process

Exp.
F2o
F2t
F2i
F2b
F3o
F3t
F3i
F3b
No.
(µg)
(µg)
(µg)
(µg)
(µg)
(µg)
(µg)
(µg)
1 Rep1 105,341 78,621 18,324
710
148,090 98,182 37,636 10,853
Rep2 107,497 80,939 26,914
661
151,121 102,186 43,929 11,166
Rep3 105,806 93,527 22,459
657
148,744 101,161 42,726 11,766
2 Rep1 105,320 77,363 17,634
737
148,060 105,123 34,762
9,659
Rep2 104,791 66,603 13,988
978
147,317 105,949 43,430 10,420
Rep3 106,482 96,918 15,843
679
149,695 111,224 35,393 11,908
3 Rep1 105,447 86,525 11,372
410
148,239 115,224 26,216
7,502
Rep2 105,193 87,197 16,765
612
147,882 112,516 36,823
9,785
Rep3 105,595 85,474 10,795
201
148,447 113,586 30,676 12,047
4 Rep1 107,582 91,128 13,250
635
151,240 110,960 39,374
9,193
Rep2 105,806 92,027 11,853
238
148,744 106,100 42,328
9,279
Rep3 104,960 82,947
6,520
208
147,555 105,687 41,431 12,473
5 Rep1 107,053 90,705 13,556
495
150,497 115,875 31,023
8,038
Rep2 105,700 83,990 15,302
740
148,595 113,091 32,281
9,242
Rep3 104,516 90,602 12,000
197
146,931 104,409 31,658 13,913
6 Rep1 107,095 90,032
8,147
812
150,557 113,313 31,430 11,220
Rep2 107,201 76,835 17,204
188
150,705 100,882 38,155
8,816
Rep3 104,580 76,035 10,706
526
147,020 105,760 25,943 17,432
7 Rep1 105,320 95,031 10,767
215
148,060 108,837 29,403
7,933
Rep2 106,398 85,364
8,702
355
149,576 102,902 25,985
9,173
Rep3 104,749 75,054 14,839
186
147,258 94,846 35,125 12,831
8 Rep1 106,377 105,510 3,605
399
149,546 122,904 21,351
7,173
Rep2 106,398 87,722
6,239
317
149,576 118,831 22,586
4,369
Rep3 106,863 88,730 10,127
787
150,230 104,311 29,355 14,738
9 Rep1 104,876 81,642
4,378
382
147,436 123,428 26,409
6,607
Rep2 106,208 97,020 17,188
185
149,308 118,540 29,321
3,822
Rep3 106,017 87,683
8,841
819
149,041 115,452 19,670 14,425
F(2/3/4)o: mass of F2/F3/F4 fractions of TPH in the original oily sludge sample;
F(2/3/4)t: mass of F2/F3/F4 fractions of TPH in the recovered oil from top oil layer;
F(2/3/4)i: mass of F2/F3/F4 fractions of TPH in the recovered oil from oil-IL interface;
F(2/3/4)b: mass of F2/F3/F4 fractions of TPH in the residual sand/clay;
F(2/3/4)s: sum of F(2/3/4)t, F(2/3/4)i and F(2/3/4)b;
TPHs: sum of F2t, F2i, F2b, F3t, F3i, F3b, F4t, F4i, and F4b;
TPHo: sum of F2o, F3o, and F4o.

210

Table A2. Mass balance calculation of the F2, F3, and F4 fractions of TPH for orthogonal experiments
of IL-enhanced solvent extraction process (continued)

Exp. No.

F4o
(µg)
39,071
39,871
39,244
39,063
38,867
39,495
39,111
39,016
39,165
39,902
39,244
38,930
39,706
39,205
38,766
39,722
39,761
38,789
39,063
39,463
38,852
39,456
39,463
39,636
38,899
39,393
39,322

F4t
(µg)
12,897
17,075
13,412
16,113
21,248
17,858
19,651
24,887
19,039
17,112
22,173
16,766
19,720
25,819
18,489
22,355
23,782
20,305
21,032
22,733
15,800
24,786
29,347
19,059
26,931
31,440
22,265

F4i
(µg)
21,739
24,857
21,116
20,704
22,882
18,664
19,912
23,012
17,529
25,876
22,419
21,448
22,821
19,717
18,242
21,923
24,441
15,751
19,230
15,913
18,737
17,544
13,775
16,018
17,614
15,792
12,689

F4b
F2s/
F3s/
F4s/
TPHS/
(µg)
F2o
F3o
F4o
TPHo
1 Rep1
10,653
0.93
0.99
1.16
0.99
Rep2
5,061
1.01
1.04
1.18
1.05
Rep3
7,276
1.10
1.05
1.07
1.07
2 Rep1
9,674
0.91
1.01
1.19
1.00
Rep2
4,540
0.78
1.08
1.25
1.00
Rep3
7,028
1.07
1.06
1.10
1.07
3 Rep1
8,721
0.93
1.00
1.23
1.01
Rep2
3,628
0.99
1.08
1.32
1.08
Rep3
5,440
0.91
1.05
1.07
1.01
4 Rep1
6,635
0.98
1.05
1.24
1.05
Rep2
4,082
0.98
1.06
1.24
1.06
Rep3
5,801
0.85
1.08
1.13
1.01
5 Rep1
6,914
0.98
1.03
1.25
1.04
Rep2
5,255
0.95
1.04
1.30
1.04
Rep3
6,374
0.98
1.02
1.11
1.02
6 Rep1
8,293
0.92
1.04
1.32
1.03
Rep2
5,293
0.88
0.98
1.35
0.99
Rep3
5,816
0.83
1.01
1.08
0.96
7 Rep1
7,254
1.01
0.99
1.22
1.02
Rep2
3,960
0.89
0.92
1.08
0.93
Rep3
6,181
0.86
0.97
1.05
0.94
8 Rep1
7,222
1.03
1.01
1.26
1.05
Rep2
5,416
0.89
0.97
1.23
0.98
Rep3
7,709
0.93
0.99
1.08
0.98
9 Rep1
8,188
0.82
1.06
1.36
1.01
Rep2
4,951
1.08
1.02
1.32
1.08
Rep3
8,026
0.92
1.00
1.09
0.98
Average 0.94
1.02
1.20
1.03
F(2/3/4)o: mass of F2/F3/F4 fractions of TPH in the original oily sludge sample;
F(2/3/4)t: mass of F2/F3/F4 fractions of TPH in the recovered oil from top oil layer;
F(2/3/4)i: mass of F2/F3/F4 fractions of TPH in the recovered oil from oil-IL interface;
F(2/3/4)b: mass of F2/F3/F4 fractions of TPH in the residual sand/clay;
F(2/3/4)s: sum of F(2/3/4)t, F(2/3/4)i and F(2/3/4)b;
TPHs: sum of F2t, F2i, F2b, F3t, F3i, F3b, F4t, F4i, and F4b;
TPHo: sum of F2o, F3o, and F4o.

211

Table A3. ANOVA for TPH, F2, F3, and F4 recovery in solvent extraction alone

a

Sources of variation
For TPH recovery
(A) So/Sl ratio
(B) Shaking speed
(C) Extraction duration
Error
Total
For F2 recovery
(A) So/Sl ratio
(B) Shaking speed
(C) Extraction duration
Error
Total
For F3 recovery
(A) So/Sl ratio
(B) Shaking speed
(C) Extraction duration
Error
Total
For F4 recovery
(A) So/Sl ratio
(B) Shaking speed
(C) Extraction duration
Error
Total

SSa

DFb Mean square F-value Prob > F Significance

494.59
1.72
70.03
147.32
730.67

2
2
2
18
26

247.30
0.86
35.01
8.18

30.21
0.10
4.28

<0.001
0.90
0.03

100.54
32.46
258.73
1030.96
1784.37

2
2
2
18
26

50.27
16.23
129.37
57.28

0.88
0.28
2.26

0.43
0.76
0.13

470.02
11.20
35.42
65.79
632.08

2
2
2
18
26

235.01
5.60
17.71
3.66

64.30
1.53
4.85

<0.001
0.24
0.02

**

3415.07
276.35
227.23
336.54
4522.00

2
2
2
18
26

1707.54
138.17
113.62
18.70

91.33
7.39
6.08

<0.001
0.0045
0.0096

**
**
**

**
*

*

Sum of squares; b Degree of freedom; * Significant (P < 0.05); ** Highly significant (P < 0.01)
Table A4. ANOVA for TPH recovery in IL-enhanced solvent extraction process

Sources of variation SSa
DFb Mean square
For TPH recovery
(A) So/Sl ratio
1.79
2
0.89
(B) IL/Sl ratio
89.16 2
44.58
(C) Shaking speed
28.00 2
14.00
Error
345.68 18
19.20
Total
554.09 26
Part 1: TPH recovery in top oil layer
(A) So/Sl ratio
139.83 2
69.92
(B) IL/Sl ratio
125.79 2
62.90
(C) Shaking speed
135.12 2
67.56
Error
313.87 18
17.44
212

F-value Prob > F Significance
0.05
2.32
0.73

0.95
0.13
0.50

4.01
3.61
3.87

0.036
0.048
0.040

*
*
*

a

Sources of variation SSa
DFb Mean square
Total
761.49 26
Part 2: TPH recovery in oil-IL interface
(A) So/Sl ratio
111.44 2
55.72
(B) IL/Sl ratio
17.69 2
8.84
(C) Shaking speed
284.74 2
142.37
Error
125.48 18
6.97
Total
575.12 26

F-value Prob > F Significance
7.99
1.27
20.42

0.0033
0.31
<0.001

**
**

Sum of squares; b Degree of freedom; * Significant (P < 0.05); ** Highly significant (P < 0.01)

213

APPENDIX B
Table B1. Concentrations of extractable fractions of heavy metals in the oily waste (SRC) and its solid residues

Sample
SRC
Exp. No.1
Extractable fractions (mg/kg)
Cu F1 325.47 ± 8.21
13.79 ± 2.50
F2 157.94 ± 8.92
101.16 ± 1.27
F3 101.28 ± 3.12
155.61 ± 16.56
F4 96.79 ± 6.04
525.52 ± 0.37
F5 37 ± 2
116 ± 16
Cr
F1 5.06 ± 0.07
0.20 ± 0.28
F2 24.83 ± 0.79
11.17 ± 0.11
F3 85.12 ± 2.59
47.27 ± 3.96
F4 648.10 ± 6.53
807.79 ± 50.02
F5 53 ± 4
140 ± 54
Ni
F1 677.44 ± 6.59
10.45 ± 1.92
F2 64.00 ± 7.01
29.94 ± 1.24
F3 22.21 ± 3.66
73.14 ± 8.12
F4 25.62 ± 2.25
389.58 ± 3.02
F5 44 ± 0
505 ± 10
Pb
F1 128.72 ± 2.69
0.00
F2 159.30 ± 8.17
72.56 ± 3.58
F3 278.78 ± 1.02
187.33 ± 18.52
F4 177.97 ± 8.03
573.10 ± 26.33
F5 56 ± 1
127 ± 11

Exp. No.2

Exp. No.3

Exp. No.4

Exp. No.5

0.71 ± 0.11
97.81 ± 8.41
104.34 ± 1.30
476.30 ± 31.14
318 ± 24
0.00
1.18 ± 0.06
5.41 ± 0.01
248.03 ± 22.25
846 ± 22
3.47 ± 0.68
18.15 ± 0.78
36.35 ± 0.06
447.62 ± 38.74
597 ± 40
0.00
66.95 ± 7.72
114.09 ± 1.40
344.36 ± 31.22
524 ± 40

20.41 ± 4.45
154.07 ± 3.85
176.50 ± 9.11
397.51 ± 3.75
220 ± 5
0.00
10.14 ± 0.65
40.17 ± 0.97
747.93 ± 156.16
271 ± 18
6.42 ± 1.41
26.20 ± 0.66
56.33 ± 0.87
401.22 ± 11.16
581 ± 8
0.00
87.15 ± 0.88
164.61 ± 0.74
587.21 ± 54.94
181 ± 13

11.94 ± 0.59
134.58 ± 8.92
168.50 ± 2.62
453.60 ± 52.01
215 ± 41
0.00
1.77 ± 0.23
5.73 ± 0.52
693.62 ± 109.65
385 ± 52
2.82 ± 0.42
11.95 ± 0.35
23.54 ± 1.20
362.06 ± 25.99
688 ± 24
0.00
54.59 ± 4.08
65.46 ± 3.21
656.16 ± 97.38
260 ± 42

20.34 ± 5.83
230.86 ± 6.15
168.18 ± 24.03
393.93 ± 44.31
77 ± 32
0.00
3.58 ± 0.47
21.47 ± 2.28
528.02 ± 31.07
430 ± 29
16.00 ± 4.73
49.50 ± 3.31
57.88 ± 7.71
366.92 ± 22.83
494 ± 23
0.00
77.45 ± 5.75
156.17 ± 16.50
462.44 ± 43.06
242 ± 32

214

Sample
Zn
F1
F2
F3
F4
F5

SRC
669.33 ± 6.37
64.83 ± 7.99
21.94 ± 3.79
30.23 ± 1.36
100 ± 0

Exp. No.1
19.04 ± 3.45
53.34 ± 2.28
144.35 ± 10.93
510.34 ± 12.46
327 ± 22

Exp. No.2
0.49 ± 0.12
23.00 ± 1.57
77.82 ± 1.49
321.73 ± 21.15
729 ± 21

Exp. No.3
12.54 ± 2.73
51.03 ± 0.57
128.81 ± 5.09
576.48 ± 44.89
351 ± 37

Exp. No.4
3.67 ± 0.30
25.12 ± 1.89
38.61 ± 0.88
632.79 ± 102.75
437 ± 57

Exp. No.5
18.16 ± 4.62
55.80 ± 3.74
109.18 ± 10.08
446.92 ± 26.90
399 ± 25

Table B1. Concentrations of extractable fractions of heavy metals in the oily waste (SRC) and its solid residues (continued)

Sample
Exp. No.6
Extractable fractions (mg/kg)
Cu
F1 14.03 ± 2.14
F2 221.09 ± 16.43
F3 188.99 ± 2.91
F4 411.11 ± 17.42
F5 86 ± 39
Cr
F1 0.00
F2 10.86 ± 0.99
F3 43.22 ± 2.58
F4 703.96 ± 33.79
F5 259 ± 30
Ni
F1 6.51 ± 1.49
F2 37.81 ± 2.12
F3 59.89 ± 3.80
F4 401.43 ± 12.19

Exp. No.7

Exp. No.8

Exp. No.9

Exp. No.10

1.10 ± 0.34
142.00 ± 4.17
91.01 ± 0.28
477.13 ± 64.31
175 ± 11
0.00
1.27 ± 0.07
5.29 ± 0.29
179.95 ± 3.63
791 ± 3
7.97 ± 2.69
27.79 ± 0.44
24.52 ± 0.55
495.67 ± 25.79

1.23 ± 0.12
87.10 ± 2.74
109.98 ± 14.72
532.35 ± 3.73
209 ± 14
0.00
1.28 ± 0.04
5.34 ± 1.12
230.58 ± 17.65
800 ± 19
5.97 ± 1.31
25.45 ± 1.00
38.61 ± 7.82
498.35 ± 6.08

4.19 ± 0.44
89.27 ± 4.37
150.47 ± 4.31
497.10 ± 35.35
245 ± 26
0.00
1.63 ± 0.02
6.70 ± 0.88
552.05 ± 24.92
527 ± 26
3.04 ± 0.60
11.15 ± 0.31
25.77 ± 2.88
373.09 ± 27.28

1.21 ± 0.16
102.40 ± 6.86
120.55 ± 8.86
516.80 ± 91.48
200 ± 8
0.00
1.22 ± 0.07
5.72 ± 0.65
296.06 ± 11.84
736 ± 12
2.90 ± 0.34
16.55 ± 1.31
28.34 ± 3.46
469.01 ± 1.26

215

Sample
Pb

Zn

F5
F1
F2
F3
F4
F5
F1
F2
F3
F4
F5

Exp. No.6
513 ± 20
0.00
107.27 ± 6.31
186.62 ± 6.17
497.94 ± 34.78
178 ± 22
15.05 ± 2.49
64.43 ± 4.04
136.32 ± 11.47
490.85 ± 23.39
358 ± 5

Exp. No.7
424 ± 22
0.00
47.61 ± 1.13
154.73 ± 4.50
309.30 ± 11.39
421 ± 6
0.86 ± 0.16
35.00 ± 0.36
124.39 ± 6.99
278.39 ± 19.58
585 ± 12

Exp. No.8
4701 ± 4
0.00
56.02 ± 0.78
114.94 ± 21.23
352.60 ± 20.68
466 ± 41
0.83 ± 0.04
23.00 ± 0.71
86.16 ± 19.79
347.46 ± 19.67
628 ± 40

Exp. No.9
677 ± 24
0.00
53.56 ± 1.76
73.84 ± 5.50
532.85 ± 2.20
378 ± 9
2.94 ± 0.41
25.77 ± 0.46
46.46 ± 5.71
519.63 ± 3.25
544 ± 10

Exp. No.10
524 ± 4
0.00
64.01 ± 5.46
116.44 ± 13.55
401.07 ± 10.22
409 ± 18
0.86 ± 0.12
27.89 ± 2.21
87.20 ± 12.87
372.23 ± 5.62
599 ± 5

Table B1. Concentrations of extractable fractions of heavy metals in the oily waste (SRC) and its solid residues (continued)

Sample
Exp. No.11
Extractable fractions (mg/kg)
Cu
F1 32.23 ± 4.16
F2 247.78 ± 16.02
F3 187.37 ± 7.36
F4 337.29 ± 0.21
F5 69 ± 5
Cr
F1 0.00
F2 10.49 ± 0.84
F3 58.34 ± 0.24

Exp. No.12

Exp. No.13

Exp. No.14

Exp. No.15

2.83 ± 0.42
188.36 ± 1.08
162.35 ± 47.40
432.79 ± 4.64
97 ± 12
0.00
2.82 ± 0.14
11.74 ± 1.33

4.87 ± 0.31
99.32 ± 4.37
173.49 ± 11.63
489.38 ± 13.83
163 ± 2
0.00
1.81 ± 0.05
8.16 ± 0.58

4.48 ± 0.28
108.12 ± 7.48
165.76 ± 18.31
472.23 ± 36.07
167 ± 11
0.00
2.01 ± 0.20
8.16 ± 0.62

2.98 ± 0.39
108.48 ± 9.94
149.17 ± 1.16
493.32 ± 13.57
158 ± 25
0.00
2.32 ± 0.17
8.33 ± 0.54

216

Sample

Ni

Pb

Zn

F4
F5
F1
F2
F3
F4
F5
F1
F2
F3
F4
F5
F1
F2
F3
F4
F5

Exp. No.11
625.04 ± 1.15
270 ± 2
32.95 ± 3.45
48.05 ± 2.92
72.46 ± 2.42
339.20 ± 5.70
473 ± 3
4.35 ± 0.32
149.04 ± 8.92
223.15 ± 10.93
455.56 ± 6.02
87 ± 8
42.75 ± 4.58
76.56 ± 4.70
145.86 ± 18.47
453.45 ± 10.08
290 ± 8

Exp. No.12
457.33 ± 1.80
502 ± 1
6.73 ± 0.35
33.88 ± 0.32
40.23 ± 5.21
380.91 ± 9.92
515 ± 5
0.00
68.85 ± 0.51
150.58 ± 23.23
409.90 ± 11.36
300 ± 11
5.35 ± 0.28
51.32 ± 0.63
102.68 ± 18.60
414.52 ± 21.50
446 ± 3

Exp. No.13
530.77 ± 14.62
529 ± 14
2.75 ± 0.32
12.77 ± 1.23
31.07 ± 2.01
354.27 ± 7.02
648 ± 5
0.00
60.91 ± 4.90
82.98 ± 4.69
546.30 ± 26.51
338 ± 17
2.97 ± 0.11
30.10 ± 1.67
57.43 ± 1.87
515.58 ± 4.50
509 ± 1

217

Exp. No.14
567.19 ± 79.83
513 ± 79
3.89 ± 0.63
14.20 ± 1.79
31.59 ± 2.40
351.00 ± 28.00
646 ± 25
0.00
57.87 ± 7.72
82.69 ± 5.19
537.51 ± 60.50
347 ± 48
3.03 ± 0.24
29.84 ± 3.23
56.51 ± 1.86
503.28 ± 35.63
512 ± 31

Exp. No.15
531.51 ± 9.11
508 ± 8
4.05 ± 0.05
16.85 ± 7.01
31.31 ± 1.60
371.17 ± 12.00
614 ± 15
0.00
65.04 ± 6.60
82.95 ± 2.68
522.86 ± 13.04
325 ± 22
3.50 ± 0.01
38.15 ± 3.08
57.80 ± 4.23
512.39 ± 14.88
476 ± 22

Figure B1. FT-IR spectra of the feedstock and solid residues after pyrolysis (SRC: solid residue sample from control experiment without pyrolysis
treatment; SROx: solid residue from oily waste pyrolysis alone at X (°C) temperature and reaction time = 75 min; SROHx: solid residue from copyrolysis of oily waste and hog fuel at X (°C) temperature and reaction time = 75 min and hog fuel addition = 20%

218

APPENDIX C
Table C1. Sequential extraction results of OS500-B

Concentration (mg/kg)
Fractions

Cd

Pb

Cr

Cu

Ni

Zn

F1: Water soluble

–a

–

–

–

–

–

F2: Exchangeable

–

–

–

5.9

–

–

F3: Bound to carbonate

–

–

8.5

54.5

–

115.5

F4: Bound to Fe-Mn oxides

–

–

–

27.9

–

–

F5: Bound to organic matter

–

–

–

68.5

–

71.7

F6: Residual

–

–

8.5

2.5

1.6

2.4

Below detection limit.
F1–F5 detection limit (mg/L): Cd (0.06), Pb (0.9), Cr (0.06), Cu (0.1), Ni (0.2) Zn (0.3);
F6 detection limit (mg/kg): Cd (0.6), Pb (9), Cr (0.6), Cu (1), Ni (2), and Zn (3).

a

Table C2. Sequential extraction results of OS500-Pb and OS500-Cd

Fractions

OS500-Pb

OS500-Cd

mg/g

mg/g

%
F1: Water soluble
0.00
–
F2: Exchangeable
0.45
1.38 ± 0.52
F3: Bound to carbonate
303.42 ± 11.90 98.06
F4: Bound to Fe-Mn oxides 4.51 ± 0.35
1.46
F5: Bound to organic matter –
0.00
F6: Residual
0.11 ± 0.01
0.04
a

Below detection limit.
F1–F5 detection limit (mg/L): Cd (0.06), Pb (0.9);
F6 detection limit (mg/kg): Cd (0.6), Pb (9).

a

219

%
0.11
0.02 ±0.00
2.93 ± 0.09 15.12
16.31 ± 0.67 84.15
0.12 ± 0.01 0.62
0.00
–
–
0.01

Table C3. Recovery of sequential extraction results of OS500-Pb and OS500-Cd

Sample

Sum from each step
in sequential
extraction (mg/g)

Sorbed amount
from equilibrium
study (mg/g)

Recovery
(%)

OS500-Pb

309.41 ± 12.18

351.48 ± 2.08

88.03

OS500-Cd

19.38 ± 0.75

23.95 ± 3.79

80.92

12.0

pHfinal

11.5
11.0
10.5
10.0
9.5

2

4

6

8

10

12

14

pHinitial
Figure C1. Determination of pHPZC for OS500. Thick straight line represents pHinitial = pHfinal line.

220

500

a

Qt (mg/g)

400
y=254.790+7.564x
R2=0.986

300
y=60.113+106.781x
R2=0.955

200
100
0

5

10

15

20

25

t1/2 (h1/2)
30

b

Qt (mg/g)

25
y=21.248+0.155x
R2=0.775

20
y=10.104+2.394x
R2=0.887

15
10
0

5

10

15

20

25

t1/2 (h1/2)
Figure C2. Intraparticle diffusion plot for the sorption of (a) Pb2+ and (b) Cd2+ on OS500 (0.1g char in
30 mL solution (0.01 M NaNO3), initial Pb2+ (or Cd2+) concentration 1600 (1000) mg/L; room
temperature 24 °C; initial solution pH 5).

221

400

Qe (mg/g)

360
320

Temkin
Dubinin-Radushkevich

280
240
0

2000

4000

6000

8000

10000

Ce (mg/L)

Model
Equation
Plot
a
b
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

LangmuirT (User)
y=a*b*x/(1+b*x)
Qe
373.19033
19.77826
614.93275
0.83945
0.79931

Model
Equation
Plot
Kf
n
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

Freundlich (User)
Kf*x^(1/n)
Qe
273.81433
25.08163
393.80285
0.89718
0.87148

Model
Equation
Plot
Klf
nlf
alf
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

LFsips (User)
(Klf*x^nlf)/(1+(alf*x)^nlf)
Qe
472.32798
0.07654
0.01177
513.06631
0.89953
0.83256

Model
Equation
Plot
K
a
b
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

RedlichPeterson (User)
(K*x)/(1+a*x^b)
Qe
62493.79342
221.91343
0.96368
521.75065
0.89783
0.82972

Model
Equation
Plot
A
b
Reduced Chi-Sqr
R-Square(COD)
Adj. R-Square

Temkin (User)
(8.314*297.15)/(b)*(Ln(A*x))
Qe
4.32493E9
198.01794
386.58714
0.89907
0.87383

24

Qe (mg/g)

22
20
Temkin
Dubinin-Radushkevich

18
16
14
0

200

400

600

800

1000

Ce (mg/L)
Figure C3. Isotherms of (1) Pb2+ and (2) Cd2+ sorption onto OS500 (0.1g char in 30 mL solution (0.01
M NaNO3); contact time 128 h; room temperature 24 °C; initial solution pH 5).

222

APPENDIX D
Table D1. Sequential extraction results of OSDCs before Pb2+/Cd2+sorption

Sample
OS300-B
OS500-B
Extractable fractions (mg/kg)
Cd F1 –a
–
F2 –
–
F3 –
–
F4 –
–
F5 –
–
F6 –
–
Pb F1 –
–
F2 –
–
F3 –
–
F4 –
–
F5 –
–
F6 –
–
Cr F1 –
–
F2 –
–
F3 8.4
8.5
F4 –
–
F5 –
–
F6 7.6
8.5
Cu F1 –
–
F2 –
5.9
F3 53.7
54.5
F4 53.8
27.9
F5 38.0
68.5
F6 2.2
2.5
Ni F1 –
–
F2 –
–
F3 –
–
F4 –
–
F5 –
–
F6 3.1
1.6

OS700-B

OS900-B

OS500N100-8M-B

–
–
–
–
–
–
–
–
–
–
–
–
–
–
12.3
16.6
–
8.9
–
–
36.7
77.3
34.0
3.1
–
–
31.0
–
–
–

–
–
–
–
–
–
–
–
–
–
–
–
–
–
8.7
24.5
–
5.0
–
–
16.9
64.3
60.7
2.0
–
–
–
–
–
1.4

–
–
–
–
–
–
–
–
–
–
–
–
–
–
18.6
17.1
–
5.1
–
–
43.8
106.7
30.4
1.5
–
–
–
–
–
1.4

223

Sample
OS300-B
OS500-B
Extractable fractions (mg/kg)
Zn F1 –
–
F2 –
–
F3 79.1
115.5
F4 –
–
F5 64.8
71.7
F6 7.4
2.4

OS700-B

OS900-B

OS500N100-8M-B

–
–
104.9
–
77.4
–

–
–
125.5
–
–
4.1

–
–
28.3
–
–
1.8

Below detection limit.
F1–F5 detection limit (mg/L): Cd (0.06), Pb (0.9), Cr (0.06), Cu (0.1), Ni (0.2) Zn (0.3);
F6 detection limit (mg/kg): Cd (0.6), Pb (9), Cr (0.6), Cu (1), Ni (2), and Zn (3).

a

Table D2. Sequential extraction results of Pb-loaded/Cd-loaded OSDCs

Chars
OS300

OS500

OS700

OS900

Fractions
1
2
3
4
5
6
1
2
3
4
5
6
1
2
3
4
5
6
1
2
3
4

Loaded with Pb2+
mg/g

%

Loaded with Cd2+
mg/g
%

–a
1.71 ± 0.13
205.22 ± 1.38
2.86 ± 0.04
–
0.04 ± 0.00
–
1.38 ± 0.52
303.42 ± 11.90
4.51 ± 0.35
–
0.11 ± 0.01
0.29 ± 0.06
0.44 ± 0.21
203.16 ± 6.91
13.64 ± 1.53
0.39 ± 0.02
0.56 ± 0.08
2.10 ± 0.39
2.97 ± 0.20
30.44 ± 2.23
2.83 ± 0.51

0.00
0.81
97.81
1.36
0.00
0.02
0.00
0.45
98.06
1.46
0.00
0.04
0.13
0.20
92.99
6.24
0.18
0.26
5.46
7.73
79.30
7.37

0.41 ± 0.08
6.29 ± 0.33
23.56 ± 1.72
0.18 ± 0.02
–
–
0.02 ±0.00
2.93 ± 0.09
16.31 ± 0.67
0.12 ± 0.01
–
–
0.02 ± 0.00
8.79 ± 1.16
72.73 ± 2.75
0.82 ± 0.60
0.06 ± 0.01
0.04 ± 0.00
0.33 ± 0.07
3.00 ± 0.25
6.17 ± 0.19
0.55 ± 0.02

224

1.34
20.66
77.40
0.59
0.00
0.00
0.11
15.12
84.15
0.62
0.00
0.01
0.03
10.65
88.21
0.99
0.07
0.05
3.29
29.78
61.35
5.42

Chars

OS500N100-8M

Fractions

Loaded with Pb2+

5
6
1
2
3
4
5
6

mg/g
–
0.05 ± 0.00
–
0.76 ± 0.57
257.90 ± 7.91
11.58 ± 0.18
0.25 ± 0.02
0.03 ± 0.00

Below detection limit.
Fraction 1–5 detection limit (mg/L): Cd (0.06), Pb (0.9);
Fraction 6 detection limit (mg/kg): Cd (0.6), Pb (9)

a

Loaded with Cd2+
%
0.00
0.14
0.00
0.28
95.34
4.28
0.09
0.01

mg/g
–
0.02 ± 0.00
–
4.71 ± 0.25
78.29 ± 6.24
1.22 ± 0.02
–
–

Table D3. Recovery of sequential extraction results of Pb-loaded/Cd-loaded OSDCs

Sum from each step

Sorbed amount

in sequential

from equilibrium

extraction (mg/g)

study (mg/g)

OS300-Pb

209.82 ± 1.29

282.96 ± 3.22

74.15

OS500-Pb

309.41 ± 12.18

351.48 ± 2.08

88.03

OS700-Pb

218.48 ± 8.58

261.14 ± 2.02

83.66

OS900-Pb

38.38 ± 2.76

49.68 ± 2.31

77.25

OS500N100-8M-Pb

270.52 ± 7.91

310.00 ± 2.19

87.26

OS300-Cd

30.44 ± 1.99

46.67 ± 0.69

65.22

OS500-Cd

19.38 ± 0.75

23.95 ± 3.79

80.92

OS700-Cd

82.46 ± 3.88

92.14 ± 3.11

89.49

OS900-Cd

10.06 ± 0.33

11.05 ± 1.50

91.04

OS500N100-8M-Cd

84.22 ± 6.30

90.06 ± 3.17

93.52

Sample

225

Recovery
(%)

%
0.00
0.15
0.00
5.59
92.96
1.45
0.00
0.00

Table D4. Pearson correlation coefficient matrix between OSDCs properties and Pb2+/Cd2+ sorption
mechanism

SBET

a

PV

Ash%

CEC

H/C

Qp,Pb

Qt,Pb

Qcm,Cda

Qcf,Cd

Qt,Cd

SBET

1.00

PV

0.88

1.00

Ash%

‒0.08

0.16

1.00

CEC

0.73

0.97*

0.34

1.00

H/C

0.82

0.95

‒0.12

0.90

1.00

Qp, Pb

0.48

0.10

‒0.85

‒0.15

0.25

1.00

Qt,Pb

0.68

0.32

‒0.73

0.07

0.43

0.97*

1.00

Qcm,Cda

0.89

1.00**

0.11

0.95*

0.96*

0.15

0.36

1.00

Qcf,Cd

0.81

0.99*

0.14

0.98*

0.97*

0.04

0.25

0.99*

1.00

Qt,Cd

0.83

0.99**

0.12

0.97*

0.97*

0.08

0.29

0.99**

1.00** 1.00

OS500N100-8M was not included; * Significant (P < 0.05); ** Highly significant (P < 0.01).

226

Figure D1. XRD patterns of OSDCs before Pb2+/Cd2+ sorption (OSDC-B). Minerals with peaks labeled: M, Muscovite 2M2,
K0.77Al1.93(Al0.5Si3.5)O10(OH)2; B, Barite, BaSO4; Q, Quartz, SiO2; C, Calcite, CaCO3; D, Dolomite, CaMg(CO3)2; O, Calcium oxide, CaO; Cl, Celsian,
BaAl2Si2O8; Bo, Barium oxide hydrate, BaO2 (H2O)8; F, Ferritungstite, (W, Fe)(O,OH)3; W, Witherite, BaCO3.
227

Figure D2. FTIR spectra of OSDC before Pb2+/Cd2+ sorption (OSDC-B).
228