BIODEGRADATION OF NONYLPHENOLS, DETERMINATION OF
DEGRADATION PRODUCTS AND DETECTION OF RESPONSIBLE
MICROORGANISMS USING MOLECULAR TECHNIQUES
A THESIS SUBMITTED TO
THE GRADUATE SCHOOL OF NATURAL AND APPLIED SCIENCES
OF
MIDDLE EAST TECHNICAL UNIVERSITY
BY
FADİME KARA MURDOCH
IN PARTIAL FULFILLMENT OF THE REQUIREMENTS
FOR
THE DEGREE OF DOCTOR OF PHILOSOPHY
IN
BIOTECHNOLOGY
JULY 2015
Approval of the Thesis:
BIODEGRADATION OF NONYLPHENOLS, DETERMINATION OF
DEGRADATION PRODUCTS AND DETECTION OF RESPONSIBLE
MICROORGANISMS USING MOLECULAR TECHNIQUES
submitted by FADİME KARA MURDOCH in partial fulfilment of the
requirements for the degree of Doctor of Philosophy in Biotechnology
Department, Middle East Technical University by,
Prof. Dr. Gülbin Dural Ünver
Dean, Graduate School of Applied and Natural Sciences
Prof. Dr. Filiz Dilek
Head of Department, Biotechnology
Prof. Dr. F. Dilek Sanin
Supervisor, Environmental Engineering Dept.
Prof. Dr. G. Candan Gürakan
Co-Supervisor, Food Engineering Dept.
Examining Committee Members
Prof. Dr. Gülay Özcengiz
Biology Dept., METU
Prof. Dr. F. Dilek Sanin
Environmental Engineering Dept., METU
Prof. Dr. Hilal Özdağ
Biotechnology Institute, Ankara University
Assoc. Prof. Dr. Can Özen
Biotechnology Dept., METU
Assist. Prof. Dr. Yeşim Soyer
Food Engineering Dept., METU
Date :
28.07.2015
I hereby declare that all information in this document has been obtained and
presented in accordance with academic rules and ethical conduct. I also
declare that, as required by these rules and conduct, I have fully cited and
referenced all material and results that are not original to this work.
Name, Last name: Fadime Kara Murdoch
Signature:
iv
ABSTRACT
BIODEGRADATION OF NONYLPHENOLS, DETERMINATION OF
DEGRADATION PRODUCTS AND DETECTION OF RESPONSIBLE
MICROORGANISMS USING MOLECULAR TECHNIQUES
Kara Murdoch, Fadime
Ph.D., Department of Biotechnology
Supervisor: Prof. Dr. F. Dilek Sanin
Co-Supervisor: Prof. Dr. G. Candan Gürakan
July 2015, 321 pages
Nonylphenol(poly)ethoxylates (NPnEO) have received special attention during the
last decades due to their toxic and endocrine-disrupting effects for many organisms.
They have limited degradation in wastewater treatment plants especially in
activated sludge units. Since they accumulate in sludge, understanding of their fate
during sludge treatment is important. With this motivation, this research aimed to
monitor the degradation of nonylphenol diethoxylate (NP2EO, as a parent
compound) into its degradation products, nonylphenol monoethoxylate (NP1EO)
and nonylphenol (NP), in lab-scale semi-continuous anaerobic digesters.
Determination and quantification of the degradation products were carried out with
gas
chromatography-mass
spectrometry
(GC/MS).
Fluorescence
in
situ
hybridization (FISH) and quantitative PCR (qPCR) were used to determine the
abundance and composition of four sub-groups of Proteobacteria and acetoclastic
methanogens in the total microbial community of anaerobic digesters in the
v
presence and absence of NP2EO. Moreover, enrichment studies were performed in
order to isolate diverse bacterial strains able to use branched NP as the sole carbon
and energy source.
Results showed that NP2EO degraded slowly under anaerobic conditions and NP
accumulated as a final product in digesters. Digester performance monitored by
solids reduction, chemical oxygen demand (COD) removal and methane production
showed no deterioration due to NP2EO spike. Beta- and Gammaproteobacteria
constituted the predominant sub-groups of Proteobacteria within the NP2EOspiked sludge community of anaerobic digesters according to FISH and qPCR
results. The dominant proliferation of the Methanosaeta (62.5-77%) member of
acetoclastic methanogens was determined in the all operated anaerobic digesters.
Keywords:
Anaerobic
Digester,
Nonylphenol
Diethoxylate,
Nonylphenol
Monoethoxylate, Nonylphenol, Fluorescence in situ Hybridization, Quantitative
PCR
vi
ÖZ
NONİLFENOLLERİN BİYOLOJİK PARÇALANMASI, PARÇALANMA
ÜRÜNLERİN TAYİNİ VE SORUMLU MİKROORGANİZMALARIN
MOLEKÜLER METOTLARLA TESPİTİ
Kara Murdoch, Fadime
Doktora, Biyoteknoloji Bölümü
Tez Yöneticisi: Prof. Dr. F. Dilek Sanin
Ortak Tez Yöneticisi: Prof. Dr. G. Candan Gürakan
Temmuz 2015, 321 sayfa
Nonilfenol(poli)etoksilatlar (NPnEO) birçok organizma için toksik ve endokrin
yapısını bozucu özellikleri nedeniyle günümüzde yaygın olarak incelenmektedir.
Bu kimyasallar, atıksu arıtma tesislerinde sınırlı seviyede biyolojik olarak
parçalanabilse de, çamur yapısında birikme eğilimlerinden dolayı çamur arıtım
sürecindeki akıbetlerinin anlaşılması önem taşımaktadır. Bu amaçla, bu çalışmada
nonilfenol dietoksilatın (seçilen ana bileşik, NP2EO) laboratuvar ölçekli yarısürekli anaerobik çürütücülerde parçalanma mekanizmalarının belirlenmesi ve
parçalanma ürünleri olan nonilfenol (NP) ve nonilfenol monoetoksilatların
izlenmesi (NP1EO) hedeflenmiştir. Bu parçalanma ürünlerinin belirlenmesi ve
ölçülmesi
Gaz
Kromotografisi/Kütle
Spektrometrisi
(GC/MS)
ile
gerçekleştirilmiştir. Ayrıca çalışmada floresan in situ hibridizasyon (FISH) ve
vii
kantitatif PCR (qPCR) ile anaerobik çürütücülerin mikrobiyal komunitesini
oluşturan Proteobakteri grubuna ait dört alt Proteobakteri grubu ve asetoklastik
metanojenlerin miktar ve dağılımı belirlenmiştir. Aynı zamanda NP2EO eklenen
anaerobik çürütücülerde bu filogenetik grupların NP2EO ve parçalanma ürünlerinin
varlığında tepkileri de incelenmiştir. Ayrıca sıvı besiyerinde zenginleştirme
çalışmalarında dallanmış-yapıdaki NP’yi tek karbon ve enerji kaynağı olarak
kullanabilen farklı bakteri suşları izole edilmiştir.
Yapılan bu çalışmada NP2EO’nun anaerobik koşullar altında yavaş parçalandığı ve
son ürün olan NP’nin anaerobik çürütücülerde biriktiği gösterilmiştir. Çürütücülerin
performansları katı madde ve kimyasal oksijen ihtiyacı (COD) giderimleri ve metan
üretimi parametreleri ile takip edildiğinde NP2EO eklenmesiyle herhangi bir
kötüleşme olmadığı gözlenmiştir. Beta- and Gammaproteobakteri’lerin NP2EO
eklenen anaerobik çürütücülerde dominant Proteobakteri alt-grupları oldukları
FISH ve qPCR çalışmaları ile belirlenmiştir. Asetoklastik methanogen grubuna ait
Methanosaeta (62.5-77%) cinsinin tüm işletilen anaerobik çürütücülerde dominant
yayılımı belirlenmiştir.
Anahtar Kelimeler: Anaerobik Çürütücü, Nonilfenol Dietoksilat, Nonilfenol
Monoetoksilat, Nonilfenol, Floresan in situ Hibridizasyon, Kantitatif PCR
viii
To my beloved family and
lovely husband, Robert
ix
ACKNOWLEDGEMENTS
I would like to thank and express my gratitude to my supervisor, Prof. Dr. F. Dilek
Sanin for her valuable guidance and encouragement throughout of my thesis
research. I will always greatful to her for giving a chance to me even though I have
a different research background. Also, I would like to thank my co-supervisor, Prof.
Dr. G. Candan Gürakan, for encouragement and support.
I would also like to acknowledge the examining committee members Prof. Dr.
Gülay Özcengiz, Prof. Dr. Hilal Özdağ, Assoc. Prof. Dr. Can Özen and Assist.
Prof. Dr. Yeşim Soyer for their valuable suggestions and contributions.
I would like to sincerely thank Assoc. Prof. Dr. Anthony Hay for giving me the
chance be a member of his laboratory in Cornell University and providing all the
necessary lab supplies to improve my skills in molecular and microbiological
applications. I am thankful to Prof. Dr. G. Candan Gürakan, Prof. Dr. Cihangir
Tanyeli, Assoc. Prof. Dr. Tuba H. Ergüder Bayramoğlu and Assist. Prof. Dr. Robert
Murdoch who let me use the instruments and equipment in their laboratories.
I would like to thank all members of the Environmental Engineering Department
for embracing and supporting me in this new field. I am also grateful to Mehmet
Hamgül and Mehmet Dumanoğulları for their friendship and technical assistance. I
am grateful to Güldane Kalkan, Cemalettin Aydın, Kutlay Öztürk and Gülçin
Özsoy for their encouragement and support.
I am grateful to my friends for their encouragement and support during this
research: Çiğdem Kıvılcımdan Moral, Hakan Moral, Güray Doğan, Mihriban
Civan, Firdes Yenilmez, Deniz Genç, Eylem Doğan, Hande Bozkurt, Muneer
Ahmad and Seçil Ömeroğlu. I am also thankful to my friends in the Biology
Depertment Tuğba Özaktaş, Aysun Türkanoğlu Özçelik and Tuba Çulcu for their
endless support, encouragement and help. Also, I would like to thank Hanh
x
Nguyen, my best friend, for her help and support in the Hay Lab. Through our
friendship, I experienced the cultural similarities between Vietnam and Turkey:) I
am also thankful to Eda Çelik Akdur, Özlem Durmuş and Berrak for their endless
support during my time in USA. They never let me get home-sick, we had
wonderful times in Ithaca, NY
I am also greatful to my all office-mates, especially Sema Yurdakul, Berkay Çelebi
and Didem Civancık for their endless patience, support and encouragement. I am
also thankful to Ebru Koçak for always being near me whenever I needed.
I would like to dedicate the thesis to my parents Halil and Gülseren Kara, my sister,
Gülşah Kara Buz and brother-in-law Adnan Buz; and my brother Yasin Kara and
sister-in-law, Habibe Kara. Also I would like to dedicate the thesis to my beautiful
nieces Ilgın Buz, Zeyno Kara and Zehra Kara. I would like to express my deepest
appreciation to them for their endless support, love, understanding and patience not
only during this study but also throughout my life.
I would like to express my deepest gratitude and love to my husband, Robert
Murdoch. Without his support, I could never have completed this research. I spent
most of my time in laboratory after our marriage and he was always there when I
needed. I am also greatful to him for sharing his experience in molecular biology
and microbiology with me. Thank you for your guidance, patience and love!! Now,
I promise we will spend time together away from laboratory I also would like to
thank my parents-in-law, Bruce and Carol Murdoch for accepting me into their
family without hesitation and offering their endless support, love and
encouragement.
I would like to thank The Scientific and Technological Research Council of Turkey
(TÜBİTAK) for providing the financial support for this research. I am also thankful
to TUBITAK BIDEP program no. 2211 and 2214 for their financial support for my
research in Turkey and abroad.
xi
I
doing
am
that
always
which
I
cannot do, in order that
I may learn how to do
it.
Pablo
Picasso
xii
TABLE OF CONTENTS
ABSTRACT ............................................................................................................... v
ÖZ ............................................................................................................................ vii
ACKNOWLEDGEMENTS ....................................................................................... x
LIST OF TABLES .................................................................................................. xix
LIST OF FIGURES ................................................................................................ xxi
CHAPTERS
1. INTRODUCTION ................................................................................................. 1
1.1. Background Information ................................................................................. 1
1.2. Purpose of the Study ....................................................................................... 3
1.3. Scope of the Study .......................................................................................... 5
1.4. Organization of Thesis .................................................................................... 6
2. LITERATURE REVIEW ...................................................................................... 7
2.1. Wastewater Sludge and Anaerobic Digestion................................................. 7
2.2. Nonylphenol and Nonylphenol Polyethoxylates .......................................... 10
2.2.1. Structural and Physico-chemical Properties ........................................... 11
2.2.2. Existence of NP Compounds in Environment ........................................ 14
2.2.2.1. Water and Wastewater ......................................................................... 15
2.2.2.2. Sludge from Wastewater Treatment Plants ......................................... 17
2.2.3. Toxicity................................................................................................... 18
2.3. Biodegradation of Nonylphenol Compounds ............................................ 22
2.3.1. Degradation of the Ethoxylate Side-Chain ........................................ 22
2.3.2. Biodegradation of Nonylphenol ......................................................... 27
2.3.3. Metabolism of Hydroquinone ............................................................ 30
2.4. Molecular Approaches to Phylogenetic Identification of Microorganisms .. 32
2.4.1. Characterization of Environmental Microbial Communities by
Molecular Analyses .......................................................................................... 34
2.4.1.1.1. Theory and Methodology of FISH ................................................... 35
2.4.1.1.2. Advantages of FISH ......................................................................... 37
xiii
2.4.1.1.3. Limitations of FISH .......................................................................... 39
2.4.1.1.4. Application of FISH to Studying the Impact of NP and NPnEO on
Microbial Communities .................................................................................... 39
2.4.1.1.4.1. Nonylphenol polyethoxylates ................................................... 39
2.4.1.1.4.2. Nonylphenol .............................................................................. 41
2.4.1.2. Quantitative Polymerase Chain Reaction (qPCR)............................... 43
3. MATERIALS AND METHODS ......................................................................... 47
3.1. Sludge Samples ............................................................................................. 47
3.2. Chemicals...................................................................................................... 47
3.3. Experimental Set-up ..................................................................................... 48
3.3.1. Set-up of Laboratory Scale Semi-Continuous Digesters ........................ 48
3.3.2. Operation of Laboratory Scale Semi-Continuous Digesters .................. 49
3.4. Analytical Methods ....................................................................................... 53
3.4.1. pH Measurement .................................................................................... 53
3.4.2. Solids Determination .............................................................................. 54
3.4.3. Chemical Oxygen Demand (COD) ........................................................ 54
3.4.4. Total Gas Production .............................................................................. 55
3.4.6. Identification and Quantification of Nonylphenol Compounds ............. 56
3.4.6.1. The Principles of GC/MS................................................................ 56
3.4.6.2. GC/MS method for determination of NP compounds .................... 56
3.4.6.3. Derivatization .................................................................................. 58
3.4.6.4. Extraction of NP Compounds from Solid Phase of Sludge Samples61
3.4.6.5. Extraction of NP Compounds from Liquid Phase of Sludge
Samples ........................................................................................................ 63
3.4.6.6. Recovery ......................................................................................... 64
3.4.7. Quality Assurance/Quality Control (QA/QC) ........................................ 66
3.4.7.1. Glassware ........................................................................................ 66
3.4.7.2. Linearity .......................................................................................... 66
3.4.7.3. Repeatability ................................................................................... 67
3.4.7.4. Limit of Detection and Quantification (LOD and LOQ) ................ 68
3.5. Molecular Analyses ...................................................................................... 69
xiv
3.5.1. DNA Extraction ...................................................................................... 69
3.5.2. Polymerase Chain Reaction (PCR)......................................................... 71
3.5.3. Quantitative PCR (Real-Time PCR) ...................................................... 72
3.5.3.1. Primers ............................................................................................ 72
3.5.3.2. qPCR Assays ................................................................................... 73
3.5.3.3. Standard Curve Preparation ............................................................ 75
3.5.4. Fluorescence in Situ Hybridization (FISH) ............................................ 80
3.5.4.1. FISH probes .................................................................................... 80
3.5.4.2. Preparation of Gel Coated slides..................................................... 82
3.5.4.3. Fixation ........................................................................................... 82
3.5.4.4. Dehydration ..................................................................................... 82
3.5.4.5. Hybridization .................................................................................. 83
3.5.4.6. Visualization ................................................................................... 84
3.5.4.7. Image Processing and Analysis....................................................... 84
3.5.5. Enrichment Studies ................................................................................. 85
3.5.5.1. Chemicals ........................................................................................ 85
3.5.5.2. Media .............................................................................................. 86
3.5.5.4. DNA Extraction .............................................................................. 87
3.5.5.5. Restriction Enzyme Analysis .......................................................... 88
3.5.5.6. Sequencing and Phylogenetic Analysis .......................................... 88
3.5.5.7. Biodegradation Experiments ........................................................... 89
3.5.5.7.1. Synthesis of Single NP Isomers…………………………... ....89
3.5.5.7.2. Growth and Biodegradation Assays......................................... 94
3.5.5.7.3. Extraction of NP Compounds .................................................. 95
4. DETERMINATION OF NONYLPHENOL COMPOUNDS IN SLUDGE
SOLIDS AND LIQUIDS: METHOD DEVELOPMENT AND APPLICATION .. 97
4.1. Introduction ................................................................................................... 97
4.2. Materials and Methods ................................................................................ 100
4.2.1. Sludge samples ..................................................................................... 100
4.2.2. Chemicals ............................................................................................. 101
4.2.3. Optimization of Derivatization Method for NP Compounds ............... 101
xv
4.2.4. Optimization of GC/MS Method for detection of NP compounds....... 109
4.2.5. Optimization of Extraction Methods for NP compounds ..................... 123
4.2.5.1. Extraction from Solid Samples ..................................................... 123
4.2.5.2. Extraction from Aqueous Samples ............................................... 128
4.2.6. Recovery ............................................................................................... 130
4.2.7. Quality Assurance/Quality Control (QA/QC) ...................................... 132
4.2.7.1. Linearity ........................................................................................ 132
4.2.7.2. Repeatability ................................................................................. 132
4.2.7.3. Limit of Detection and Quantification (LOD and LOQ) .............. 133
4.3. Method Application Studies ....................................................................... 134
4.4. Conclusion .................................................................................................. 137
5. BIODEGRADATION OF NONYLPHENOL DIETHOXYLATE IN LABSCALE ANAEROBIC DIGESTERS .................................................................... 139
5.1. Introduction ................................................................................................. 139
5.2. Materials and Methods................................................................................ 142
5.2.1. Sludge Samples .................................................................................... 142
5.2.2. Chemicals ............................................................................................. 143
5.2.3. Digester Set-up ..................................................................................... 143
5.2.4. Instrumentation ..................................................................................... 145
5.2.5. Derivatization ....................................................................................... 146
5.2.6. Extraction ............................................................................................. 147
5.2.6.1. Extraction from solid phase .......................................................... 147
5.2.6.2. Extraction from liquid phase ......................................................... 147
5.2.7. Analytical Methods .............................................................................. 148
5.3. Results and Discussion ............................................................................... 148
5.3.1. pH ......................................................................................................... 148
5.3.2. Biogas Production and Composition .................................................... 150
5.3.3. Chemical Oxygen Demand................................................................... 153
5.3.4. Total Solids and Volatile Solids ........................................................... 155
5.3.5. Total Suspended Solids and Volatile Suspended Solids ...................... 158
5.3.6. Digester Performance ........................................................................... 160
xvi
5.3.7. Nonylphenol Compounds ..................................................................... 162
5.3.8. Biotic Control Digesters ....................................................................... 162
5.3.8.1. Solid Phase of Biotic Control Digesters ....................................... 163
5.3.8.2. Liquid Phase of Biotic Control Digesters ..................................... 167
5.3.9. Spiked Digesters ................................................................................... 169
5.3.9.1. Solid Phase of Spiked Digesters ................................................... 170
5.3.9.2. Liquid Phase of Spiked Digesters ................................................. 174
5.3.10. Abiotic Control Digesters ................................................................... 177
5.4. Conclusion .................................................................................................. 181
6.
ASSESSMENT
OF
ANAEROBIC
MICROBIAL
COMMUNITY
STRUCTURE IN THE PRESENCE OF NONYLPHENOL DIETHOXYLATE. 185
6.1. Introduction ................................................................................................. 185
6.2. Materials and Methods ................................................................................ 188
6.2.1. Laboratory-scale Anaerobic Digesters ................................................. 188
6.2.2. Sampling of Sludge for Microbial Analysis ......................................... 189
6.2.2.1. Fluorescence in situ Hybridization189
6.2.2.2. Quantitative PCR .......................................................................... 197
6.2.2.2.1. DNA Extraction ..................................................................... 197
6.2.2.2.2. Optimization of Quantitative PCR Assays ............................ 198
6.2.3. Analytical Methods .............................................................................. 205
6.3. Results and Discussion ............................................................................... 205
6.3.1. Examination of Microbial Community Structure by FISH and qPCR . 206
6.4. Conclusions ................................................................................................. 238
7. ISOLATION AND CHARACTERIZATION OF NOVEL BACTERIAL
STRAINS CAPABLE OF DEGRADATION OF NONYLPHENOL ISOMERS 241
7.1. Introduction ................................................................................................. 241
7.2. Materials and Methods ................................................................................ 244
7.2.1. Chemicals ............................................................................................. 244
7.2.2. Enrichment Studies ............................................................................... 245
7.2.3. Identification and Characterization of Bacterial Isolates ..................... 246
7.2.4. Biodegradation Assays ......................................................................... 247
xvii
7.2.5. Extraction and Quantification of NP Compounds ................................ 248
7.3. Results and Discussion ............................................................................... 249
7.3.1. Isolation and Characterization of Bacterial Strains .............................. 249
7.3.2. Biodegradation Tests ............................................................................ 252
7.4. Conclusions ................................................................................................. 261
8. CONCLUSIONS................................................................................................ 263
9. RECOMMENDATIONS ................................................................................... 269
REFERENCES ...................................................................................................... 271
APPENDIX A ........................................................................................................ 307
APPENDIX B ........................................................................................................ 311
APPENDIX C ........................................................................................................ 315
CURRICULUM VITAE ........................................................................................ 317
xviii
LIST OF TABLES
Table 3.1. Solids concentrations used in digester set-up ......................................... 49
Table 3.2. Initial parameters following digester set-up ........................................... 50
Table 3.3. Parameters and measurement frequency at different time intervals of
operation period. ................................................................................................. 53
Table 3.4. GC/MS method for determination of NP, NP1EO and NP2EO ............. 57
Table 3.5. GC/MS method for determination of NP1EC......................................... 58
Table 3.6. Quantitative calibration data for NP compounds .................................... 67
Table 3.7. Performance of DNA extraction kits evaluated in terms ofconcentration
and purity ............................................................................................................ 71
Table 4.1. Details of GC/MS and oven programs of studies applied in preliminary
experiments ....................................................................................................... 110
Table 4.2. GC/MS methods tested with different oven programs for NPE
compounds ........................................................................................................ 117
Table 4.3. GC/MS methods experimented for NP1EC analysis ............................ 120
Table 4.4. Retention times and target ions of the NP compounds ......................... 122
Table 4.5. Effect of extraction solvents on the mean recoveries (%) of target
compounds in sludge samples obtained from 8-min sonication-assisted
extraction ........................................................................................................... 125
Table 4.6. The mean recoveries (%) of NP compounds in spiked water samples
after solid phase extraction with different solvents........................................... 130
Table 4.7. Recoveries and RSDs obtained from extraction of spiked NP
compounds from solid phases ........................................................................... 131
Table 4.8. Recoveries and RSDs obtained from extraction of spiked NP
compounds from aqueous phases ...................................................................... 131
Table 4.9. Quantitative calibration data for NP compounds ................................. 132
Table 4.10. Repeatability studies at two different concentrations of NP
compounds ........................................................................................................ 133
xix
Table 4.11. Limits of detection (LOD) and limits of quantification (LOQ) values
for the developed extraction methods ............................................................... 134
Table 5.1. Percent COD removals for biotic control and spiked digesters ........... 155
Table 5.2. % TS and VS removals for biotic control and spiked digesters ........... 158
Table 5.3. % TSS and VSS removals for biotic control and spiked digesters ....... 160
Table 5.4. Evaluation of digester performance of biotic control and spiked
digesters............................................................................................................. 161
Table 5.5. Concentrations of NP compounds in sludge samples measured at
digester set-up and prior to spike for biotic control digesters ........................... 163
Table 5.6. Concentrations of NP compounds measured in sludge samples at
digester set-up and prior to spike for spiked digesters ...................................... 169
Table 6.1. Oligonucleotide probes used for FISH analysis ................................... 195
Table 6.2. Taxon-specific primers used in qPCR assays. ...................................... 201
Table 7.1. m/z values used for identification and quantification of each compound ..
................................................................................................................................ 249
Table 7.2. Morphological and biochemical characteristics of the isolated bacterial
isolates ............................................................................................................... 252
xx
LIST OF FIGURES
Figure 2.1. Generalized scheme for anaerobic digestion metabolic processes (van
Haandel and Van Der Lubbe 2007). .......................................................................... 9
Figure 2.2. Structure of NP compounds (adapted from (Tanghe et al. 2000) and
(Díaz et al. 2002)). ................................................................................................... 13
Figure 2.3. Structural similarities between two different 4-NP isomers and 17-βEstradiol (a human estrogen) and diethylstilbestrol (a synthetic estrogen)
(Adapted from (Montgomery-Brown and Reinhard 2003). ..................................... 20
Figure 2.4. Generalized scheme for biodegradation of NPnEO under aerobic and
anaerobic conditions ................................................................................................ 23
Figure 2.5. Suggested pathways for aerobic degradation of nonylphenol
ethoxylates.1: (John and Whilte, 1998; 2: (Nguyen and Sigollot, 1997); 3:
(Franska et al., 2003); 4: (Jonkers et al., 2001), (adapted from Soares et al, 2008).24
Figure 2.6. Integrated scheme for nonylphenol polyethoxylate metabolism.
Adapted from (Giger et al. 2009). ............................................................................ 26
Figure 2.7. Metabolism of α-quaternary nonylphenol isomers by Sphingobium
xenophagum Bayram and by Sphingomonas sp. TTNP3. From Kolvenbach et al.
(2014). ...................................................................................................................... 29
Figure 2.8. Metabolism of α-tertiary nonylphenol isomers by Sphingobium
xenophagum Bayram (from Kolvenbach et al., 2014). ............................................ 30
Figure 2.9. Metabolism of hydroquinone by Sphingomonas sp. TTNP3
(Kolvenbach and Corvini 2012)............................................................................... 32
Figure 3.1. Configuration of lab-scale anaerobic semi-continuous digester
connected to gas collection cylinder. ....................................................................... 50
Figure 3.2. Details of digester operation for spike and microbial activity............... 51
Figure 3.3. Details of derivatization procedure applied for NP, NP1EO, NP2EO
compounds ............................................................................................................... 59
Figure 3.4. Details of derivatization procedure applied for NP1EC. ....................... 60
xxi
Figure 3.5. Details of sonication-assisted extraction method applied to solid phase
of sludge samples for extraction of NP compounds ................................................ 62
Figure 3.6. Application steps for SPE method for extraction of NP compounds
from liquid phases of sludge samples ...................................................................... 63
Figure 3.7. Visual illustration of SPE method ......................................................... 64
Figure 3.8. Derivative melting curve analysis for standards in qPCR ..................... 74
Figure 3.9. Blue-white screening specific for β-Proteobacteria ............................. 76
Figure 3.10. Agarose gel images following colony PCR with i) vector-specific
T7F/M13R primer pair (left side) ii) insert-specific primer pair (right side) for βProteobacteria. L: DNA Ladder (Gene Ruler, 100 bp plus) ................................... 78
Figure 3.11. A standard curve constructed for quantification of abundance of αProteobacteria in sludge samples ............................................................................. 79
Figure 3.12. Visual representation of FISH analysis ............................................... 85
Figure 3.13. Chromatographic illustration of p363NP with GC/MS analysis. ........ 92
Figure 3.14. Chromatographic illustration of p353NP with GC/MS analysis. ........ 93
Figure 4.1. GC/MS chromatograms of NP silyl-derivatized in 100 μL of
BSTFA+10%TMCS and BSTFA+33%TMCS ...................................................... 104
Figure 4.2. GC/MS chromatograms of NP1EO silyl-derivatized in 100 and 200
μL of BSTFA+1%TMCS and BSTFA+10%TMCS .............................................. 104
Figure 4.3. GC/MS chromatogram of NP, NP1EO and NP2EO silyl-derivatized in
100 and 200 μL of BSTFA+1%TMCS .................................................................. 105
Figure 4.4. a) A GC/MS chromatogram of methylated NP1EC given in the study
of Diaz et al. (Díaz et al. 2002) b) A GC/MS chromatogram of methylated
NP1EC following derivatization in our lab using the method of Diaz et al. (Díaz
et al. 2002). ............................................................................................................ 106
Figure 4.5. a) A GC/MS chromatogram of propyl esters of NP1EC given in the
study of Ding and Tzing (Ding and Tzing 1998) b) A GC/MS chromatogram of
propyl esters of NP1EC following derivatization in our lab using the method of
Ding and Tzing (Ding and Tzing 1998). ................................................................ 107
Figure 4.6. a) A GC/MS chromatogram of methyl esters of NP1EC given in the
study of Lee et al. (Lee et al. 1997) b) GC/MS chromatogram of methyl esters of
xxii
NP1EC obtained in our study following derivatization method suggested by Lee
et al. (Lee et al. 1997). ........................................................................................... 108
Figure 4.7. Comparison of total ion chromatograms of NP obtained from
application of oven programs given in studies of Lian et al. (Lian et al. 2009),
Gatidou et al. (Gatidou et al. 2007), and Gibson et al. (Gibson et al. 2005).......... 112
Figure 4.8. Comparison of total ion chromatograms of NP obtained from
application of oven programs given in studies of Richter et al. (Richter et al.
2009) and Lu et al. (Lu et al. 2008). ...................................................................... 113
Figure 4.9. Comparison of total ion chromatograms of NP obtained from
application of oven programs given in studies of Benanou et al. (2007), Donghao
et al. (2004) and Barber et al. (2000). .................................................................... 113
Figure 4.10. Total ion chromatogram of NP obtained from application of oven
program given in the study of Diaz et al. (2000) ................................................... 114
Figure 4.11. Total ion chromatogram of NP1EO obtained from application of
oven program given in the study of Diaz et al. (2000). ......................................... 115
Figure 4.12. Total ion chromatogram of NP1EO obtained from application of
oven program given in the study of Gatidou et al. (2007). .................................... 115
Figure 4.13. Total ion chromatogram of NP2EO obtained from application of
oven program given in the study of Diaz et al. (2002) .......................................... 116
Figure 4.14. Comparison of total ion chromatograms of NP obtained from the
application of different GC/MS methods ............................................................... 118
Figure 4.15. Comparison of total ion chromatograms of NP obtained from
application of GC/MS oven program of Method 1 and Method 7 ......................... 119
Figure 4.16. Total ion chromatogram of NP1EC obtained from application of
oven program given in the study of Lee et al. (1997) ............................................ 121
Figure 4.17. Example for chromatographic view of GC/MS analysis for NP
compounds with application of optimized GC/MS oven program ........................ 122
Figure 4.18. Example for chromatographic view of GC/MS analysis of NP1EC . 123
Figure 4.19. Total ion chromatograms obtained for the extraction of target
compounds from solid samples using different solvents ....................................... 124
xxiii
Figure 4.20. Mean recoveries (%) of NPE compounds obtained for sonicationassisted extraction method at different sonication times with acetone .................. 126
Figure 4.21. Mean recoveries (%) of NPE compounds for extraction by
mechanical shaking extraction method at different extraction times and
combination of extraction methods in the presence of acetone ............................. 127
Figure 4.22. Concentrations of NP, NP1EO, NP2EO and sum of NPE in
dewatered sludge samples ...................................................................................... 136
Figure 5.1. Structure of NPnEOs and potential degradation products ................... 140
Figure 5.2. pH distribution vs. time for feed and operated anaerobic digesters .... 150
Figure 5.3. Cumulative methane production vs. time for operated anaerobic
digesters ................................................................................................................. 151
Figure 5.4. COD change with respect to time for operated anaerobic digesters.... 154
Figure 5.5. Change in TS and VS concentrations vs. time for operated digesters. 156
Figure 5.6. Change in TS and VS concentrations vs. time for operated digesters. 159
Figure 5.7. Change in NP compounds concentration (solid phase) vs. time for BD1 (biotic control) digester ....................................................................................... 164
Figure 5.8. Change in NP compounds concentration (solid phase) vs. time for BD2 (biotic control) digester ....................................................................................... 165
Figure 5.9. Change in NP compounds concentration (liquid phase) vs. time for
BD-1 (biotic control) digester ................................................................................ 167
Figure 5.10. Change in NP compounds concentration (liquid phase) vs. time for
BD-2 (biotic control) digester ................................................................................ 168
Figure 5.11. Change in NP compounds concentration (solid phase) vs. time for
SD-1 (spiked digester) ........................................................................................... 170
Figure 5.12. Change in NP compounds concentration (solid phase) vs. time for
SD-2 (spiked digester) ........................................................................................... 171
Figure 5.13. Change in NP compounds concentration (liquid phase) vs. time for
SD-1 (spiked digester) ........................................................................................... 175
Figure 5.14. Change in NP compounds concentration (liquid phase) vs. time for
SD-2 (spiked digester) ........................................................................................... 176
xxiv
Figure 5.15. Change in NP compounds concentration (solid phase) vs. time for
AD-1 (abiotic control digester) .............................................................................. 178
Figure 5.16. Change in NP compounds concentration (solid phase) vs. time for
AD-2 (abiotic control digester) .............................................................................. 179
Figure 5.17. Change in NP compounds concentration (liquid phase) vs. time for
AD-1 (abiotic control digester) .............................................................................. 180
Figure 5.18. Change in NP compounds concentration (liquid phase) vs. time for
AD-2 (abiotic control digester) .............................................................................. 181
Figure 6.1. Sonication procedure in ultrasonic bath at different sonication times.
Bright-field microscope view, 40X objective. ....................................................... 191
Figure 6.2. Sonication procedure with sonication 3-mm-probe at different time
periods. Bright-field microscope view, 40X objective. ......................................... 193
Figure 6.3. Epifluorescence photographs showing in situ hybridization for
Bacteria domain. a) DAPI-staining b) EUB338MIX-Cy3 c) Overlay of DAPI and
EUB338MIX-Cy3. (66th day, spiked digester), Bar = 5 µm, 6300X ..................... 208
Figure 6.4. Epifluorescence photographs showing in situ hybridization for
Archaea domain a) DAPI-staining b)overlay of DAPI and ARC915-Cy3 c)
processed image of overlay of DAPI and ARC915-Cy3 d) ARC915-Cy3. (103rd
day, biotic control digester), Bar = 5 µm, 6300X .................................................. 209
Figure 6.5. Epifluorescence photographs showing in situ hybridization for Alphasubgroup. a) DAPI-staining b) overlay of DAPI and ALF968-Cy3 c) processed
image of overlay of DAPI and ALF968-Cy3 d) ALF968-Cy3. (83rd day, spiked
digester). Bar = 5 µm, 6300X ................................................................................ 210
Figure 6.6. Epifluorescence photographs showing in situ hybridization for Betasubgroup. a) DAPI-staining b) overlay of DAPI and BET42a-Cy3 c) processed
image of overlay of DAPI and BET42a-Cy3 d) BET42a-Cy3. (66thday, spiked
digester). Bar = 5 µm, 6300X ................................................................................ 211
Figure 6.7. Epifluorescence photographs showing in situ hybridization for
Gamma-subgroup. a) DAPI-staining b) overlay of DAPI and Gam42a-Cy3 c)
processed image of overlay of DAPI and Gam42a-Cy3 d) Gam42a-Cy3. (71st
day, spiked digester). Bar = 5 µm, 6300X ............................................................. 212
xxv
Figure 6.8. Epifluorescence photographs showing in situ hybridization for Deltasubgroup. a) DAPI-staining b) overlay of DAPI and SRB385-Cy3 c) processed
image of overlay of DAPI and SRB385-Cy3 d) SRB385-Cy3. (83rd day, spiked
digester). Bar = 5 µm, 6300X ................................................................................ 213
Figure 6.9. Epifluorescence photographs showing in situ hybridization for
Methanosaeta genus. a) DAPI-staining b) overlay of DAPI and MX825-Cy3 c)
processed image of overlay of DAPI and MX825-Cy3 d) MX825-Cy3. (124thday,
control digester). Bar = 5 µm, 6300X .................................................................... 214
Figure 6.10. Epifluorescence photographs showing in situ hybridization for
Methanosaeta genus. a) DAPI-staining b) overlay of DAPI and MX825-Cy3 c)
processed image of overlay of DAPI and MX825-Cy3 d) MX825-Cy3. (124thday,
control digester). Bar = 5 µm, 6300X .................................................................... 215
Figure 6.11. Epifluorescence photographs showing in situ hybridization for
Methanosarcina genus. a) DAPI-staining b) overlay of DAPI and MS821-Cy3 c)
processed image of overlay of DAPI and MS821-Cy3 d) MS821-Cy3. (103rd day,
spiked digester). Bar = 5 µm, 6300X ..................................................................... 216
Figure 6.12. Overall distribution of Bacteria and Archaea domains in operated
semi-continuous anaerobic digesters determined by FISH. ................................... 218
Figure 6.13. Overall distribution of Bacteria and Archaea domains in operated
semi-continuous anaerobic digesters determined by qPCR assays........................ 218
Figure 6.14. Relative abundance of Alphaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization using ALF968 oligonucleotide probe. ............. 219
Figure 6.15. Relative abundance of Alphaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by qPCR
assays. .................................................................................................................... 221
Figure 6.16. Relative abundance of Betaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization using BET42a oligonucleotide probe .............. 222
xxvi
Figure 6.17. Relative abundance of Betaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by qPCR
assays ..................................................................................................................... 223
Figure 6.18. Relative abundance of Gammaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization using GAM42a oligonucleotide probe ............ 225
Figure 6.19. Relative abundance of Gammaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by qPCR
assays ..................................................................................................................... 226
Figure 6.20. Relative abundance of Deltaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization using SRB385 oligonucleotide probe .............. 228
Figure 6.21. Relative abundance of Deltaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by qPCR
assays ..................................................................................................................... 229
Figure 6.22. Relative abundance of Methanosaeta in total microbial community of
biotic control and spiked anaerobic digesters analyzed by fluorescence in situ
hybridization using MX825 oligonucleotide probe ............................................... 231
Figure 6.23. Relative abundance of Methanosaeta in total microbial community of
biotic control and spiked anaerobic digesters analyzed by qPCR assays .............. 232
Figure 6.24. Relative abundance of Methanosarcina in total microbial community
of biotic control and spiked anaerobic digesters analyzed by fluorescent in situ
hybridization using MS821 oligonucleotide probe. ............................................... 233
Figure 6.25. Relative abundance of Methanosarcina in total microbial community
of biotic control and spiked anaerobic digesters analyzed by qPCR assays .......... 234
Figure 6.26. Overall distribution of Methanosaeta and Methanosarcina in
Archaea domain of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization. ......................................................................... 236
Figure 6.27. Overall distribution of Methanosaeta and Methanosarcina in
Archaea domain of biotic control and spiked anaerobic digesters analyzed by
qPCR assays. .......................................................................................................... 237
xxvii
Figure 7.1. Enrichment studies in the presence of tNP .......................................... 250
Figure 7.2. Isolated bacterial strains on LB agar plates ......................................... 251
Figure 7.3. Demonstration of MSM-tNP flasks inoculated with isolated bacterial
strains at 7th day of biodegradation assay .............................................................. 253
Figure 7.4. Degradation of tNP by bacterial isolates in MSM medium ................ 254
Figure 7.5. Degradation of tNP by bacterial isolates in MSM medium ................ 255
Figure 7.6. Total ion chromatogram (TIC) belonging to the control sample at 7th
day. 4-tert-butylphenol: internal standard .............................................................. 255
Figure 7.7. The synthesized and purified p353NP and p363NP ............................ 256
Figure 7.8. The synthesized and purified p353NP and p363NP ............................ 257
Figure 7.9. Overlaid total ion chromatograms obtained for p353NP and p363NP at
1st, 3rd, and 7th analysis days................................................................................... 258
Figure 7.10. Biodegradation test bottles belonging to positive control and bacterial
isolates (4th day of operation), p363NP as a sole carbon source............................ 260
Figure 7.11. Biodegradation test bottles representing positive control and bacterial
isolates (4th day of operation), p363NP as a sole carbon source............................ 261
Figure A.1. Calibration curve for 4-n-NP .............................................................. 307
Figure A.2. Calibration curve for NP..................................................................... 307
Figure A.3. Calibration curve for NP1EO ............................................................. 308
Figure A.4. Calibration curve for NP2EO ............................................................. 308
Figure A.5. Calibration curve for NP1EC ............................................................. 309
Figure A.6. Calibration curve for tNP ................................................................... 309
Figure A.7. Calibration curve for p353NP............................................................. 310
Figure A.8. Calibration curve for p363NP............................................................. 310
Figure B.1. A standard curve constructed for quantification of abundance of αProteobacteria in sludge samples .......................................................................... 311
Figure B.2. A standard curve constructed for quantification of abundance of βProteobacteria in sludge samples .......................................................................... 311
Figure B.3. A standard curve constructed for quantification of abundance of
Bacteria domain in sludge samples ....................................................................... 312
xxviii
Figure B.4. A standard curve constructed for quantification of abundance of γProteobacteria in sludge samples .......................................................................... 312
Figure B.5. A standard curve constructed for quantification of abundance of
Archaea domain in sludge samples ........................................................................ 313
Figure B.6. A standard curve constructed for quantification of abundance of δProteobacteria in sludge samples .......................................................................... 313
Figure B.7. A standard curve constructed for quantification of abundance of
Methanosaeta in sludge samples............................................................................ 314
Figure B.8. A standard curve constructed for quantification of abundance of
Methanosarcina in sludge samples ........................................................................ 314
Figure C.1. 1H and 13C NMR spectrum of p353NP ............................................... 315
Figure C.2. 1H and 13C NMR spectrum of p363NP ............................................... 316
xxix
ABBREVIATIONS
AD:
Abiotic control digester
ADS:
Anaerobically digested sludge
COD:
Chemical oxygen demand
dm:
dry matter
EDC:
Endocrine disrupting compound
FISH:
Fluorescence in situ Hybridization
F/M:
Food/Microorganisms
GC:
Gas chromatography
GC/MS:
Gas chromotograpty/Mass spectrometry
HRT:
Hydraulic retention time
LB:
Luria-Bertani
LOD:
Limit of detection
LOQ:
Limit of quantification
MSM:
Minimal Salt Medium
NP:
Nonylphenol
NP1EC:
Nonylphenoxy acetic acid
NP1EO:
Nonylphenol monoethoxylate
NP2EC:
Nonylphenoxyethoxy acetic acid
NP2EO:
Nonylphenol diethoxylate
NPE:
NP+NP1EO+NP2EO
NPnEC:
Nonyl phenoxycarboxylic acids
NPnEO:
Nonylphenol polyethoxylates
OD:
Optical density
qPCR:
Quantitative PCR
PBS:
Phosphate buffer saline
p353NP:
4(3’,5’-dimethyl-3’-heptyl)phenol
p363NP:
4(3’,6’-dimethyl-3’-heptyl)phenol
xxx
SIM:
Selective ion mode
SPE:
Solid phase extraction
SRT:
Solids retention time
tNP:
Technical NP mixture
TS:
Total solids
TSS:
Total suspended solids
VS:
Volatile solids
VSS:
Volatile suspended solids
WAS:
Waste activated sludge
WWTP:
Wastewater treatment plant
xxxi
CHAPTER 1
INTRODUCTION
1.1. Background Information
Alkylphenol polyethoxylates (APnEOs) are among the most widely used nonionic
surfactants in the world (Thiele et al. 1997). They have been used for decades in
formulations of detergents, emulsifiers, dispersants, dyes (in cosmetics, cleaning,
petroleum, pulp and paper production, plastics and textiles manufacturing industries)
and pesticides (Talmage 1994). The most significant commercial alkylphenol
ethoxylates
are
nonylphenol
polyethoxylates
(NPnEO)
and
octylphenol
polyethoxylates (OPnEO). Approximately 80% of alkylphenols are constituted by
NPnEO and 15-20% are OPnEO (Renner 1997, Staples et al. 1999).
Due to widespread use of nonylphenol (NP) compounds in household and industrial
applications, these compounds reach surface waters (streams, rivers, lakes, ocean
etc.) at high concentrations through direct disposal or discharge from wastewater
treatment plants (Ahel and Giger 1993, Ahel et al. 1994, Ahel et al. 1994, Gibson et
al. 2005). The studies revealed that NP compounds have toxic, carcinogenic and
estrogenic effects (Soto et al. 1991, Jobling and Sumpter 1993, White et al. 1994,
Cox 1996, Rodgers-Gray et al. 2001). Because of the structural similarities between
nonylphenol and estrogens, nonylphenol mimics natural hormone and competes for
the estrogen binding receptors in vertebrates (White et al. 1994, Lee and Lee 1996,
1
Porter and Hayden 2002, Soares et al. 2008). Due to this property, these compounds
have been described as endocrine disrupting compounds (EDC). It has been reported
that NP compounds lead to the feminization of male reproductive organs and serious
metabolic problems (Roy et al. 1998, Rodgers-Gray et al. 2001, Birkett and Lester
2002, Folmar et al. 2002). Human exposure to NP compounds has been evidenced
by detecting them in breast milk (Ademollo et al. 2008), umbilical cord blood (Chen
et al. 2008) and urine (Calafat et al. 2005). Some studies indicated the possibility of
adverse effects on sperm quality and quantity in humans exposed to NP (Toppari et
al. 1996, Swan et al. 1997). These findings have increased the concerns related to the
long-term effects of NP on wildlife and human health.
Due to low water solubility and high log KOW values, NP compounds accumulate in
cells, tissues (Coldham et al. 1998) and solid particles. They have been detected at
large concentrations in river sediments and sewage sludge (Bennie et al. 1997,
Marcomini et al. 2000, Ying et al. 2002, Santos et al. 2007, Mao et al. 2012) Due to
their persistence and estrogenic effects, removal of NP compounds by
biodegradation in environmental systems have gained importance in recent years.
Biological degradation of nonylphenol polyethoxylates is possible; degradation starts
by losing ethoxy groups leading to the formation of more toxic and persistent
metabolites: nonylphenol (NP), nonylphenol monoethoxylate (NP1EO), nonylphenol
diethoxylate (NP2EO), more water-soluble nonylphenoxy acetic acid (NP1EC) and
nonylphenoxyethoxy acetic acid (NP2EC) (Ahel et al. 1994, Di Corcia et al. 1998,
Maguire 1999, Ying et al. 2002, Montgomery-Brown and Reinhard 2003, Gonzalez
et al. 2007, Koh et al. 2008, Nagarnaik et al. 2010). Studies related to the fate of NP
compounds focused on determination of NP, NP1EO, NP2EO, NP1EC and NP2EC
rather than NPnEOs and NPnECs due to the aforementioned rapid biodegradation of
the parent compounds.
2
Owing to environmental and health concerns associated with NP compounds, they
have been listed as hazardous substances by the OSPAR Commission (OSPAR,
1998), and listed among the priority substances in the European Water Framework
Directive (EU Directive 2000/60/EC). Additionally, European Union proposed a
limit
value
in
“Working
Document
on
Sludge-3rd
Draft
for
NPEs
(NP+NP1EO+NP2EO) as 50 mg/kg dry solids for land application of sludge. Turkey
recently set the same value as the regulatory limit in ”Regulation for the Use of
Municipal and Urban Sludge on Land”. Setting up a limit value for NPEs prior to
land application of sewage sludge have brought a need for rapid and reliable
methods for extraction and simultaneous detection of this family of compounds in
sewage sludge. Since anaerobic digesters are the most commonly used stabilization
techniques for sewage sludge, the proposed regulatory limit for land application in
European countries (including Turkey) necessitates the documentation of
degradation mechanisms of nonylphenol compounds in anaerobic systems. Besides,
the composition and change of microbial community structure in anaerobic digesters
in the presence of NP compounds have not been throughly studied and have critical
importance for the management of these persistent chemicals. Isolation of pure
culture microorganisms that are capable of degrading nonylphenol compounds will
provide additional remediation prospects.
1.2. Purpose of the Study
With all these insights, the overall purpose of the study is to observe the degradation
of target NP compounds in biological anaerobic digesters by monitoring abundance
and diversity of Proteobacteria sub-groups and acetoclastic methanogens
represented by Methanosarcina and Methanosaeta genera with molecular methods.
The specific aims of this study can be summarized as follows:
3
1. To develop reproducible and reliable methods for the extraction of NP,
NP1EO, NP2EO, NP1EC and 4-n-NP compounds from liquid and solid
phases of sludge samples,
2. To develop a GC/MS method for the detection and quantification of NP,
NP1EO, NP2EO, NP1EC and 4-n-NP in liquid and solid phases of sludge
samples,
3. To apply the proposed methods for the determination and seasonal
monitoring of NP, NP1EO and NP2EO compounds in dewatered sewage
sludge samples taken from a real wastewater treatment plant,
4. To set lab-scale semi-continuous anaerobic digesters to monitor the
degradation of NP2EO (parent compound), determine the fate of NP2EO and
degradation products during the operation of lab-scale anaerobic digesters,
5. To investigate the change in abundance and diversity of four sub-groups of
Proteobacteria in the presence of NP compounds during the biodegradation
of NP2EO in anaerobic digesters by fluorescence in situ hybridization (FISH)
and quantitative PCR (qPCR) molecular tools. Besides that, to reveal the
effect of NP2EO and its degradation products on methanogenic activity of
digesters by monitoring the abundance of acetoclastic methanogens
represented
by
Methanosarcina
and
Methanosaeta
genera
with
aforementioned molecular methods.
6. To isolate phylogenetically diverse new bacterial strains able to use branched
NP as a sole carbon and energy source and investigate their behavior in the
presence of technical NP (tNP) and single isomers including p353NP and
p363NP.
4
1.3. Scope of the Study
To attain these objectives, the scope of the study consisted of following research
tasks:
1. Method Development and Optimization: For extraction of NPEs from solid
and liquid phases of sludge samples, extensive method development studies
were conducted at the end of which sonication-assisted extraction and solid
phase extraction methods were chosen and applied with satisfactory
recoveries. The simultaneous determination of NPEs in both phases of sludge
samples was performed by GC/MS following development of an oven
program yielding the best peak quality and abundance. With development
and optimization of extraction and quantification methods, concentration of
NPEs in dewatered sludge samples taken from Ankara Central WWTP have
been monitored for a year by monthly sampling.
2. Degradation of NP Compounds in Lab-Scale Anaerobic Digesters: NP2EO
was chosen as the model parent compound to observe the degradation and the
formation of degradation products in lab-scale digesters simulating actual
anaerobic sludge digesters. NP2EO-spiked lab-scale semi-continuous
anaerobic sludge digesters were operated. Continuous operation and
monitoring of target parameters were done on both spiked and control
digesters over a period of 147 days.
3. Determination of Microbial Community Structure: Sludge samples taken
from NP2EO-spiked and biotic control digesters were used to monitor
microbial community structure at taxonomic level. Two molecular tools,
FISH and qPCR were used to determine the relative abundance of each sub-
5
group of Proteobacteria and methanogens including Methanosarcina and
Methanosaeta.
4. Isolation of New Bacterial Strains Degrading NP: Activated sludge samples
taken from the return sludge line of Ankara Central WWTP has been used for
enrichment studies. Following enrichment, different bacterial strains able to
grow in the presence of branched NP compounds were isolated and tested for
degradation of single isomer tNP.
1.4. Organization of Thesis
The thesis is written in nine chapters. Except for the chapters that provide general
background (Chapter 1, 2, and 3) and overall conclusions (Chapters 8 and 9),
chapters are written in a scientific article format. For this reason, Chapters 4, 5, 6 and
7 have their introduction, materials and methods, results and conclusion parts.
Chapter 1 introduces the motivation, purpose and scope of the study briefly. Chapter
2 summarizes the relevant up-to-date studies reported in literature. Chapter 3
describes the material and methods used in the study which includes details of
analytical procedures and methods applied.
Chapter 8 summarizes main conclusions obtained during overall study and Chapter 9
includes suggestions for future studies in the highlight of this study. Appendices
have supplementary information including calibration curves, figures and tables.
Due to the wide coverage of the topics studied in this thesis, Chapters 4, 5, 6 and 7
are prepared in manuscript format as stated above. Some of these chapters have
already been submitted to the journals or will be submitted for publication soon.
Therefore, some information given in the introduction and material and methods
sections of Chapters 4, 5, 6 and 7 may appear to be repeated unavoidably.
6
CHAPTER 2
LITERATURE REVIEW
2.1. Wastewater Sludge and Anaerobic Digestion
Municipal sewage treatment plants generate two outputs following the treatment
including physical, chemical, and biological processes: a liquid effluent and a semisolid residue called sewage sludge. Sludge, the by-product of wastewater treatment,
mainly consists of concentrated mixed microbial mass responsible for the clean up of
wastewater. Following extensive treatment during typical sludge treatment scheme
that follows the wastewater treatment, organics and pathogens as well as the water
content are reduced. However, the quantities are still large and this presents a
management issue for most wastewater treatment plants. The application of treated
sludge to land for agricultural or soil conditioning purposes presents a very simple
and inexpensive option for sludge disposal.
Additionally, the content rich in
nitrogen, phosphorus, and organic material, makes sludge an ideal fertilizer (Spicer
2002). However, the possibility of presence of pollutants and pathogens in sludge
has become an increasing cause for opposition to this route of disposal (Iranpour et
al. 2004).
Anaerobic digestion (AD) is regarded as a highly efficient waste disposal technology
that achieves multiple goals simultaneously. AD is very popular; there are an
estimated 36,000 AD facilities treating municipal waste in Europe, treating over 40%
of generated sludge (Tilche and Malaspina 1998). AD reduces the mass of solid
wastes, thus reducing the burden of finding appropriate disposal routes (Mata-
7
Alvarez et al. 2000)). During the AD process, nitrogen nutrients are removed
through nitrification and denitrification processes, turning potentially eutrophying
organic and mineral nitrogen into gaseous nitrogen (Abufayed and Schroeder 1986).
Ultimately, AD generates methane gas as an end-product, which can be harnessed
for electricity generation.
AD is routinely performed in very large scales for
municipal waste treatment. The metabolic processes at the core of AD are highly
complex and only partially understood however; indeed they often behave
unpredictably, giving them a reputation for unreliability (Parkin and Owen 1986).
The input to a typical AD treatment system is highly complex, consisting of
polysaccharides, proteins, and long-chain fatty acids. The initial stages of digestion
involve breaking down these large and complex molecules into much smaller and
more uniform molecules (Sanders et al. 2000). The huge variety of input chemical
necessitates an immense variety of microorganisms.
These upstream processes
typically are performed by Bacteria domain and often involve hydrolytic processes
for metabolism of complex polymers; while these processes are often not given
much attention, they may be rate-limiting in some cases (Kiely et al. 1997).
Firmicutes, Proteobacteria, Bacteroidetes and Actinobacteria are the dominant
phyla in the bacterial community of anaerobic digesters; these organisms are likely
involved in upstream metabolic processes (Chouari et al. 2005, Riviere et al. 2009,
Ito et al. 2011, Lee et al. 2012, Regueiro et al. 2012, Guo et al. 2015). Available
oxidants such as nitrate and sulfate are rapidly consumed, leaving fermentative
processes as the dominant metabolic process once options for hydrolysis are
exhausted. These fermentative pathways for breakdown of complex molecules
eventually converge on common chemicals, hydrogen and acetate. At this point,
Archaea become critical in their conversion of hydrogen and acetate to methane gas,
the final end product of the AD metabolic web (Figure 2.1).
8
Figure 2.1. Generalized scheme for anaerobic digestion metabolic processes (van
Haandel and Van Der Lubbe 2007).
The methanogenic processes catalyzed by the Archaeal community has been the
focus of a great deal of attention from researchers. This final stage of AD tends to
show a good deal of variability in efficiency, thus an understanding of the
phylogenetic groups and how they interact with process variables is critical
(Ariesyady et al. 2007, Nelson et al. 2011, Vanwonterghem et al. 2014). Two
important genera of Archaeal methanogens involved in AD are Methanosarcina and
9
Methanosaeta. Chelliapan et al. (2011) studied the relative abundances of these two
genera in an anaerobic stage reactor focused on treating an industrial wastewater
(Chelliapan et al. 2011).
It was reported that Methanosaeta dominated at low
organic loading rates (OLR), making up 70-80% of the Archeal community, while at
these conditions, Methanosarcina was a very minor player. When OLR was
increased however, the abundances reversed, with Methanosarcina becoming
strongly dominant. This finding was confirmed by Kundu et al. and by Ariesyady et
al.; again, Methanosarcina dominated at high OLR (Ariesyady et al. 2007, Kundu et
al. 2013). Conklin et al. revealed a deeper specific mechanism at work in these
trends (Conklin et al. 2006). They reported that Methanosarcina dominates at high
OLR due specifically to higher concentrations of acetate (which serves as a major
source of nutrition along with hydrogen in the case of methanotrophs). Town et al.
(2014) reported consistent results, showing that Methanosarcina abundance is
directly correlated with acetate concentrations in AD. It has been suggested that the
presence of Methanosarcina in a typical AD is very important for resisting the shock
of high OLR due to their preference and metabolic efficiency under these conditions.
2.2. Nonylphenol and Nonylphenol Polyethoxylates
Non-ionic alkylphenol polyethoxylates (APnEO) are very inexpensive chemicals
with a wide variety of uses, especially as surfactants (Soares et al. 2008). APnEO
make up approximately 40% of the worldwide surfactant market. There are two main
types of APnEO; nonylphenol polyethoxylates (NPnEO), which make up about 80 %
of total APnEO and octylphenol polyethoxylates (OPnEO) which represent the
remainder (Renner 1997) (Guenther et al. 2006). NPnEO have many desirable
properties that contribute to the popularity, such as good wetting, detergency, low
foaming, low cost, and usefullness at low temperatures (Bennie et al. 1997). Within
the surfactant use area, their main area of application is in household cleaning
formulations while use as industrial emulsifiers, especially in the textile and
pesticide sectors, is also very important. Nonylphenol production reached 154,200
10
tons/year in the USA, 73,500 tons/year in Europe, 16,500 tons/year in Japan and
16,000 tons/year in China (Soares et al. 2008).
A typical NPnEO formulation consists of isomers with 20-30 ethoxylate groups.
Under aerobic conditions, initial biodegradation of NPnEO proceeds by progressive
shortening of the ethoxylate chain, which results in the formation of mono- and
diethoxylated nonylphenols under aerobic conditions (NP1EO, NP2EO) (John and
White 1998, Potter et al. 1999, Langford et al. 2005, Lu et al. 2008). Under
anaerobic conditions, the polyethoxylate side-chain is also degraded and fully
deethoxylated and nonylphenol (NP) is produced as an end-product. These
transformations can also result in accumulation of nonylphenoxy acetic acid
(NP1EC) and nonylphenoxyethoxy acetic acid (NP2EC) as stable intermediates
especially under aeronic conditions (Kohler et al. 2008).
Becaue NPnEO are used in aqueous solutions, they are usually disposed of via
municipal and industrial wastewaters into sewage treatment plants, which present the
major point of entry for these compounds into surface waters such as lakes, rivers,
and oceans (Ahel et al. 1994, Fries and Püttmann 2003). Because the core
degradation product NP is highly recalcitrant and hydrophobic (Di Corcia et al.
1998, Shang et al. 1999, Fujita et al. 2000), it sorbs to and accumulates in soils and
sediments (Ferguson et al. 2001). These properties also lead to bioaccumulation in
plants and animals (Ahel and Giger 1993, Korsman et al. 2015, Lee et al. 2015).
Also the presence of NP has been reported in atmosphere, pristine and sludgeamended soils and food (Moeder 2000, Corvini et al. 2006).
2.2.1. Structural and Physico-chemical Properties
The NP compound family includes several different forms with differing degrees of
ethoxylation in addition to various isomers of the alkyl chain (Figure 2.2). The
different structural forms each of which possess unique chemical properties such as
11
hydrophobicity, solubility, partitioning trends etc. lead to different environmental
behavior characteristics.
The core moiety, nonylphenol, is a synthetic chemical consisting of a phenol ring
with a branched nine-carbon alkyl group in the 4-position (Figure 2.2) (Guenther et
al. 2006). Under typical conditions, it is a thick, pale-colored liquid which is water
insoluble. NP is produced at the industrial scale through the alkylation of phenol
with nonene with acidic catalysis. The final mixture (referred to as the technical
mixture) is usually composed 22 common isomers and a total of 211 structural
isomers (Guenther et al. 2006), all with slightly different chemical and toxicological
properties (Eganhouse et al. 2009, Lu and Gan 2014). These isomers differ in the
structure of the nonyl moiety and a minority of them may include the nonyl group in
the 2-position (Soares et al. 2008).
The solubility and surface active properties of the compound depend on the types
and amounts of hydrophilic and hydrophobic moieties present. For a nonionic
surfactant, this depends especially on the number of polar groups which make up
hydrophilic section of the molecule. NPnEO with less than five ethoxyate groups are
usually described as "water insoluble" or lipophilic while the higher NPnEO are
described as "water soluble" or hydrophilic (Ahel and Giger 1993, Ahel and Giger
1993). Solubility of these compounds decreases with increases in temperature.
12
OH
Linear chain nonylphenol (NP);
para isomer
HO
Linear chain nonylphenol (NP);
ortho isomer
OH
Branched chain nonylphenol (NP);
para isomer
O
Nonylphenol polyethoxylates
OH
n
(NPnEO; n = 1-20)
H19C9
O
OCH 2COOH
Nonylphenol carboxylates
n
(NPnEC; n=1, NP1EC; n=2, NP2EC)
H 19C 9
Figure 2.2. Structure of NP compounds (adapted from (Tanghe et al. 2000) and
(Díaz et al. 2002)).
The octanol/water partitioning coefficients (log Kow) of NP and NP(n=1-4)EO have
been calculated as between 4.17 and 4.48 for the NPnEO metabolites (Table 2.1),
which means that these chemicals are likely to become associated with organic
matter in sediments (Ahel and Giger 1993, Düring et al. 2002). NPnEC are likely to
be ionized at the pH values of many natural waters and their log Kow values are
expected to be much lower than those of the corresponding nonylphenol ethoxylates.
13
The pKa of NP is approximately 10.7, which is a common value for phenolic
hydroxy groups. Because it will only be deprotonated when pH is above 10, it will
remain uncharged under the vast majority of environmental conditions, which means
that it will primarily interact with the soil by hydrophobic sorption to the organic
fraction sorption by hydrogen bonding (John et al. 2000). NP displays very low
mobility, spreading only slowly in soils and sediments (Barber et al. 1988, Düring et
al. 2002). The vapour pressure and the Henry's law constant of NP are 4.55 mPa at
25oC and 11.02 Pa.m3/mol, respectively, indicating a semi-volatile organic
compound capable of some degree of exchange between the air and water phases
(Ahel and Giger, 1993).
Table 2.1. Physico-chemical properties of NPnEO (adapted from (Ahel et al. 1994)
and (Ahel and Giger 1993))
Chemical
Molecular Weight
Water Solubility
Log Kow
Koc
o
Name
NP
(g/mol)
220.3
(mg/L at 20 C)
5.43
4.48
245,470
NP1EO
264.4
3.02
4.17
288,403
NP2EO
308.5
3.38
4.21
151,336
NP3EO
352.0
5.88
4.20
74,131
NP4EO
396.2
7.65
4.30
nr
NP9EO
617.6
soluble
3.59
nr
nr: not reported
2.2.2. Existence of NP Compounds in Environment
The processes driving the fate of nonylphenolic compounds begins with their
partition behavior between different environmental compartments (surface water,
sediment, groundwater, soil, air); which is driven by the physico-chemical properties
14
of the compound in question. Because NP and its derivatives are not naturally
occurring to any degree, any occurrence of these chemicals is associated with
anthropogenic activites. The current primary route by which NP compounds enter
surface waters and sediments seems undoubtedly to be through wastewater treatment
system discharges, although stormwater run-off may also contribute some (Soares et
al. 2008, Jiang et al. 2012).
2.2.2.1. Water and Wastewater
Ahel and Giger measured NPnEC by liquid-liquid extraction with chloroform at pH
2 and gaseous stripping with nitrogen into ethyl acetate, fractionation on silica,
derivatization with boron trifluoride (BFs)/methanol or hydrochloric acid/methanol,
and analysis by GC/MS and HPLC (Ahel and Giger 1993). When they measured
these compounds treated wastewater in Switzerland, they found total NP1ECNP2EC concentrations ranging from 16-330 µg/L, considerably higher than the
metabolites NP (3-57 µg/L) and NP1EO-NP2EO (21-254 µg/L).
Clark et al. measured APnEO and APnEC in treated drinking water in New Jersey
extracted by liquid-liquid extraction with methylene chloride (Clark et al. 1992).
GC/MS analysis revealed the presence of NP1EO (0.08 µg/L) and NP2EO (0.15
µg/L), whereas HPLC/MS analysis identified NP3EO-NP7EO (0.50 µg/L) and
NP2EC-NP7EC (0.23 µg/L). Field and Reed developed a novel anion-exchange
solid phase extraction method for analyzing NPnEC based on anion-exchange solid
phase extraction, derivatization with methyl iodide to form the methyl esters, and
analysis by GC with flame ionization detection and positive chemical ionization MS
with ammonia as reagent gas (Field and Reed 1996). Their study revealed that NPEC
concentrations ranged from 140-270 µg/L in treated municipal wastewater, from 17
to 1,200 µg/L in pulp mill effluent, and from <0.10-13.5 µg/L in river water.
15
Isobe et al. (2001) reported NP concentrations in the river water ranging from 0.051
to 1.08 µg/L. These concentrations tended to be higher in warmer seasons than in
colder seasons (Isobe and Takada 2004).
In a recent study in China, NP was found in 55 out of 62 drinking water samples at a
median concentration of 27 ng/L and in all 62 surface water samples at a median
concentration of 123 ng/L (Fan et al. 2013).
Lee and Peart measured acetyl derivatives of NP and OP in treated wastewater and
sludge using GC/MS (Lee and Peart 1999). NP concentrations in Toronto
wastewater ranged from 0.8-30 µg/L and OP ranged from 0.1-2.5 µg/L. A study
applying similar methods to measure NP, OP, and NP1EO-NP2EO in the Great
Lakes basin and Upper St. Lawrence River found measurable quantities of NP and
OP in 24% of the samples (NP ranged from <0.01-0.92 µg/L and OP from <0.010.08 µg/L), NP1EO in 58% of the samples (<0.02-7.8 µg/L), and NP2EO in 32% of
the samples (<0.02-10 µg/L) (Bennie et al. 1997).
Some other recent measurements of NP in WWTP effluents include 0.38-2.07 µg/L
in Spain (Céspedes et al. 2008), 0-0.44 µg/L in China (Lian et al. 2009), and 1.3
µg/L in France (Ruel et al. 2010).
A recent meta-analysis of existing data on NP compounds in diverse systems
reported several trends (Bergé et al. 2012); in wastewater treatment systems, NP
concentrations are highly variable, from 1 to 400 µg/L, with similar concentrations
for the short ethoxylated variants. Concentrations of each of the metabolites tend to
be in the same range for separate treatment plants, as would be expected. Reasons
for the high degree of variability are unclear at this point, although it is suspected to
be linked to the amount of industrial influent that a given plant receives.
Concentrations in China tended to be higher than those in Europe and North
America, probably due to high use of NP-containing materials, low regulation, and
more industrial influent. Concentrations in effluents tend to show only limited
16
degrees of removal; in the aqueous phase, all compounds are at lower concentrations
than those found in influents, although there is a strong shift from the lowethoxylated NP to pure NP.
2.2.2.2. Sludge from Wastewater Treatment Plants
Anaerobic digestion is one of the widely used sludge treatment methods during
which the sludge is stabilized. For this reason, researchers interested in investigating
the existence and fate of NP compounds in anaerobic digesters.
Giger et al. reported that NP concentrations in anaerobically digested sludge range
from 450 to 2553 mg/kg dw while in aerobically stabilized sludge, NP
concentrations are between 80 and 500 mg/kg dw (Giger et al. 1984). This increased
concentration is mostly likely due to biodegradtion of ethoxylated forms and to the
hydrophobicity of NP driving sorbtive processes.
Marcomini et al. examined the fate of NP, NP1EO and NP2EO in soil which had
been treated with sludge (Marcomini et al. 1989). Soil concentrations were 4.7, 1.1
and 0.1 mg/kg respectively. After 100 days, their concentrations dropped to 0.5
mg/kg in NP, 0.12 mg/kg in NP1EO and 0.01 in mg/kg NP2EO. However, after 320
days, no further disappearance was observed.
Abad et al. measured NPEs (NP+NP1EO+NP2EO) in 11 different WWTPs, 7 of
which employed thermal drying and the others composting (Abad et al. 2005). In
these plants NPE concentrations ranged between 14.3-3150 mg/kg dm. For most of
samples, these NPE values exceeded the EU limit of 50 mg/kg dm.
Gonzalez et al. monitored NP, NP1EO and NP2EO levels in anaerobic digesters,
anaerobic stabilization ponds, aerobic digesters and composting plants (González et
al. 2010). NPE concentrations were 44-962 mg/kg dm in composted sludge samples
17
and 61-2319 mg/kg dm in aerobically-digested sludge samples. Many of these
concentraions were also well above EU limits.
Other reported concentrations of NP in sludge include 149-167 mg/kg in Chinese
anaerobically digested sludge (Lian et al. 2009), 0.2-46.2 mg/kg in finished Danish
sludge (Seriki et al. 2008), and 9.9 mg/kg in treated French sludge (Ruel et al. 2010).
In their meta-analysis of existing NP compound data, Berge et al. showed that the
concentration of NP as the treatment process proceeds is especially evident in solid
phase sludges, which tend to be actually strongly enriched in NP by up to an order of
magnitude over the WWTP influent, reflecting both biodegradation trends and solid
phase sorption tendencies of NP (Bergé et al. 2012).
Venkatesan and Halden found that sludges in the USA contain on average 534
mg/kg NP, which leads to an estimate total loading of 5510 tons per year,
approximately 20-60% of which is land applied (Venkatesan and Halden 2013).
2.2.3. Toxicity
The potential for nonylphenol and its compounds to cause toxicity in environmental
systems is of increasing concern. NPs are now listed as one of thirteen priority
hazardous substances by the European Union (2008/105/EC) due to these concerns.
Despite the fact that they are very useful and inexpensive compounds, their potential
for serious ecotoxicity is now clear (Lu and Gan 2014). The primary mechanism for
this toxicity is through the ability of NP to cause estrogenic responses in a variety of
aquatic organisms. The relative estrogenic potency of various NP compounds can be
ranked as; NP > NP1EO = NP2EO > NP1EC = NP2EC > NP9EO. Additionally,
nonylphenol shows a considerable potential for acute toxicity. Generally, effects of
NP can seen within 14 days of the administration of the substance when considering
the hormone mimickry route. In humans, a dermal irritation mechanism has been
18
documented as well, which can involve burning, distention, and itching following
long term exposure (Cox 1996).
A number of studies have linked WWTP outflows to estrogenic activities. Harries et
al. revealed that fish caught near wastewater treatment plant (WWTP) outfall had
abnormal ratios of sex steroid hormones and exhibited histological changes in their
reproductive organs (Harries et al. 1997). Another study was able to attribute the
feminization of male reproductive organs in fish to WWTP effluent exposure
(Rodgers-Gray et al. 2001). Estrogens are very important at many stages of animal
development and reproductive physiology and behavior; they are very potent even at
low doses (Servos 1999). Any compound that has the ability to interact with a
hormone receptor or hormone metabolic pathway therefore can exert very wideranging toxicological effects. NP has been shown to lead to high levels of
vitellogenin in male fish; vitellogenin is an egg protein that is normally only
produced at notable levels in female fish (Solé et al. 2000, Petrovic et al. 2002). NP
has been shown to induce vitellogenin production in males of several fish species
(Palermo et al. 2012) (Robertson and McCormick 2012) (Del Giudice et al. 2012)
and analgous proteins in insects (Yuan et al. 2013). Other reproductive effects
indicating estrogenic activity in fish have been demonstrated as well (Sumpter and
Jobling 2013). Studies using rainbow trout (Oncorhynchus mykiss) revealed
testicular growth inhibition at concentrations of 1 µg/L, while complete growth
inhibition was observed at 54 µg/L (Jobling et al. 1996).
A study of the toxicokinetics of isotope labelled
13
C6-nonylphenol in humans
revealed that the half-life in blood was 2-3 h. Approximately 10% of the
nonylphenol applied was found in the faeces and urine indicating that it was
absorbed in the gastrointestinal tract (Müller et al. 1998). While culturing human
breast cancer cells, Soto et al. discovered that para-nonylphenol was a weak
estrogen (Soto et al. 1991). Because of the structural similarities between some
nonylphenol isomers and natural and synthetic estrogens, it is likely that
19
nonylphenol can compete with 17β-estradiol for the estrogen binding receptor in
vertebrates (Figure 2.3). The interaction between NP and fish estrogen receptor has
been demonstrated in recombinant yeast systems (Shioji et al. 2006). NP can also
alter the kinetics of the cell cycle (Kudo et al. 2004) and produce telomeric
associations and chromosomal aberration (Roy et al. 1998), suggesting that other
mechanisms of toxicity may also exist. Even though the estrogenic activity of
nonylphenol is 100 times lower than that of natural 17β-estradiol in vivo (Folmar et
al. 2002), the high and constant concentrations of NPnEO and their metabolites
(mg/L) in WWTP effluents may lead to serious problems.
OH
OH
HO
HO
17β-estradiol
Diethylstilbesterol
(a synthetic estrogen)
HO
HO
Nonylphenol (another isomer)
Nonylphenol (one isomer)
Figure 2.3. Structural similarities between two different 4-NP isomers and 17-βEstradiol (a human estrogen) and diethylstilbestrol (a synthetic estrogen) (Adapted
from (Montgomery-Brown and Reinhard 2003).
20
The toxicity of nonylphenol to soil microorganisms was assessed by measuring the
inhibition of nitrogen fixing bacteria in a study of the impact of anthropogenic
activities on soil microflora. Nonylphenol inhibited growth and nitrification
capacities of Azobacter sp. at concentrations between 18.8 and 37.6 mg/kg
(Mårtensson and Torstensson 1996). Utilizing the ToxAlert100 biosensor system,
EC50 values for NP of 0.36 mg/L and for NP6EO of 2.7 mg/L have been reported (la
Farré et al. 2001). The EC50 value for NP6EO against Pseudomonas putida, a
common representative soil bacterium, was reported as 7 mg/L (Farré and Barceló
2001). Given that NP concentrations in land applied sewage sludge are often above
this level, a negative impact on agricultural soils is expected.
Nonylphenol was shown to have negative effects on filamentous fungi at
concentrations of 1.3 and 6.2 mg/L via uncoupling of respiration and induction of
multiple physiological effects (Karley et al. 1997). However, it has also been
reported that NP has no negative effects on fungi at environmental relevant
concentrations (Kollmann et al. 2003). Nonylphenol was also shown to be toxic to
14 different species of plant cells cultivated in vitro, with EC50 values varying from
11 to 220 mg/L (Bokern and Harms 1997). NP does not appear to be particularly
acutely toxic to animals and algae however; it was reported that LC50 for NP are
130–1400 mg/L in fish, 20–1590 mg/L in invertebrates, and 25–750 mg/L in algae
(Naylor 1994).
Nonylphenol has been found in several foodstuffs, with concentrations ranging from
0.1 to 19.4 μg/kg with an estimated daily intake of 7.5 μg/day for an adult. For
infants exclusively fed with breast milk or infant formulas daily intakes of 0.2
µg/day and 1.4 µg/day NPs, respectively, have been estimated (Guenther et al. 2002,
Raecker et al. 2011). Pork, chicken and beef contained NP at concentrations between
0.50 and 0.67 mg/kg (Ramarathnam et al. 1993). Nonylphenol contamination of food
can occur through many routes, including transfer from plastic food packaging
(Inoue et al. 2001, Fernandes et al. 2008, Muncke 2009), via the use of cleaning
21
agents in the food processing industries and application of pesticides (Guenther et al.
2002).
2.3. Biodegradation of Nonylphenol Compounds
2.3.1. Degradation of the Ethoxylate Side-Chain
Generally speaking, the ethoxylate side-chain does not pose any major obstacles to
biodegradation and is usually sequentially degraded, either by direct chain
shortening or preceded by oxidation of the exposed alcohol residue (Figure 2.4). The
route of attack is influenced by the presence of oxygen; aerobic conditions increase
the occurrence of alcohol oxidation (formation of carboxyl intermediates) although
both routes have been shown to occur under aerobic and anaerobic conditions.
Degradation of the side-chain tends to be faster under aerobic conditions (Lian and
Liu 2013).
Different studies on the pathway of side-chain removal have led to the proposal of
different models. By one scheme, the side-chain is shortened sequentially, yielding
acetaldehyde molecules and eventually NP2EO (Figure 2.5, scheme 1) (John and
White 1998, Potter et al. 1999, Langford et al. 2005, Lu et al. 2008).
In another studied scheme, NP ethoxylates are first transformed to corresponding
NPnEC by oxidation of the terminal alcohol, after which the chain is removed;
interestingly, in these described systems, the branched nonyl chain is also
carboxylated and variously degraded to a degree presumably linked to the particular
isomer in question (Di Corcia et al. 1998, Jonkers et al. 2001, González et al. 2010)
(Figure 2.5, scheme 4).
22
H19C9
O
O
OH
O
n
Nonylphenol(poly)ethoxylates (NPnEO)
Stepwise de-ethoxylation
H19C9
O
O
OH
O
n-1
H19C9
H19C9
O
O
O
OH
O
n=1-2
Nonylphenol mono- and diethoxylate
(NP1EO and NP2EO)
O
O
OH
n=1-2
Nonylphenoxy mono- and diacetic acid
(NP1EC and NP2EC)
H19C9
OH
Nonylphenol (NP)
Figure 2.4. Generalized scheme for biodegradation of NPnEO under aerobic and
anaerobic conditions
23
H19C9
H19C9
O
O
Nonylphenol diethoxylate
(NP2EO)
O
Nonylphenol monoethoxylate
(NP1EO)
+
+
3
H19C9
O
O
COOH
n
Di-carboxylic poly(ethylene glycols)
H19C9
OH
O
O
O
HOOC
acetaldehyde
1
OH
O
OH
n
O
O
COOH
O
n
Nonylphenol ethoxylates
H19C9
H19C9
O
O
O
+
O
O
CH2
HOOC
x
O
COOH
n
2
COOH
Nonylphenol diethoxycarboxylate
(NP2EC)
OH
Nonylphenol diethoxylate
(NP2EO)
HO
O
4
Carboxylic poly(ethylene glycols)
O
COOH
Diethoxycarboxylate with an
alkyl chain of varying lengths
Figure 2.5. Suggested pathways for aerobic degradation of nonylphenol
ethoxylates.1: (John and Whilte, 1998; 2: (Nguyen and Sigollot, 1997); 3: (Franska
et al., 2003); 4: (Jonkers et al., 2001), (adapted from Soares et al, 2008).
Two additional pathways revealed using tert-octylphenol polyethoxylate (OPEO) as
the substrate have been demonstrated. Whether these pathways are also applicable
to NPEO has yet to be established, but given the overall similarity between NPEO
and OPEO, they may have relevance in realistic systems. These two schemes differ
from the previous two in that instead of the ethoxylate side-chain being sequentially
24
removed one monomer at a time, most of the side-chain is removed at once. In the
first, the side-chain is subjected to an internal ether cleavage, yielding OPEO whose
side-chain length is reduced by several units and a carboxylated polyethylene glycol
as end products (Figure 2.5, scheme 2) (Nguyen and Sigoillot 1996). In the other
pathway, the terminal alcohol is carboxylated prior to internal ester cleavage, again
yielding short chain-length OPEO as an end product but along with dicarboxylated
polyethylene glycol as a byproduct (Figure 2.5, scheme 3) (Franska et al. 2003).
Considering the variety of the existing evidence on the pathways employed in initial
attack on the ethoxylate side-chain, it is clear that many pathways likely
simultaneously exist in a network of metabolic reactions (Giger et al. 2009). A
proposal for how these various pathways are integrated is shown in Figure 2.6 (Giger
et al. 2009). The degradation of the polyethoxylate chain occurs quickly under both
aerobic and anaerobic conditions and appears to be performed simultaneously
several different phylogenetic groups. Factors demonstrated to have some influence
over which pathway for ethoxylate chain removal include length of the ethoxylate
side-chain, hydraulic retention time, dissolved oxygen, and solids retention time
(Giger et al. 2009). All of these initial pathways converge on a common point; once
the bulk of the side chain is removed, yielding NP, NP1EO, NP2EO, NP1EC, or
NP2EC, biodegradation slows down greatly (Figure 2.6). Because these compounds
degrade slowly, they tend to accumulate in treatment effluents and in sludges (Bergé
et al. 2012).
Under anaerobic conditions, NP is considered to be the end-product metabolite (Ahel
et al. 1994, Montgomery-Brown and Reinhard 2003), although the mechanism of
conversion of NP1EO and NP1EC to NP is still poorly characterized (Kohler et al.
2008). Under aerobic conditions, NP1EO, NP2EO, NPEC, and NP2EC are
considered by most to be dead-end metabolites, although it has also been reported
that under microaerobic conditions, NPnEC can be metabolized further to
metabolites with oxidized alkyl and ethoxy side chains (CAPnEC) (Di Corcia et al.
25
1998, Di Corcia et al. 2000). That NP1EO, NP2EO, NP1EC, or NP2EC are dead-end
or at least highly recalcitrant under aerobic conditions is substantiated by their
frequent detection in sewage treatment effluents (Giger et al. 1984, Ahel et al. 1994,
Field and Reed 1996, Di Corcia et al. 2000), sludges (Giger et al. 1984, Field and
Reed 1996, Jonkers et al. 2001), rivers and estuaries (Kvestak and Ahel 1994, Dachs
et al. 1999, Potter et al. 1999, Ferguson et al. 2001, Jonkers et al. 2001), and
sediments (Lee Ferguson and Brownawell 2003). While a slow degradation of NP
takes place under aerobic conditions, virtually no further biodegradation occurs
under anaerobic conditions (Giger et al. 2009).
Figure 2.6. Integrated scheme for nonylphenol polyethoxylate metabolism. Adapted
from (Giger et al. 2009).
26
2.3.2. Biodegradation of Nonylphenol
In the environment and in wastewater treatment systems, NP degradation is strongly
limited by lack of bioavailabilty (due to its high Kow and Koc) (Bosma et al. 1997)
and by lack of oxygen (Topp and Starratt 2000, Hesselsøe et al. 2001). Topp et al.
(2000) demonstrated that NP spiked into well-aerated agricultural soils was fully
mineralized without any lag phase with a half-life of less than a day; however, when
spiked along with sewage solids, the half-life dramatically increased to
approximately 10 days, a finding which the authors took to be indicative of the fate
of NP when oxygen is limited, although this might also have represented the effect
of sorption reducing bioavailability. Similar potentials for environmental systems to
degrade NP were also demonstrated by Soares et al., in which several NP degrading
strains were enriched from natural soils (Soares et al. 2003).
The behavior of NP in sewage treatment systems has been investigated in several
studies. Ahel et al. demonstrated that over 60% of the NP-type compounds entering
a municipal treatment system was released into the environment as a mixture of NP
and short chain NPEOs and NPECs; the vast majority of the released NP was sorbed
to solid materials (Ahel et al. 1994). These same trends were also demonstrated by
Scrimshaw and Lester (Scrimshaw and Lester 2002) and Shao, et al. (2003) (Shao et
al. 2003).
Given the recalcitrance and narrow phylogenetic group of organisms capable of
degradation, it seems that NP presents a unique challenge to bacteria. While phenol
degradation is quite diverse and common amongst bacteria, NP requires specialized
metabolic approach, likely due to the presence of the bulky aliphatic substitution in
the para-position. It is also important to note that NP is actually a complex mixture
of isomers with different configurations of the nonyl-substitution (Ieda et al. 2005).
One of the first hints regarding how bacteria metabolize nonylphenol came from a
study of octylphenol; when OP was incubated with Sphingomonas sp. TTNP3, 2,4,4-
27
trimethyl-2-pentanol accumulated in the medium. This compound represented the
alcohol of the octyl group and was a dead-end product. There was no sign of the fate
of the aromatic ring, indicating that it served as a growth source (Tanghe et al.
2000). This same strain was later shown to be capable of growth on NP by a similar
mechanism. Since then, other Sphingomonas strains able to use NP as sole carbon
and energy sources have been isolated; S. cloaceae (Fujii et al. 2001), S. amiense
(Ushiba et al. 2003), and S. xenophaga Bayram (Gabriel et al. 2005).
The
metabolism of NP relies upon an initial mono-oxygenation which proceeds by “ipsohydroxylation”, or a hydroxylation at the 4-position on the phenolic ring (Figure
2.7). (Gabriel et al. 2007).
NP is a complex mix of isomers; some isomers have been shown to be usable as
growth sources while others seem to be highly recalcitrant (Gabriel et al. 2005,
Kolvenbach et al. 2007, Gabriel et al. 2012, Kolvenbach et al. 2014). While it was
originally suspected that the recalcitrance of NP is due to the high proportion of
isomers fully saturated with carbons at the alpha position on the nonyl-sidechain, socalled quaternary isomers blocking oxidative processes (Wheeler et al. 1997), it was
later shown that such isomers are actually preferentially degraded by a unique
metabolic process (Figure 2.6).
Degradation proceeds by action of an ipso-hydroxylation, leading to ejection of the
side-chain as a carbonium ion which reacts abiotically with water to generate the
corresponding nonanol (Figure 2.7) (Gabriel et al. 2005, Kohler et al. 2008, Gabriel
et.al. 2012).The aromatic ring remains as hydroquinone, which is seldom directly
observed due to its ready biodegradability (Giger et al. 2009).
28
Figure 2.7. Metabolism of α-quaternary nonylphenol isomers by Sphingobium
xenophagum Bayram and by Sphingomonas sp. TTNP3. From Kolvenbach et al.
(2014).
On the other hand, NP isomers without full carbon substitution at the alpha sidechain
carbon, so-called alpha-tertiary, cannot serve as growth substrates for organisms
isolated to date. In the case of these isomers, ipso-hydroxylation still takes place, but
the resulting hydroxylated metabolite undergoes an NIH shift of the aliphatic
sidechain, yielding a dead-end metabolite without any subsequent assimilative
metabolism (Figure 2.8) (Kohler et al. 2008, Gabriel et al. 2012). This metabolite, a
substituted hydroquinone, has been shown to undergo abiotic oxidation to a toxic
alkylquinone, a process which is suspected to contribute to the overall toxicity of NP
in natural systems (Gabriel et al. 2012).
29
Figure 2.8. Metabolism of α-tertiary nonylphenol isomers by Sphingobium
xenophagum Bayram (from Kolvenbach et al., 2014).
2.3.3. Metabolism of Hydroquinone
Hydroquinone (HQ), while not a typical metabolite in aromatic metabolism, has also
been observed as an intermediate in the metabolism of 4-ethylphenol by
Pseudomonas putida (Darby et al. 1987), 4-nitrophenol by Moraxella sp. (Spain and
Gibson 1991), γ-chlorocyclohexane by Sphingomonas paucimobilis (Miyauchi et al.
1999), 4-Hydroxyacetophenone by Pseudomonas fluorescens ACB (Moonen et al.
2008), 4-aminophenol by Burkholderia sp. strain AK-5 (Takenaka et al. 2003) and
hexachlorocyclohexane by a Sphingomonas species (Nagata et al. 1999).
Hydroquinones have been observed to be metabolized by two general mechanisms
which have little similarity to one another. By the first, HQ is cleaved by a direct
oxidative ring attack by an Fe(II)-containing enzyme to yield 4-hydroxymuconic
semialdehyde (Chauhan et al. 2000) (Moonen et al. 2008) while by the second
pathway, HQ is first hydroxylated to 1,2,4-trihydroxybenzene (hydroxyquinol)
which is then cleaved to produce maleylacetic acid by an Fe(III)-containing
dioxygenase (Jain et al. 1994). Hydroxyquinol is often employed in the metabolism
30
of chlorophenols (Olaniran and Igbinosa 2011) and was recently shown to be utilized
in the metabolism of diclofenac (Wojcieszyńska et al. 2014).
Enzymes involved in the first pathway, the hydroquinone dioxygenases (HQDO), are
further divided into two groups, type I and type II. Type I are monomeric proteins
that are closely related to well-known catechol extra-diol dioxygenase enzymes
which are very often involved in bacterial aromatic degradation pathways (Eltis and
Bolin 1996). Fewer examples of type II dioxygenases, which are tetrameric proteins,
have been described to date. The NP degraders Sphingomonas sp. TTNP3 and
Sphingomonas xenophagum Bayram, which metabolize quaternary NP isomers to
HQ via ipso-hydroxylation, have been shown to utilize a type II HQDO system. The
further downstream enzymes involved in metabolizing the ring-cleavage product, 2hydroxymuconic semi-aldehyde, have also been described (Figure 2.9). The muconic
semialdehyde is metabolized to the central compound 3-oxoadipate via maleylacetic
acid (Kolvenbach and Corvini 2012).
31
Figure 2.9. Metabolism of hydroquinone by Sphingomonas sp. TTNP3 (Kolvenbach
and Corvini 2012).
2.4. Molecular Approaches to Phylogenetic Identification of Microorganisms
While there is no question that the culturing of environmental microorganisms in the
laboratory has formed the foundation of modern microbiology, this approach faces
severe limitations in regards to understanding environmental microbiological
communities and processes (Hugenholtz et al. 1998). “Culture-independent” is a
generic term used to describe methodologies that avoid the use of cultures at all. At
the simplest level, this would include microscopy, a centuries-old approach. Simple
microscopy is very limited when it comes to the bacteria and archaea however,
because they usually lack any notable physical features. This is overcome by the use
32
of differential stains, such as the Gram-stain for example, which can differentiate
between certain phylogenetic groups.
It has been estimated with the aid of research-based studies that less than 1% of
microbial diversity in the environment is actually culturable (Torsvik et al. 1998).
Thus the need for further refinement of culture-independent approaches is
unquestionable (Hugenholtz et al. 1998). The foundation of the culture-independent
methods lies not in the need for new ways to understand natural microbial
communities but rather in the desire for tools by which to understand the true
phylogenetic relationships between organisms. Until the advent of modern molecular
tools, other characteristics such as physical structures, metabolic potentials,
reproductive strategies and behaviors have been used to characterize phylogenetic
relationships (Barrow and Feltham 2004). These approaches are quite lacking when
it comes to microbiology however, as bacteria and archaea lack sufficiently
diagnostic physical features and often show overlapping metabolic potentials across
widely different phylogenetic groupings. Before the molecular era, microbial species
definitions were very speculative (Olsen et al. 1986). It was proposed as early as the
1960s that information-rich biomolecules could be used as a way to identify
microorganisms and infer evolutionary relationships; one of the first proposals was
to use the degeneracy of the genetic code as a phylogenetic tool, which has proven
somewhat useful but far from adequate (Zuckerkandl and Pauling 1965). Sequences
of conserved proteins were also successfully used, although this is only useful within
small phylogenetic groups since there are no proteins that are conserved universally
throughout all life-forms (Angerer et al. 1985).
The large ribosomal subunit (16S rRNA or 23S rRNA) ultimately proved to be the
best, if not the only, molecule appropriate for detailed phylogenetic analyses. The
large ribosomal subunit is found in all life-forms, is not subject to lateral gene
transfer, and shows the appropriate balance between evolutionary sequence
alterations and functional sequence conservation (Olsen et al. 1986). The earliest
33
sequence analyses of the 16S rRNA gene drastically changed the way that all lifeforms are believed to be related to one another. Bonen et al. established the
symbiotic origin of eukaryotic organelles (Bonen and Doolittle 1976) and the team
led by Carl Woese, by sequencing 170 laboratory grown organisms, first revealed
the existence of three basic lineages of life on the planet (bacteria, archaea, and
eukarya) (Fox et al. 1980). The 16S rRNA gene sequences obtained in these and the
thousands of subsequent analyses of cultured organisms have been digitized and
placed in a number of public databases such as GenBank (Benson et al. 2009),
Ribosomal Database Project (Cole et al. 2009), Silva (Pruesse et al. 2007), and ARB
(Ludwig et al. 2004); thus this laboratory-based work provides the foundation for
modern approaches to environmental genetic analyses.
2.4.1. Characterization of Environmental Microbial Communities by Molecular
Analyses
The identification of microbial population and their diversity in lab-scale reactor
systems have become important to understand the real microbial environment of
wastewater treatment systems. With the help of obtained results from these systems,
the performance and efficiency of wastewater treatment systems could be improved
(Sanz and Kochling, 2007). In order to study microbial community structure,
molecular methods have been developed and successfully applied on sludge samples
in last decades. Conventional culture-dependent methods are inadequate for this kind
of studies due to limited information about real growth conditions of most of
microorganisms.
2.4.1.1. Fluorescence In Situ Hybridization (FISH)
One of the ideal goals of microbiology is to look at a sample using microscopic
methods and be able to identify and quantify the organisms therein. Basic light
microscopy, simply looking at an amplified image, reveals little information beyond
34
shapes of the organisms or in rare cases the presence of specialized structures or
inclusion bodies. Very little relevant phylogenetic information can be obtained by
such an approach. Since the origin of microbiology, differential staining techniques
have been applied in order to gain more information about the visualized
microorganisms. The gram stain technique for example preferentially stains bacteria
with thick cell walls, and the acid-fast stain can detect the presence of Mycobacteria
(Barrow and Feltham 2004). These staining approaches are very useful in some
situations but are far from adequate in terms of being able to characterize many
different phylogenetic groups. Microscopic staining techniques generally revolve
around the idea of a dye or washing procedure that can ultimately detect certain
diagnostic biological molecules, such as membrane components or inclusion bodies.
The molecular biology revolution of the modern era opened up many new
possibilities in terms of specific biological molecules. As early as the 1970’s
techniques for detecting mRNA were developed and applied in medical research
(Harrison et al. 1973, Harrison et al. 1974). These techniques generally revolved
around a radioactively labelled single-stranded DNA probe complementary to the
mRNA molecule in question; the probe would bind strongly to the target and thus
the location and amount of mRNA could be estimated (Angerer et al. 1985). The
realization that 16S rRNA was rich in phylogenetic information and a growing
amount of known 16S sequences (Olsen et al. 1986) made it possible to develop
DNA probes specific to certain phylogenetic groups (Pace et al. 1986), allowing in
situ visualization of target cells when the DNA probe was labelled with a fluorescent
dye. This approach is called “fluorescent in situ hybridization” (FISH).
2.4.1.1.1. Theory and Methodology of FISH
The most common application of FISH is the 16S rRNA molecule (Wagner and
Haider 2012). Note that in phylogenetic identification and in many alternative
environmental microbiological analyses, the DNA gene that codes for the 16S
35
molecule is used instead. The 16S rRNA is the ideal target for this microscopic
approach because of its intracellular abundance. The ribosome is one of the most
numerous structures in a cell and each one contains a 16S rRNA molecule (Barrow
and Feltham 2004). While there are at most a handful of copies of the 16S rRNA
gene in a cellular chromosome, there are thousands of ribosomes. The high number
of ribosomes means more targets for fluorescently labelled probes and thus stronger
signals, making labelled cells easier to detect (Moter and Göbel 2000).
The core method must be preceded with probe design. A complicated action by
itself, this relies upon preexisting molecular sequences, design of probes with the
desired level of specificity, then confirmation (Amann et al. 1990). Probes are
typically 15-30 nucleotides long (Amann et al. 1995). As the probe design and
validation are laborious, research will ideally use probes that have been previously
validated. Currently, probes are designed with the desired level of specificity using
computerized tools such as the ARB software package (Ludwig et al. 2004). GC
content must be carefully controlled because this has a high effect on the
hybridization strength; content is typically from 50 to 70%. Software tools also
exclude probes which may exhibit formation of secondary structures. After
computerize design, probes must be tested on a range of laboratory organisms in
order to confirm their accuracy.
The laboratory methodology generally begins with fixation of the sample containing
the target cells. Fixation stabilizes the macromolecules, thus preserving overall
cellular structure and importantly preserving the 16S rRNA. Fixation is performed
with the classic preservative formaldehyde or paraformaldehyde. Fixation also
permeabilizes cell membranes, necessary in order for the probes to enter the cells
and reach their targets (Amann et al. 2001).
FISH is not limited however to targeting the 16S rRNA molecule; it can be used to
target genes or mRNA molecules for gene expression studies, although since they
36
are less abundant, more advanced signal generation technologies are required, such
as signal-amplifying enzyme labels (CARD-FISH) (Schönhuber et al. 1997,
Pernthaler et al. 2001, Pernthaler et al. 2002, Wagner and Haider 2012).
2.4.1.1.2. Advantages of FISH
The main alternative methods for phylogenetic characterization of microbial
communities mostly revolve around the amplification of the 16S rRNA gene by
polymerase chain reaction. These includes creation of clone libraries, DGGE/TGGE,
t-RFLP and quantitative PCR. These techniques all have their relative advantages
and disadvantages but all suffer from a major disadvantage relative to in situ
approaches as is discussed below.
Creation of clone libraries involves the amplification of the 16S gene using primers
with a desired level of specificity, often designed to amplify all bacterial 16S genes,
followed by ligation of the product into a plasmid and transformation into a
laboratory bacterium, usually E. coli, for maintenance (Nocker et al. 2007). This type
of approach yields a desired number of E. coli strains which each has a unique PCR
amplicon on a plasmid which can then be subjected to further analysis. This will
produce a large amount of data when the inserts are sequenced, a profile of the types
of organisms present in the sample and of the relative abundances of those
organisms. Relative to other PCR-based strategies, construction of clone libraries is
very labor and resource intensive.
DGGE and TGGE (denaturing gradient gel electrophoresis and temperature gradient
gel electrophoresis respectively) approach the problem of the mixed PCR product in
a different and more streamlined manner (Muyzer 1999). They separate the different
sequences on a gel that contains either a DNA denaturing or temperature gradient,
with the idea that the double-stranded nature of different sequences will break down
at different spots on the gel and the DNA bands will become immobilized at
37
different places. This yields a different “fingerprint” of bands for different samples
(Muyzer 1999). This technique has an advantage in the different bands can be cut out
of the gel and analyzed further (Ercolini 2004).
t-RFLP (terminal restriction fragment length polymorphism) is a high throughput
technique that separates the mixed 16S pool based on restriction of the sample under
the premise that different sequences will cut at different spots (Marsh 1999). One
end (the “terminal” end) is visualized by means of incorporation of a fluorescent dye
in one of the PCR primers; different sequences will thus yield fluorescently-labelled
DNA fragments of different lengths which are then separated by capillary
electrophoresis. The main disadvantage of t-RFLP is that samples only yield a
fingerprint; the different fragments cannot be sequenced, thus only changes between
samples can be characterized (Schütte et al. 2008).
Each of these popular techniques starts with a PCR amplification. This is the main
spot where they differ from in situ approaches such as FISH. The overall PCR
process introduces bias at many stages of the procedure (V. Wintzingerode et al.
1997). Firstly, different types of microbes may react differently to the initial DNA
extraction procedures. For example, an extraction process that may be suitable for
gram positive bacteria, which have strong cells walls requiring more harsh
conditions, may damage the DNA in gram-negative bacteria (Kirk et al. 2004). In
the PCR process, primers hybridize to sections of extracted DNA and repeated
polymerization of the target is catalyzed. It has been demonstrated that this process
can be highly biased, with some sequences being amplified much more efficiently
than others. This is particularly problematic in the case of the 16S gene, the product
of which is specifically designed to create secondary structures, which can interfere
with amplification. The PCR process also produces artifacts, or unintended products
which produce false data during analysis (Suzuki and Giovannoni 1996). Further
complicating the issue, the polymerase usually used in PCR (Taq polymerase) has a
considerable error rate; after the typical thirty cycles of amplification used in PCR,
38
many deletions and point mutations can accumulate, up to five per 16S amplicon
(Gutell et al. 1994). Further, PCR can amplify DNA that was not present in living
cells, thus making the false suggestion that non-living DNA represents actual viable
members of the ecosystem. This is not as much of a problem for detection of rRNA
by FISH because RNA degrades relatively much faster when outside of a living cell.
2.4.1.1.3. Limitations of FISH
In PCR-based approaches, all 16S genes can be amplified, including those
previously unknown, allowing for later characterization of entirely novel sequences.
However, FISH depends completely on the accurate design of DNA probes that are
complementary to 16S rRNA molecules. This means that a study can only search for
and visualize known phylogenetic groups. However, currently there are thousands
of 16S sequences available in online databases (Ludwig et al. 2004, Pruesse et al.
2007, Benson et al. 2009, Cole et al. 2009) and a large number of rRNA-targeted
FISH probes have been designed and validated; these probes can be found by general
literature searches or at public databases such as probeBase (Loy et al. 2003, Loy et
al. 2007).
2.4.1.1.4. Application of FISH to Studying the Impact of NP and NPnEO on
Microbial Communities
2.4.1.1.4.1. Nonylphenol polyethoxylates
Lozada et al. studied the activity bacterial communities of activated sludge reactors
fed with nonylphenol polyethoxylates (NP10EO) by FISH (Lozada et al. 2004).
Overall reactor activity, as judged by carbon removal efficiency, was unaffected by
the addition of 60 ppm NP10EO. The reactors were subjected to FISH analysis with
several Proteobacterial probes in addition to probes for Acinetobacter and low GC
gram positives. The FISH community analysis results showed that the
39
Betaproteobacteria were enriched from 5% up to 33% of the total microbial
population. Since this group was using a long-chain NPEO (NP10EO), it was
unclear whether the shift was caused by the polyethoxylate or one of the downstream
metabolites.
Di Gioia et al. utilized one of the more powerful aspects of FISH when they
demonstrated the development of a three species aggregation in a culture enriched
with NP6EO-NP9EO (Di Gioia et al. 2004).
The aggregate was composed of
bacteria of the Bacillus, Acinetobacter, and Stenothrophomonas. Bacillus appeared
to be responsible for allowing aggregation while the other two strains were
responsible for a cooperative metabolism of the NPEO. Whether there was a
metabolite produced was not clear, although it is likely that they did not metabolize
the core NP since none of these types of organisms have ever been demonstrated to
degrade NP in other studies.
Salvadori et al. isolated a novel low ethoxylated NPnEO degrader of the
Stenotrophomonas genus, which is within the Gammaproteobacteria (Salvadori et
al. 2006). In order to investigate this organism in situ, they developed novel FISH
probes. The probes were used to confirm that the organism is found in aerobic
sewage treatment systems, although no attempts at a quantitative analysis were
made.
Bertin et al. used FISH to characterize the biofilm that formed in backed-bed
reactors acclimated to 30-90 ppm NPnEO (Bertin et al. 2007). The reactors showed
high parent compound removal rates of up to 99% over nine days. NP did
accumulate in the system during biodegradation however, demonstrating the
recalcitrance of the core molecule. FISH analysis revealed that the biofilms were
dominated by Gammaproteobacteria.
40
2.4.1.1.4.2. Nonylphenol
Symsaris et al. examined the effect of NP on methanogenesis by anaerobic sludge
reactors (Symsaris et al. 2015). They found an IC50 of 363 ppm for inhibition of
methanogenesis by NP. Methanogen populations were targeted for FISH analysis,
with probes for four major phylogenetic groups. The image analysis and
quantification methods for the FISH methodology were not made clear, but NP
appeared to alter the methanogen community although without a clear functional
significance; in a NP-inoculated reactor, there were no clear changes at 100 mg/L
NP, but when NP was increased to 300 mg/L Methanosaeta and Methanobacteriales
decreased while Methanomicroiales and Methanococcales species increased.
Xie et al. used FISH as an overall quantitative tool to examine the abundance of
bacteria and archaea in soils contaminated with 50 mg/kg NP (Xie et al. 2013).
Their results suggested 67% and 49% declines in bacterial and archaeal populations
respectively.
Di Giola et al. used phylogenetic FISH probing to examine a microbial consortium
able to degrade high concentrations of NP under aerobic conditions (Di Gioia et al.
2008).
The consortium was generated by enrichment from a sewage treatment
facility. While no individual isolates could be obtained, the phylogenetic probing
allowed for a characterization of the consortium; it contained a high occurrence of
Proteobacteria, especially the Gammaproteobacteria, which appeared to make up
the majority.
A study by Soares et al. using packed bed reactors constantly fed a saturated solution
of NP revealed a strong enrichment of Betaproteobacteria, in some cases apparently
making up to 60% of the total bacterial population (Soares et al. 2006). Their
analysis was performed using FISH with probes for major Proteobacterial groups
along with certain gram positives and a probe specific to Pseudomonas.
41
In a study of lake sediment microcosms utilizing FISH analysis with Proteobacterial
and gram positive probes revealed that the presence of part per billion concentrations
of NP led to strong enrichments of Beta- and Gammaproteobacteria, in addition to
strong enrichment of gram positive bacteria when the NP concentration was
increased to 1 ppm (Jontofsohn et al. 2002).
Wang et al. studied the impact of 150 ppm NP addition to river microcosms by
constructing clone libraries for general phylogenetic analysis and by performing tRFLP on the conserved amoA gene of ammonia oxidizers (Wang et al. 2014). As
with other studies performed by this research group however, it is not clear if they
utilized linear or branched NP in their study. They demonstrated that
Alphaproteobacteria, Gammaproteobacteria and Bacteroidetes were enriched by the
NP addition. Ammonia oxidizing guilds were also affected by the NP addition,
although the nature of t-RFLP prohibited any detailed descriptions of how the
community was affected specifically. Nevertheless, this study seemed to show that
metabolic guilds associated with important ecological functions can be affected by
the presence of this micropollutant. These findings were later confirmed by highthroughput sequencing analysis (Wang et al. 2015).
In a related study which did not apply FISH, Zhang et al. applied semi-quantitative
PCR to investigate a hypothesis that branched NP might be degraded in situ by
classical meta-cleavage (catechol 2,3-dioxygenase) pathways (Zhang et al. 2008).
They compared copy numbers of catechol 2,3-dioxygenase genes to 1,2dioxygenase, alkane catabolic genes, and 16S rDNA, showing that the 2,3
dioxgenase gene abundance increased during incubation of 25 ppm NP in aquatic
microcosms. t-RFLP analysis results suggested an increased abundance of
Betaproteobacterial species, although this result was speculative given the inexact
nature of t-RFLP. These results generally run counter to the accepted pathway at this
point, the ipso-hydroxylation pathway, which does not involve a catecholic
42
metabolite. However, the lack of catecholic pathways for NP degradation in isolates
to date does not preclude its occurrence in natural systems.
Additionally, Lozada et al. examined the response activated sludge reactor microbial
communities to enrichment with NPEOs using DGGE and 16S clone libraries
(Lozada et al. 2006). These approaches found enrichments of Gammaproteobacteria,
Acidobacteria
and
Alphaproteobacteria
(including
Sphingomonas)
and
Betaproteobacteria (Burkholderia).
Overall, studies of both NP and NPnEO have found enrichments of
Gammaproteobacteria in almost every case.
Enrichments of Alpha- and
Betaproteobacteria also have been reported. Note that the only isolated bacteria
capable
of
metabolizing
NP
have
been
Sphingomonads,
which
are
Alphaproteobacteria.
2.4.1.2. Quantitative Polymerase Chain Reaction (qPCR)
Traditional polymerase chain reaction (PCR) involves a reaction in which two short
single stranded DNA oligomers are incubated with a DNA mixture of interest with
Taq DNA polymerase and a nucleotide mixture. This mixture is thermocycled many
times, typically 30-40, in such a way that the target DNA melts, DNA primers
hybridize to exposed template DNA, and the polymerase creates a copy of the
template. This process is exponential. By the end of the cycling process, nucleotides
are exhausted and a large amount of new product DNA is created regardless of how
much template was present in the mixture. Thus, classical PCR is inherently nonquantitative. This short-coming was overcome in 1992 with the development of
“real-time” PCR or “quantitative” PCR (qPCR) (Higuchi et al. 1992), in which the
amount of product is quantified at every step using a fluorescent dye, usually the dye
SYBR Green, which is active only in the presence of double stranded DNA. The
43
fluorescent dye is an area of constant improvement; the dye EvaGreen has been
shown to be superior and is gaining in popularity (Eischeid 2011).
The intensity of fluorescence is proportional to the amount of DNA that has been
produced in the thermocycled reaction. In the early cycles, the signal is very low and
does not change to a detectable degree. However, as amplification proceeds, in an
exponential manner, a signal is eventually detected. The earlier the signal is detected,
the more abundant the template is. The cycle in which the threshold of detection
breached is called the Ct (for threshold cycle). In order to achieve accurate
quantification of the template, the Ct of an unknown sample is simply compared to a
standard curve consisting of the Ct values of a standard curve of reactions containing
known values of template (Nolan et al. 2006).
qPCR offers unique advantages for the phylogenetic analyses of microbial
communities. Advantages of qPCR include the requirement of only a small amount
of template, high throughput, low cost, and high sensitivity (Bacchetti De Gregoris et
al. 2011). It has been used for this purpose for over a decade at this point, beginning
with the work of Suzuki et al. in 2000 (Suzuki et al. 2000), in which certain specific
phylogenetic groups were analyzed in a natural population. The main obstacle for the
growth of qPCR for microbial community analysis has been the development and
confirmation of probes that can detect broader phylogenetic levels, such as the
phylum level (Fierer et al. 2007). Fierer et al. developed broader phylogenetic primer
sets targeting six bacterial and two fungal phyla which could be used to gain a more
comprehensive view of soil microbial communities (Fierer et al. 2005). Since then,
qPCR has been successfully applied in various studies for microbial community
analysis (Philippot et al. 2009, Wessén et al. 2010, Philippot et al. 2011). The
primers have been the subject of constant testing and refinement (Morales and
Holben 2009, Töwe et al. 2010). Primers for many more phyla have been developed,
including sets for archaeal phyla (Wessén et al. 2010)
44
While FISH has become quite popular for studying engineered systems such as
bioreactors, qPCR has mainly been utilized in microbial ecology, with many
investigations of agricultural soils (van Diepeningen et al. 2006, Birkhofer et al.
2008, Simmons and Coleman 2008), natural soils (Yergeau et al. 2010, Philippot et
al. 2011) and rumen systems (Tajima et al. 2001, Deng et al. 2008). qPCR has been
used to analyze the responses of microbial communities to pollutants such as heavy
metals (Feris et al. 2003) (Wakelin et al. 2014), petroleum products (Feris et al.
2003) (Lladó et al. 2013), ethanol (Feris et al. 2003), trichloroethylene (Sleep et al.
2006), and aromatics (Martínez-Pascual et al. 2015). However, this technique has
not yet been applied to the study of the impacts of NP on microbial systems.
45
46
CHAPTER 3
MATERIALS AND METHODS
3.1. Sludge Samples
Waste activated sludge (WAS) and anaerobically digested sludge (ADS) samples
used in setting up the lab-scale digesters were supplied from Ankara Central
Wastewater Treatment Plant. WAS samples were taken from the plant every week to
provide enough feed for the daily basis feeding regime of the digesters. The
treatment plant is the biggest treatment plant in Ankara and second largest in Turkey
with a current flow rate as 765,000 m3/day (ASKI, 2015).
ADS samples were taken from inside of a mesophilic anaerobic digester and WAS
samples from the return sludge line of a secondary sedimentation tank. To reach
desired solids concentrations, WAS and ADS samples were settled and supernatant
was discarded periodically.
3.2. Chemicals
The standard solutions containing 5 µg/mL for NP, NP1EO and NP2EO in acetone
(OEKANAL) were purchased from Fluka (Sigma Aldrich Co. LLC, USA) to
calibrate the Gas Chromatography/Mass Spectrometry (GC/MS). 4-n-nonylphenol
(10 ng/µL in Cyclohexane) was supplied by Dr. Ehrenstorfer (Augsburg, Germany)
and used as surrogate in extraction experiments. 4-Nonylphenol-di-ethoxylate (10
mg, 99% purity) was used as the parent compound to spike into anaerobic semicontinuous digesters and was provided by Dr. Ehrenstorfer (Augsburg, Germany).
47
Prior to GC/MS injection, for NP, NP1EO and NP2EO the extracts were derivatized
by using N,O -bis(trimethylsilyl)-trifluoroacetamide with 1% trimethylchlorosilane
(BSTFA+TMCS, 99:1, Sylon BFT) purchased from Supelco Analytical (Sigma
Aldrich Co. LLC, USA). Sodium sulfate (anhydrous) and fine powdered copper
were obtained from Merck (Darmstadt, Germany) to remove moisture and sulfur
during the solid phase extraction, respectively. Certified reversed C18 Sep-Pak
(tC18, 6cc, 500 mg) cartridges were supplied from Waters (Milford, MA, USA) to
adsorb analytes from liquid phases. All of the GC grade solvents, hexane, acetone
and methanol, were purchased from Merck (Darmstadt, Germany).
For the preparation of COD kits, potassium dichromate (K2Cr2O7), silver(II) sulfate
(Ag2SO4), mercury(II) sulfate (HgSO4) and sulfuric acid (H2SO4, 95-97%) were
supplied from Merck (Darmstadt, Germany). Potassium hydrogen phthalate (KHP)
was used as a standard for preparation of calibration curve in COD analysis and was
obtained from Merck (Darmstadt, Germany). Alconox (White Plains, NY, USA) was
used as detergent during the cleaning of all glassware.
3.3. Experimental Set-up
3.3.1. Set-up of Laboratory Scale Semi-Continuous Digesters
WAS taken from a return line of a secondary sedimentation tank was used as a feed
source and ADS obtained from mesophilic digester was used as inoculum (seed) for
the set-up of the laboratory scale semi-continuous digesters. WAS and ADS samples
were settled overnight and supernatant was discarded periodically to increase solids
concentration. To minimize microbial activity during settling, these samples were
kept at 4oC. Prior to digesters set-up, solids content analyses of concentrated WAS
and ADS samples were performed. TS, TSS, VS and VSS concentrations of WAS
and ADS samples are given in Table 3.1. The food to microorganism ratio (F/M)
was determined based on g-VS /g-VSS and set to 1 by mixing required volumes of
WAS (960 mL) and ADS (1040 mL) samples.
48
Table 3.1. Solids concentrations used in digester set-up
TS (mg/L)
VS (mg/L)
TSS (mg/L)
VSS (mg/L)
WAS
21600
17950
20175
17300
ADS
33240
17300
33173
16530
3.3.2. Operation of Laboratory Scale Semi-Continuous Digesters
Three sets of lab-scale semi-continuous anaerobic digesters were operated in 3 L
reactors with 2 L-working volume. Each digester set was run in duplicate, resulting
in six digesters in total, and was named as following: spiked digester (SD), abiotic
control digester (AD) and biotic control digester (BD) (Figure 3.1). These digesters
were filled with certain volumes of WAS and ADS as mentioned earlier to set F/M
ratio at 1. After filling, all digesters were purged with N2 for 10 min to remove
oxygen from the system and then sealed to provide required anaerobic environment.
Abiotic control digesters were autoclaved for 1 hour at 121oC to eliminate microbial
activity.
All digesters were connected to graduated gas collection cylinders (4 L) to monitor
gas production during operation (Figure 3.1). The cylinders were filled with brine
solution (10% NaCl, w/v and 2% H2SO4, v/v) to prevent solubilization of biogas.
The volume of produced biogas was monitored by measuring the displacement of
solution level in the graduated cylinders daily. The digesters were operated at 35oC
in a hot-room and were placed on magnetic stirrers to provide mixing of digester
contents. The sealing of all digesters was checked carefully to ensure that no air
intrusion into the digester systems was possible. After digesters were set up, the
initial TS, VS, TSS, VSS and COD values for each digester were measured and are
presented in Table 3.2.
49
Figure 3.1. Configuration of lab-scale anaerobic semi-continuous digester connected
to gas collection cylinder.
Table 3.2. Initial parameters following digester set-up
AD-1
AD-2
BD-1
BD-2
SD-1
SD-2
TS (mg/L)
VS (mg/L)
26090
15340
26210
15410
25650
15080
25980
15260
25840
15220
25545
15140
TSS (mg/L)
22880
23120
22320
22790
22460
22020
VSS (mg/L)
13590
13740
13140
13470
13280
12960
COD (mg/L)
25460
25780
24740
24960
24890
24320
A fixed volume of sludge (133.4 mL) was removed from the sampling ports of the
digesters and the same amount of fresh activated sludge (feed) was added to the
digesters daily to maintain a solids retention time (SRT) of 15 days. For feeding of
abiotic digesters, same amount fresh activated sludge was autoclaved for 30 min at
121oC and then was added to digesters by a sterile syringe.
50
In order to remove background NP compounds present in activated sludge, and to
achieve steady state, the digesters were operated without any spike for 63 days. The
digesters were considered to be at steady state when the variation in concentrations
of MLSS, MLVSS and NP compounds was less than 10%. Then, 3000 µg/L NP2EO
(in 3 mL acetone) was dosed into the spiked digesters at 63rd day (Figure 3.2). A 3
mL acetone without NP2EO was added to biotic control digesters. Then, parent
compound NP2EO and its degradation products NP1EO and NP in all digesters were
monitored and quantified by GC/MS during operation time. After 147 days (in total),
digesters were terminated.
Figure 3.2. Details of digester operation for spike and microbial activity
Table 3.3 summarizes the measured parameters and their measurement frequencies
over 147 days. TS, TSS, VS and VSS, COD, pH, gas production and composition
were monitored during the operation period of the digesters. All analyses were
performed with the 133.4 mL sludge sample withdrawn from each digester daily
51
according to the schedule given in Table 3.3. After NP2EO spiking at 63rd day,
sampling frequency of NP compounds was increased to every other day for 20 days
(63rd - 83rd) to observe possible biodegradation of NP2EO and changes in
concentration of NP2EO, NP1EO and NP more closely. After the relatively rapid
biodegradation of NP2EO within this 20 days, no drastic change was observed in the
concentration of NP compounds, so sampling frequency was decreased. Total gas
production and composition were determined every other day throughout 147 days.
TS, VS, TSS, VSS and COD analyses were conducted once in three days between
the 63rd and 99th days. Then, these analyses were carried out weekly up to the
digester-termination time. pH measurements were also conducted once in three days
for 36 days (63rd - 99th) following the spike. Then, sampling frequency was widened
to once in five days for the remaining 50 days (99th - 147th).
52
Table 3.3. Parameters and measurement frequency at different time intervals of
operation period.
Parameter
Operation period (d)
Measurement
1st - 38th
Frequency
in three days
Once
38th - 63rd
Once in five days
63rd - 83rd
Every other day
Nonylphenol compounds
(NP2EO, NP1EO, NP)
th
83 - 99
th
99th - 147th
st
1 - 147
Total gas volume and composition
th
1st - 44th
th
COD
Every other day
Once in three days
Once in four days
63th - 99th
Once in three days
99th - 147th
Once a week
1st - 38th
Once in three days
38th - 63rd
Once in five days
63rd - 99th
Once in three days
99th - 147th
Once a week
1st - 38th
Once in three days
th
rd
Once in four days
63rd - 99th
Once in three days
99th 147th
Once in five days
38 - 63
pH
Once a week
rd
44 - 63
TS, VS, TSS, VSS
Once in four days
3.4. Analytical Methods
3.4.1. pH Measurement
The pH measurements were conducted according to Standard Method 4500H
(APHA, AWWA, WEF, 2005). The pH was measured by a Oakton Waterproof 300
53
series (pH/mV/oC) pH meter. The pH meter was calibrated routinely using these
solutions prior to pH measurements.
3.4.2. Solids Determination
Total solids (TS) and Volatile Solids (VS) of sludge samples withdrawn from the
digesters were determined according to Standard Methods 2540B and 2540E,
respectively. On the other hand, Standard Methods 2540D and 2540E were applied
for the determination of total suspended solids (TSS) and volatile suspended solids
(VSS), respectively (APHA, AWWA, WEF, 2005). TS, VS, TSS and VSS analyses
were carried out in duplicate during operation period and the values are represented
as average.
3.4.3. Chemical Oxygen Demand (COD)
To determine COD of the sludge samples, Hach Method 8000 named “Reactor
Digestion Method” approved by United States Environmental Protection Agency
(USEPA) was followed. Instead of using ready-to-use Hach COD kits (0-1500 mg/L
range), the COD kits were prepared in the laboratory based on the Hach Water
Analysis Handbook (2012, 6th edition). Following the preparation of the COD kits,
they were calibrated by using potassium hydrogen phthalate (KHP) as a standard. In
the Hach handbook, KHP is chosen as a standard assuming its theoretical COD value
is 1.175 mg O2/mg KHP (on a weight basis). In the light of this information, a
calibration curve was prepared by using KHP at 0 – 1000 mg/L range.
The sludge samples obtained from the digesters were mixed homogenously and each
sample was diluted 1/50 using dH2O prior to all COD analyses. After 2 h-digestion
and cooling, COD readings were conducted using Hach DR 2400 spectrophotometer
at 620 nm wavelength. Triplicate analyses were carried out for each sludge sample
and then averaged value was reported as COD.
54
3.4.4. Total Gas Production
The digesters were connected to 4 L gas collection cylinders to monitor daily gas
production. The cylinders were graduated by sticking a strip indicating volume
intervals based on milliliter. The brine solution is used instead of water to prevent
solubilization of produced gases into water so gas production can be determined
without any loss of produced gas. Total gas production was monitored every day by
reading the level of brine solution displacement which was pushed down by the
produced gas.
3.4.5. Gas composition
The gas composition of the semi-continuous anaerobic digesters was determined
with a Gas Chromatograph (GC) equipped with Thermal Conductivity Detector
(TCD) (Agilent Technologies 6890N). Nitrogen (N2), methane (CH4) and carbon
dioxide (CO2) composition of the produced gas in the digesters were monitored. The
column chosen for this analysis was a 30.0 m X 530 μm X 40.0 μm nominal HP-Plot
Q capillary column. The oven program was as follows: column temperature was held
at 45°C for a minute and then increased to 65°C at a rate of 10°C/min. Helium was
the carrier gas with a flow rate of 3 mL/min. The retention times of the N2, CH4 and
CO2 were 2.1 min, 2.3 min and 2.9 min, respectively. Replicate injections were
carried out for each digester and the averaged value was reported.
On each analysis day, the GC/TCD was calibrated by using two standard gas
mixtures. The gas composition of each standard gas mixture (SGM) was as follows:
65%, 25%, 10% for SGM-1 and 25%, 55%, 20% for SGM-2. These two points
analyzed every measurement day were used for concentration correction. Each gas
mixture was injected three times to the GC/TCD.
55
3.4.6. Identification and Quantification of Nonylphenol Compounds
3.4.6.1. The Principles of GC/MS
Gas chromatography (GC) coupled with Mass Spectrometry (MS) is widely used for
the determination of volatile and semi-volatile organic compounds in complex
mixtures. Following the injection of sample into GC/MS, the compounds in the
sample are separated based on their volatility in the GC and carried through the
column by a carrier gas. These compounds are bombarded in the ion source section
by electrons causing them to be fragmented as positively charged ions. The mass of
each ion divided by the charge (always +1) is the mass charge ratio (m/z) which
represents the molecular weight of the ion. During the movement of these ions in the
MS, they travel through a quadrupole which forms an electromagnetic field. In this
part, the ions are filtered based on their mass. The quadrupole scans the ions based
on their mass until the range of m/z is recovered. This produces a graph of signal
intensity (relative abundance) versus m/z ratios (essentially molecular weight).
These m/z values are used as a unique fingerprint for each compound and the m/z
values of unknown compounds can be searched and matched in the library of the
GC/MS software.
3.4.6.2. GC/MS method for determination of NP compounds
In order to identify and quantify NP compounds in sludge samples obtained from
semi-continuous anaerobic digesters, a GC/MS (7890A Agilent gas chromatograph
coupled to a 5975C Agilent mass spectrometer with Triple-Axis) instrument was
used. In light of methods used in literature for NP compounds, our research group
developed a reliable and rapid method for detection and quantification of NP
compounds after an extensive study conducted. The details of the method are given
in Table 3.4.
56
A 1 µl of the sample was injected in the splitless mode at 250oC into a
30m×0.25mm×0.25µm HP-5MS 5% phenyl methyl siloxane capillary column
(Agilent 19091S-43E). The GC/MS was operated in Selective Ion Mode (SIM). The
mass spectrometer was operated in the electron impact ionization (EI) mode and the
energy of the electrons was kept at 70 eV. The interface was kept at 280oC and the
MS ionization source and the MS quadrupole were kept at 230 and 1500C,
respectively. The oven program used in analyses is given in Table 3.4.
Table 3.4. GC/MS method for determination of NP, NP1EO and NP2EO
GC/MS Method
Column
HP-5MS 5% Phenyl Methyl Siloxane
Carrier Gas
Helium (1 mL/min)
Injection volume
1 µL
Injection Mode
Splitless
Injection Temperature
250°C
MS Interface Temperature
280°C
MS Source Temperature
230°C
MS Quadrupole Temperature
150°C
MS Mode
SIM
100oC, hold for 5 min,
25oC/min to 160oC,
10oC/min to 260oC, hold for 5 min
35oC/min to 260oC, hold for 7 min
30.114 min
Oven Program
Duration
To quantify NP1EC in the sludge samples, 1 µl of the sample was injected in the
splitless mode at 250oC into a same column used for other NP compounds. The
GC/MS was operated in Selective Ion Mode (SIM). The mass spectrometer was
operated in the electron impact ionization (EI) mode and the energy of the electrons
was kept at 70 eV. The interface, MS ionization source and the MS quadrupole were
57
kept at the same temperature as in the method developed for NP, NP1EO and
NP2EO. The details of the oven program used are given in Table 3.5.
Table 3.5. GC/MS method for determination of NP1EC
GC/MS Method
Column
HP-5MS 5% Phenyl Methyl Siloxane
Carrier Gas
Helium (1 mL/min)
Injection volume
1 µL
Injection Mode
Splitless
Injection Temperature
250°C
MS Interface Temperature
280°C
MS Source Temperature
230°C
MS Quadrupole Temperature
150°C
MS Mode
SIM
70oC, hold for 1 min,
30oC/min to 160oC,
5oC/min to 290oC, hold for 5 min
35 min
Oven Program
Duration
3.4.6.3. Derivatization
Volatility of a sample is an essential requirement for GC/MS analysis. It is difficult
to analyze compounds by GC/MS that have poor volatility and thermal instability
(Danielson et al. 2000). For these compounds, a derivatization process is generally
applied prior to GC/MS analysis to make these compounds sufficiently volatile so
that they can be eluted at reasonable temperatures without thermal decomposition
(Orata 2012). It has been reported that an increase in ethoxy chain length of NP
compounds is accompanied by increase in detection limit due to lowering of the
volatility (Hoai et al. 2003).
58
The details of the method development for the derivatization procedure are given in
Chapter 4. The developed derivatization procedure is summarized in Figure 3.3.
Figure 3.3. Details of derivatization procedure applied for NP, NP1EO, NP2EO
compounds
For NP1EC derivatization, Lee and co-workers suggested the methylation of NP1EC
in the presence of boron trifluoride (BF3, 20% in methanol) as a derivatization
method (Lee et al. 1997). This method produces methylated derivatives of this
compound and results in a peak suite of the compound. This derivatization method
was applied with minor changes as given in Figure 3.4.
The details of experimental trials for derivatization of NP1EC compound are given
in Chapter 4.
59
Figure 3.4. Details of derivatization procedure applied for NP1EC.
Following derivatization of all NP compounds, in order to determine most abundant
ions, SCAN mode was used by monitoring the mass range from 50 to 600. After
identification of m/z values that defines the unique fingerprint for each compound,
quantitative analyses were carried out using SIM mode in GC/MS.
Method development including literature studies, GC/MS analyses and derivatization
are given in details in Chapter 4.
60
Following the capability of quantification of each NP compound, the calibration
curves for 4-n-NP, NP, NP1EO, NP2EO and NP1EC were prepared using standard
solutions prior to extraction of these compounds from the sludge samples. An 8point calibration curve with r2 >0.99 in the concentration range of 10 ppb to 1000
ppb was used for the quantification of 4-n-NP, NP, NP1EO and NP2EO. For
NP1EC, a 6 point-calibration curve with r2 >0.99 was constructed in the 50-1000 ppb
concentration range.
3.4.6.4. Extraction of NP Compounds from Solid Phase of Sludge Samples
To be able to measure NP compounds in samples taken from the digesters, an
extraction procedure was followed. First sludge samples were centrifuged at 2500
rpm for 10 min for the separation of solids from the liquid. After separation, NP
compounds were extracted from each phase by applying the extraction methods
developed in our laboratory.
There was a variety of methods in the literature that have been developed and used
for the extraction of NP compounds from solid/sludge samples (Long et al. 1998,
Lye et al. 1999, Liu et al. 2004, Gibson et al. 2005, Andreu et al. 2007, Núñez et al.
2007). The main purpose for the development of a new extraction method was to
improve the extraction performance as well as to reduce overall analysis time and
cost. With these evaluations, the considerable reduction in extraction time, solvent
consumption and energy, made sonication-assisted extraction (SAE) a very attractive
extraction method over other methods (soxhlet extraction, mechanical shaking,
microwave assisted extraction etc.).
Method development and optimization studies for extraction of NP compounds from
solid phases of sludge samples are given in details in Chapter 4.
61
The developed sonication-assisted extraction method depicted in Figure 3.5 was
applied to solid fractions of sludge samples obtained from lab-scale semi-continuous
anaerobic digesters.
Figure 3.5. Details of sonication-assisted extraction method applied to solid phase of
sludge samples for extraction of NP compounds
Triplicate extractions were carried out for each sample. Then, GC/MS analyses were
carried out as duplicate readings from each of triplicate extracts of the same sample;
thus, six readings in total were carried out from the same sample.
Following extraction the remaining extracts (in acetone) were put into amber vials
with PTFE-lined caps and were kept at -18oC in the dark.
62
3.4.6.5. Extraction of NP Compounds from Liquid Phase of Sludge Samples
Solid phase extraction (SPE) is the most common method used for the extraction of
analytes from water samples due to high concentration ability and recovery,
simplicity and low consumption of solvents compared to traditional liquid-liquid
extraction method. For these reasons it has been widely used for extraction of NP
compounds from water, surface water and wastewater samples (Di Corcia et al.
2000, Liu et al. 2004, Lee et al. 2005, Stasinakis et al. 2008). Therefore, SPE method
was chosen to extract and concentrate the NP compounds from liquid phases of
sludge samples. Method development and optimization studies for extraction of NP
compounds from liquid phases of sludge samples are given in details in Chapter 4.
Following method development, liquid portions of sludge were taken into extraction
vials and optimized SPE procedure was applied as summarized in Figure 3.6 and
depicted in Figure 3.7.
Figure 3.6. Application steps for SPE method for extraction of NP compounds from
liquid phases of sludge samples
63
Figure 3.7. Visual illustration of SPE method
3.4.6.6. Recovery
Recovery is a parameter used for evaluation of efficiency of methods applied for
extraction and measurement of target compounds. Recovery studies were carried out
for each compound based on the surrogate (4-n-NP). In order to determine extraction
efficiency of sonication-assisted extraction method applied for extraction of NP
compounds from solid phases of sludge samples, spiking of 0.5 g-clean sludge
samples were conducted at a concentration of 50 µg/L (0.1 ng/mg, in acetone) for
each compound and surrogate. Surrogate (4-n-NP) was added into spiked sludge
samples for evaluation of efficiency and precision of extraction methods. Same
procedure was also repeated for determination of extraction efficiency of extraction
method applied for NP compounds from liquid phases of sludge samples. Blanks
were also prepared during recovery studies to reveal matrix effect.
64
Due to assessing recovery of each compound based on surrogate, the recovery of
surrogate was determined first by using the following formula:
Recovery = R (%) =
X100
(3.1)
where;
Ce is the concentration of surrogate measured in the spiked sample
Ct is the theoretical concentration of surrogate spiked into the sample
Recovery values for target NP compounds were calculated by using following
formula:
Recovery = R (%) =
−
X100
(3.2)
where;
Cs is the concentration of target compound measured in the spiked sample
Cu is the concentration of target compound measured in the unspiked (blank) sample
Ct is the theoretical concentration of target compound spiked into a sample
Relative standard deviation (RSD) is widely used to point out the precision and
repeatability of an analysis. It is generally expressed as percentage (%RSD) and
calculated based on following formula:
RSD (%) =
100
(3.3)
where,
s is the standard deviation of a series of analysis and X represents the mean value for
that.
65
The recovery studies carried out in this study are presented in details in Chapter 4.
3.4.7. Quality Assurance/Quality Control (QA/QC)
3.4.7.1. Glassware
In order to meet QA/QC standards all glassware used in analyses was cleaned with
Alconox detergent for 1 h and then rinsed consecutively with: i) deionized water; ii)
chromic-sulfuric acid (technical grade), iii) deionized water, and iv) acetone
(technical grade). All glassware was dried at 65°C overnight.
3.4.7.2. Linearity
The linearity of the response of NP compounds was determined by analyzing a series
of 8 standards in the concentration range of 0-1000 µg/mL during preparation of
calibration curves (6 standards for NP1EC). The peak area of each compound and
internal standard (4-n-NP) was plotted against the concentration range of the
standards which exhibited a linear relationship with the correlation coefficients (r2
>0.99). In order to apply the linear regression analysis for quantitative purposes, r2
must be ≥0.99 according to USEPA Method 8000C (USEPA, 2003). The obtained
data for target NP compounds are summarized in Table 3.6.
66
Table 3.6. Quantitative calibration data for NP compounds
Compound
Concentration range
Equation
Correlation coefficient
(r2)
(µg/mL)
4-n-NP
10-1000
y=303413x-3E+06
0.999
NP
10-1000
y=216895x-4E+06
0.996
NP1EO
10-1000
y=16724x-80723
0.999
NP2EO
10-1000
y=12502x-123871
0.997
NP1EC
50-1000
y=60281x+282810
0.992
Stock solutions of NP compounds were prepared in acetone as 1000 µg/mL at
analysis day. Required dilutions to prepare standards at different concentrations were
conducted by using stock standard solution. Standards were prepared as three
replicates and duplicate injections of each standard were performed in calibration
curve preparation.
The calibration curves were re-prepared when some changes to the GC or the
program was made (cleaning of GC, changing of liner, column, etc.). Also, the
condition of GC/MS was checked regularly by injection of one or two standard(s)
during analysis. The prepared calibration curves applied for quantification of NP
compounds during the operation of semi-continuous anaerobic digesters are given in
Appendix A.
3.4.7.3. Repeatability
Repeatability is the variation in independent measurements performed using the
same method and equipment in the same laboratory by the same operator under the
same conditions. A measurement can be considered as repeatable when the variation
in repeated measurements is smaller than the proposed limit.
To test repeatability of developed method, known concentrations of standards for
each compound (20 µg/mL and 500 µg/mL) were injected to GC/MS (n=6) and
67
obtained average concentration values and RSDs are evaluated. Results obtained are
demonstrated in Chapter 4.
3.4.7.4. Limit of Detection and Quantification (LOD and LOQ)
The limit of detection (LOD) and limit of quantification (LOQ) parameters were
determined for each compound for both extraction methods. The LOD can be
defined as the minimum concentration of a substance which could be detected in a
sample with high level of confidence (99%). LOQ is the minimum concentration of a
substance in a sample which can be quantitatively determined with satisfactory
precision and accuracy.
In this study, LOD and LOQ values were determined using signal to noise (S/N)
ratio (ICH, 1996). The ratio represents the quality of the signal obtained from the
sample.
LOD and LOQ studies were performed by spike of known low concentrations of
standard solutions on sludge samples (ICH, 1996). Injections were carried out (n=6)
for each compound with blanks and following equations were used for calculation of
LOD and LOQ;
LOD = 3 ( )
(3.4)
LOQ = 10 ( )
(3.5)
For each compound, the calculated LOD and LOQ values based on Equation 3.4 and
3.5 are presented in Chapter 4.
68
3.5. Molecular Analyses
3.5.1. DNA Extraction
In order to isolate genomic DNA from sludge samples with high quality and yield,
different commercial DNA extraction kits were tested. MoBio PowerSoil DNA
Extraction Kit (MoBio Laboratories Inc., USA), E.Z.N.A. Soil DNA Extraction Kit
(Omega Bio-Tek, USA) and NucleoSpin Soil DNA Extraction Kit (Macherey-Nagel,
GmbH & Co. KG, Germany) were purchased and used in comparison studies. A
DNA extraction kit special for sludge samples has not been manufactured yet,
therefore soil DNA extraction kits were used for this purpose as reported in literature
(Bonot et al. 2010, Bushon et al. 2010, Guo and Zhang 2013).
Application of physical disruption techniques on sludges during extraction procedure
is a critical step due to need for complete homogenization and cell lysis. Bead mill
homogenization and freeze-thaw disruption techniques have been widely reported
for soil and sludge samples in literature (Tsai and Olson 1991, Bruce et al. 1992,
Hugenholtz et al. 1998, Miller et al. 1999). Due to the absence of bead mill
homogenizer in our laboratory, sludge samples were treated with a method
developed by a combination of freeze-thaw and high-speed vortex. For this purpose,
0.5 g sludge sample (wet weight) was taken into each tube including beads and
disruption solution (provided by kit) was added. The tubes were placed into 70oC
water bath for 10 min. After that, they were put on ice for 5 min. These steps were
repeated once again. Then, the tubes were placed horizontally on a flat-bed vortex
pad and secured with tape. They were vortexed at high speed for 15 min. The aim of
this step was to provide collision of the beads with microbial cells randomly in the
presence of disruption chemicals (like sodium dodecyl sulfate) to break cell structure
and provide cell lysis. Following disruption/homogenization step, the instructions
were applied given in manufacturer’s protocol of each DNA extraction kit. DNA was
69
eluted from silica spin-filter columns with 50 µL DNA-Free PCR grade water. For
each sludge sample and each kit, triplicate extraction was carried out.
The extracted DNA must be free of contaminants that can inhibit subsequent
molecular applications, such as PCR, qPCR etc. So the performance of three DNA
extraction kits was evaluated based on their total yield (as concentration) and purity
of DNA. These parameters were determined by UV-Visible Spectrophotometer
(NanoDrop 2000, Thermo Scientific, USA). OD260/OD280 and OD260/230 ratios
were used to assess the quality of DNA. A ratio of 1.8 for OD260/OD280 is accepted
as “pure” for DNA. If the OD260/OD280 ratio is lower than 1.8, it indicates the
presence of proteins and phenols. On the other hand, for OD260/OD230 ratio, the
acceptable range is between 2.0 and 2.2. If the OD260/OD230 is lower than this
range, it shows the presence of contaminants like humic acids, residual phenol and
aromatic compounds.
Total genomic DNA concentration and the ratios indicating the purity of extracted
DNA for each kit are presented in Table 3.7. As can be seen from results, Omega
E.Z.N.A Soil Extraction Kit (OE-SE) yielded significantly better results for DNA
concentration and purity than the other two kits. Therefore, DNA from sludge
samples taken from the digesters was isolated using OE-SE kit.
70
Table 3.7. Performance of DNA extraction kits evaluated in terms of concentration
and purity
Kit
A260/280
A260/230
75.3±3.6
1.8
2.1
428.4±18.3
1.8
2.0
93±5.1
1.7
1.8
DNA concentration
(ng/µL)
MoBio PowerSoil DNA Extraction Kit
(MP-SE)
Omega E.Z.N.A. Soil DNA Extraction Kit
(OE-SE)
NucleoSpin Soil DNA Extraction Kit (NSSE)
The results indicated pure DNA by the OD260/OD280 and OD260/OD230 values of
1.8-2.0 and 2.0-2.2, respectively in all extractions. After quantification of sludge
DNA samples, the quality of DNA samples (1:100 diluted) was also evaluated by
1.0% agarose gel in TAE buffer with GelRed (Biotium Corp., USA) for 1 h at 80 V.
The gels were visualized with Quantum ST-4 3000 Gel Image Acquisition System
(Montreal Biotech Inc., Canada).
All DNA samples were stored at -20 oC for downstream applications.
3.5.2. Polymerase Chain Reaction (PCR)
The extracted genomic DNA from sludge samples was used as a template for PCR
amplification of target regions for microbial community analysis. The specificity of
primer pairs given in Table 3.8 for target phylogenetic groups was tested with PCR
analysis. Also, the optimum annealing temperature for each primer set was
determined by performing gradient PCR. Following determination of ideal annealing
temperature for each primer set, qPCR assays were conducted.
The PCR reaction mix contained 1× PCR buffer (Thermo Scientific, USA), 4 mM
MgCl2, 400 µM dNTPmix, 500 nM of each primer and 2.5 U of Taq DNA
71
polymerase (Thermo Scientific, USA) for a 50-µL PCR reaction. PCR reactions
were performed with Thermo Scientific thermal cycler (Thermo Scientific, Arktik,
USA). The cycler program was set to conduct 30 cycles of 95°C for 1 min, T°C for
1 min, and 72°C for 1 min followed by a final extension step of 10 min at 72°C. PCR
products were analyzed by agarose gel (1.2%) electrophoresis in TAE buffer and
visualized with Quantum ST-4 3000 Gel Image Acquisition System (Montreal
Biotech Inc., Canada).
Also, PCR amplification of 16S rDNA from pure cultures and reference strains was
performed using 27F (5′-AGA GTT TGA TCM TGG CTCAG-3′) and 1492R (5′TAC GGY TAC CTT GTT ACG ACTT-3′) primer set (Lane 1991). The reaction
mixture of 25 µL contained 2.5 µL of Taq PCR buffer, 2 µL of dNTPs (200 µM
each), 1.25 µL of each primer (10 µM), 2 µL of MgCl2 (2 mM), 1.25 U Taq DNA
polymerase, 1 µL of template DNA (dilution 1:50), and sterile water to complete the
solution to 25-µL volume. PCR reactions were performed in thermal cycler, with 30
cycles consisting of denaturation at 94°C for 1 min, annealing at 54°C for 1 min, and
extension at 72°C for 2 min, and a final extension at 72°C for 8 min. The
electrophoresis of PCR products (4 µL) was conducted at 80 V on a 1.2% agarose
gel in TAE buffer. The gel was stained with GelRed dye and visualized under gel
image system.
3.5.3. Quantitative PCR (Real-Time PCR)
3.5.3.1. Primers
In recent years, quantitative PCR (qPCR, also referred to as real-time PCR) has
become a promising tool for studying microbial communities in complex
environmental samples like soil, sediment, activated sludge etc. (Grüntzig et al.
2001, Hermansson and Lindgren 2001, Bach et al. 2002, Harms et al. 2003, Wéry et
al. 2008, Srinandan et al. 2011, Jang et al. 2014). The qPCR approach enables
72
quantification of the abundances of specific microbial groups in community structure
in a rapid and reliable way. Therefore, the abundances of target microbial groups
within the general microbial community available in the semi-continuous anaerobic
digesters were characterized by qPCR assays in this study using primers listed in
Table 3.8. Details are given in Chapter 6.
Table 3.8. Specific primers used in qPCR assays
Target Group
Forward
Reverse
Primer
Primer
Bacteria
Eub338
Eub518
Alphaaproteobacteria
Eub338
Alf685
Betaproteobacteria
Eub338
Bet680
Gammaproteobacteria
Gam1080
Gam1202
Deltaproteobacteria
Del361
Del685
Archaea
Arc349
Arc806
Methanosaeta
MS1b
Sae835
Methanosarcina
Mb1b
Sar835
Sequence of primer pairs
(5´ to 3`)
F: ACT CCT ACG GGA GGC AGC AG
R: ATT ACC GCG GCT GCT GG
F: ACT CCT ACG GGA GGC AGC AG
R: TCT ACG RAT TTC ACC YCT AC
F: ACT CCT ACG GGA GGC AGC AG
R: TCA CTG CTA CAC GYG
F: TCG TCA GCT CGT GTY GTG A
R: CGT AAG GGC CAT GAT C
F: AAG CCT GAC GCA SCA A
R: ATC TAC GGA TTT CAC TCC TAC A
F: GYG CAS CAG KCG MGA AW
R: GGA CTA CVS GGG TAT CTA AT
F: CCG GCC GGA TAA GTC TCT TGA
R: GAC AAC GGT CGC ACC GTG GCC
F: CGG TTT GGT CAG TCC TCC GG
R: AGA CAC GGT CGC GCC ATG CCT
3.5.3.2. qPCR Assays
qPCR assays were performed in a 25 µL reaction mixture that consisted of 12.5 µL
of EvaGreen Ssofast Supermix with Low ROX (Biorad, USA), 1.25 µL of each
primer (0.5 µM), 5.0 µL of template DNA and 5.0 µL sterile dH2O. Sludge DNA
73
samples were used at a concentration of 1–5 ng/μL. To avoid PCR amplification
problems due to the presence of inhibitors, DNA samples were diluted 10 to 100
times in sterile water (nuclease-free). Each reaction mixture was loaded in
polypropylene 96-well plates on an ABI 7500 Real Time PCR System (Applied
Biosystems, USA). qPCR conditions were 3 min at 98°C, followed by 35 cycles of
98°C for 15 s and 1 min at the annealing temperature. qPCR reactions were carried
out in triplicate for each DNA sample and the appropriate set of standards.
Melting curve analysis was carried out for each assay by monitoring fluorescence
continuously between 60 and 95°C with 0.5°C increments (Figure 3.8). Melting
curve analysis is a necessary step to check the specificity of the amplified PCR
products. Non-specific or primer-dimer products can be produced during qPCR
reaction and show up as an additional peak aside from the amplicon (PCR product)
peak in melting curve plot.
Figure 3.8. Derivative melting curve analysis for standards in qPCR
74
3.5.3.3. Standard Curve Preparation
In order to quantify amount of target DNA in samples, external calibration curves
are prepared to reproduce highly specific and reliable data. However, the validation
of the external calibration curve used for quantification in quantitative PCR (qPCR)
depends entirely on the accuracy of the standards. Firstly, in this study, genomic
DNA based standards were prepared and assayed in qPCR but the standards did not
produce reliable and reproducible data. Therefore plasmid standards were prepared
and used for quantification in qPCR assays. In most of studies, it has been reported
that use of plasmid-standards for quantification performs better and generates more
reproducible and precise results than the use of standards made from genomic DNA
in qPCR assays (Becker et al. 2000, Suzuki et al. 2000, Whelan et al. 2003, Lee et al.
2006, Li et al. 2010, Rungrassamee et al. 2013).
To prepare a plasmid standard for each microbial group given in Table 3.8, the target
region was amplified by PCR using genomic DNA extracted from the relevant
positive control strain. The PCR products were run on a 1.2 % agarose gel for
confirmation of the amplification specificity. A clean-up step was applied for these
PCR products by using Gel and PCR Clean-up kit (Macherey-Nagel, GmbH & Co.
KG, Germany). Then, cleaned PCR products for each group were cloned into
pGEM®-T Easy vectors (Promega, USA).
pGEM®-T Easy vector has a multiple cloning site (MCS) within the lacZ
gene which encodes β-galactosidase (β-gal). When the PCR product is inserted
within the lacZ gene, the gene gets disrupted. Plasmids with insert are called
recombinant plasmid and these plasmids are transformed into competent cells
(Escherichia coli JM109 provided by the kit).
Following transformation of
recombinant plasmids into competent cells, they can be selected on LB agar plates
with Ampicillin/IPTG/X-Gal by blue-white screening. Ampicillin (Amp) is added
into LB agar plates for selection of competent cells with recombinant plasmids
75
acquired the ampicillin resistance ability as a result of successful insertion. X-Gal (5bromo-4-chloro-indolyl-β-D-galactopyranoside)
and
IPTG
(Isopropyl
β-D-1-
thiogalactopyranoside) are spread on LB/Amp agar plates for application of bluewhite screening. X-Gal is analogue of lactose and when β-galactosidase hydrolyses
X-Gal, it produces a blue color due to formation of a blue product called as 5,5'dibromo-4,4'-dichloro-indigo. As can be understood, the competent cells with
recombinant plasmids have no ability to hydrolyze X-Gal due to disruption of lacZ
gene by insertion. Therefore, white colonies growing on Amp/IPTG/X-Gal LB plates
indicates the cells transformed with recombinant plasmids. On the other hand,
competent cells with non-recombinant plasmids (no insert) form blue colonies on
Amp/IPTG/X-Gal LB plates due to being able to produce β-galactosidase and
hydrolyze X-Gal into blue color product. IPTG functions as the inducer of the lac
operon so IPTG is used to enhance the expression of lacZ gene. One of
Amp/IPTG/X-Gal LB plates with white and blue colonies is presented in Figure 3.9.
The white colonies have recombinant plasmids with PCR product specific for βProteobacteria.
Figure 3.9. Blue-white screening specific for β-Proteobacteria
76
The procedure mentioned above was applied to perform cloning and blue-white
screening for each PCR product targeting specific microbial group aimed in this
study.
Following cloning and blue-white screening, 20 white colonies representing each
target plasmid were selected randomly on Amp/IPTG/X-Gal LB plates. A colony
PCR
was
carried
out
for
each
selected
colony
using
T7F
(5'-
TAATACGACTCACTATAGGG-3') and M13R (5'-CAGGAAACAGCTATGAC3') vector-specific primer pairs to determine if they contain the insert of interest. For
colony-PCR, the reaction contents and conditions used are as follows:
Colony-PCR reaction:
Colony-PCR condition:
5 µL Mastermix
940C for 5 min
1 µL T7F primer (500nM)
940C for 1 min
1 µL M13R primer (500nM)
450C for 1 min
Colony
720C for 1 min
35 cycles
720C for 10 min
2 µL dH2O
To screen white colonies for the insert of interest, a colony-PCR was also performed
with insert-specific primers pairs. The reaction mix and conditions are given in
Section 3.5.2.
All colony-PCR products were run on a 1.2% agarose gel in TAE buffer to check the
inserts based on their expected size. As an example, the images of agarose gels for
77
PCR products obtained using vector-specific and insert-specific primer pairs are
illustrated in Figure 3.10. The length of PCR product inserted into plasmid specific
for β-Proteobacteria is 365 bp. When T7F/M13R primer pair was used for colony
PCR, it introduced extra ≈ 200 bp in addition to insert size (for this case, 365 bp). As
can be seen Figure 3.10, the results indicated that 5 out of 7 of randomly chosen
white colonies have target plasmids with insert of interest in proper size. Following
screening of recombinant plasmids by colony PCR for each specific group,
recombinant plasmids were isolated using a Plasmid Miniprep Kit (Promega).
Quality and concentration of isolated plasmid DNA were determined with UV-Vis
Spectrophotometer.
Figure 3.10. Agarose gel images following colony PCR with i) vector-specific
T7F/M13R primer pair (left side) ii) insert-specific primer pair (right side) for βProteobacteria. L: DNA Ladder (Gene Ruler, 100 bp plus)
78
Standard curves were prepared using triplicate 10-fold serial dilutions (103 –109) of
plasmid DNA containing cloned PCR product for target microbial group. The copy
number of standard plasmids was calculated based on plasmid (3015 bp) plus insert
lengths and assuming a molecular mass of 660 Da for a base pair. One of the
obtained calibration curves prepared using standard plasmid DNA is given in Figure
3.11. The rest of the curves can be seen in Appendix B. As it is presented, there is a
linear relationship between the log of the plasmid DNA copy number and the
threshold cycle (CT) value. This linearity was observed for all plasmid standard
curves with correlation coefficient (R2) values ranged from 0.996 to 0.999
(Appendix B).
Figure 3.11. A standard curve constructed for quantification of abundance of αProteobacteria in sludge samples
Amplification efficiency (E) of qPCR assays was calculated by using the slope of the
standard curves and the following formula (Pfaffl, 2001):
79
The calculated amplification efficiency values for all qPCR assays conducted in this
study were varied between 1.7 and 2.1 (90-110% as percentage) reported as
acceptable range (Pfaffl 2001).
3.5.4. Fluorescence in Situ Hybridization (FISH)
Fluorescence in situ hybridization (FISH) is one of the widely used cultivationindependent molecular tools for investigation of microbial community structure in
mixed culture systems such as activated sludge systems (Amann et al. 1995,
Zimmermann et al. 2001, Kenzaka et al. 2005). In FISH analyses, specific rRNAtargeted oligonucleotide probes are used for detection and quantification of
microorganisms in their ecological environments. In this study, FISH method was
applied for identification and quantification of target microbial groups in operated
anaerobic digesters in the presence of NP2EO. Obtained data were used to evaluate
the changes in microbial groups composing the community structure of digesters
before and after NP2EO spike.
3.5.4.1. FISH probes
An oligonucleotide probe is a single-strand of DNA and generally consists of 15–25
bases. Its sequence is complementary to a region of 16S rRNA of target
microorganism studied. For application of FISH method, oligonucleotide probes are
labeled with a fluorescence dye that can be excited by light of a given wavelength
(You et al. 2003).
80
Oligonucleotide probes used in this study are listed in Table 3.9 and details are given
in Chapter 6. The probes were synthesized and labeled from 5’ end with Cyanine 3
(Cy3) dye by Alpha DNA, Canada. Cyanine dyes are brighter and have greater
photo-stability than other fluorescein dyes. Cy3 is excited maximally at 550 nm and
emits maximally in the red end of the spectrum at 570 nm.
DAPI (diamidino-2-phenylindole) is a blue fluorescent probe that fluoresces brightly
when specifically bind to the minor groove of double stranded DNA. DAPI stains all
cells without differentiation between live and dead ones. Therefore it is used as
positive internal standard in FISH analyses.
Table 3.9. Specific oligonucleotide probes used in FISH analysis
Specificity
Probe
Sequence (5´……3`)
Most Bacteria
EUB338MIX
GCWGCCWCCCGTAGGWGT
Negative Control
NON338
ACTCCTACGGGAGGCAGC
Alphaaproteobacteria
ALF968
GGTAAGGTTCTGCGCGTT
Betaproteobacteria
BET42a
GCCTTCCCACTTCGTTT
Gammaproteobacteria
GAM42a
GCCTTCCCACATCGTTT
Deltaproteobacteria
SRB385
CGGCGTCGCTGCGTCAGG
Archaea
ARC915
GTGCTCCCCCGCCAATTCCT
Methanosaeta
MX825
TCGCACCGTGGCCGACACCTAGC
Methanosarcina
MS821
CGCCATGCCTGACACCTAGCGAGC
81
3.5.4.2. Preparation of Gel Coated slides
To provide the best possible adhesion and minimization of sample (biomass) loss,
glass microscope slides (VWR International, USA) were coated with gelatin (Daims,
2009). Following heating of 500 mL of deionized water up to 60oC, 2.5 g gelatin
(Type A, Sigma Aldrich Co. LLC, USA) and 0.25 g chromium potassium sulfate
(CrK(SO4), Sigma Aldrich Co. LLC, USA) were added and dissolved while stirring.
Then, the gelatin-coating solution was cooled at room temperature and filtered.
Slides were dipped into solution five times and then dried at room temperature for 2
days. The gelatin coated glass slides were stored in dust-free storage boxes.
3.5.4.3. Fixation
Sludge samples collected from semi-continuous anaerobic digesters were fixed with
4% paraformaldehyde (PFA) in phosphate buffered saline (PBS, 137 mM NaCl, 2.68
mM KCl, 1.47 mM KH2PO4, 8.1 mM Na2HPO4.7H2O, pH 7.4). Following overnight
fixation, the samples were washed twice with PBS and then kept in PBS:ethanol
(1:1, v:v) at -20oC until analysis. PFA forms cross-link with proteins and
peptidoglycan of cell walls to keep cell morphology. Also, fixation step increases
cell permeability.
Following washing steps of fixed-sludge samples with PBS, sludge samples were
sonicated by using sonication probe (3 mm diameter, LabSonic P, Sartorious Corp.,
France) to detach cells from particles and disaggregate cell flocs. Sonication was
carried out at an amplitude of vibration of 20% for 15 s. After sonication step, the
samples were suspended in PBS:ethanol (1:1, v:v) and stored at -20oC.
3.5.4.4. Dehydration
10 µL sample was added onto a gelatin-coated glass slide and dried for 10 min at
46oC. Then the slides were placed into tubes containing an ethanol series (50-80-
82
96%) in increasing order for 3 min each to provide dehydration. The slides were airdried prior to passing hybridization step. The removal of water from the fixed-sludge
samples is essential to increase the resolution during microscopy analysis.
3.5.4.5. Hybridization
Hybridization step is critical to be carried out at optimum conditions in terms of
temperature, salt concentration (ionic strength) and formamide percentage. The
formamide concentration is dependent on the probe used and affects the stringency
of the hybridization (Nielsen et al. 2009). It is more convenient to perform
hybridization step at incubator or water bath at set temperature.
A 15 µL hybridization buffer (900 mM NaCl, 20 mM Tris-HCl (pH 7.4), 0.01% SDS
and X% formamide depending on probe used) and 5 µL of probe (30 ng/µL) were
added over the sample. Then, the slides were incubated at 46oC for 3 hours. After
incubation, the slides were washed with washing buffer (Y mM NaCl (depending on
percentage of formamide used in hybridization buffer), 20 mM Tris-HCl (pH 7.4), 5
mM EDTA (pH 8), 0.01% SDS) for 7 min. This step was repeated and then the
slides were rinsed with ddH2O. After air-drying of slides, counterstaining was
performed with addition of 2 µL of DAPI solution (5 µg/mL, Merck, Germany) on
each sample. The slides were incubated with DAPI for 5 min and then rinsed with
ddH2O.
Following air-drying, the samples were mounted in 5 µL of DABCO (1,4diazabicyclo[2.2.2]octane, Sigma Aldrich, USA) or Citifluor (Citifluor Ltd.,
Canterbury, UK). DABCO and Citifluor are anti-fading agents, in other words, they
retain the fluorescence of the samples during the analysis for a long time. They also
provide storage of the samples at -20oC in the dark for several days without loss of
probe fluorescence.
83
3.5.4.6. Visualization
The slides were examined with a Carl Zeiss Axio Scope.A1 epifluorescence
microscope equipped with a 100 W high pressure mercury lamp and a charged
coupled device (CCD) camera. Zeiss software was used while taking images. For
each slide, 10 random fields of view (FOV) were obtained. For each randomly
chosen FOV, one image was captured for the population-specific FISH probe signal
and another image was obtained for the DAPI signal. In other words, the same FOV
is represented by two visualization ways: DAPI and the specific fluorochrome
selected. In this study, Cy3 fluorochrome was used for all probes. Cy3 was excited
by green light and emitted red fluorescence.
3.5.4.7. Image Processing and Analysis
Images were analyzed by using the DAIME and Image J fluorescence microscopyimage analysis programs. Depending on the different algorithms offered by the
image analysis programs, the best algorithm was tried to be determined and used for
analysis.
The schematic presentation of application steps for FISH analysis is illustrated in
Figure 3.12.
84
Figure 3.12. Visual representation of FISH analysis
3.5.5. Enrichment Studies
3.5.5.1. Chemicals
Technical grade nonylphenol (tNP, Pestanal) was purchased from Sigma-Aldrich (St.
Louis, MO). All solvents were high-performance liquid chromatography (HPLC)
grade and were supplied from Merck (Darmstadt, Germany). All salts and medium
components were also purchased from Merck (Darmstadt, Germany). All restriction
enzymes and buffers were supplied from Promega (Promega Corperation, WI, USA).
For synthesis of nonylphenol isomers, BF3-ether complex was supplied from Merck
(Darmstadt, Germany). Diethyl ether, dichloromethane, ethyl acetate and petroleum
85
ether (boiling range from 60oC to 90oC) were obtained as pure solvents from Merck
(Darmstadt, Germany). BF3-ether complex was purchased from Sigma Aldrich
(Sigma Aldrich Co. LLC, USA). 3’,5’-dimethyl-3’-heptanol and 3’,6’-dimethyl-3’heptanol were supplied from Alfa Aesar (MA, USA).
3.5.5.2. Media
Mineral salt medium (MSM) was used for the enrichment studies and prepared
according to study of McCullar et al. (McCullar et al. 1994). For this purpose, tNP
was added as the sole carbon and energy source into sterile MSM medium by using a
sterile syringe. For the preparation of MSM agar plates with tNP, noble agar was
added as 16 g/L into MSM and then autoclaved for 15 min at 121°C. Nonylphenol is
not stable at elevated temperatures and not possible to sterile by using a 0.22 µmfilter due to its viscous structure. Therefore, non-sterile tNP and some MSM salts
(0.22 µm filter sterilized) were added after autoclavation. The medium was agitated
for emulsification of tNP and then poured into plates.
Liquid Luria Bertani (LB) medium was used for culturing of bacterial isolates and
storing them in 20% glycerol stock at -80°C. The components of LB medium for a
liter were as follows: 5 g sodium chloride, 10 g tryptone and 5 g yeast extract. LB
agar plates were also used for cultivation of the isolates. All media were prepared
with ultrapure water (Millipore) and pH of the medium was adjusted to 7.0 with
NaOH.
3.5.5.3. Enrichment and Isolation
Waste activated sludge (WAS) from the central municipal wastewater treatment
plant in Ankara, Turkey was used for enrichment studies. 250 mL of WAS sample
was spiked with 1 mg/mL tNP and incubated on a shaker at 28±2oC and 120 rpm for
10 days in the dark. After 10 days, 1 mL of enrichment culture was transferred into
Erlenmeyer flasks (250 mL) containing 100 mL of MSM and 1 mg/mL tNP. The
86
flask was incubated for 10 days again. Then, 1 mL sample was taken from MSM
medium showing the growth and added into sterile 100 mL-MSM medium
containing 1 mg/mL tNP. This step was repeated three more times. The growth and
degradation were evaluated by turbidity and monitored by measuring optical density
(OD) values at 550 nm using Hach-Lange DR 3900 Spectrophotometer (Hach Lange
GmbH, Germany). With the last transfer, the enrichment culture was serially diluted
(in sterile 0.85% NaCl) and spread on MSM agar plates with 1 mg/mL tNP. Plates
were incubated at 28±2oC nearly for a month. The strains with an ability of using
tNP as a sole carbon and energy source were isolated from tNP-MSM agar plates. In
order for identification, the bacterial isolates with different colony morphologies
were determined and then restriction enzyme analysis was carried out. Cell
morphology of the isolates was determined by observing DAPI stained cells under a
microscope. Biochemical tests for the identification of the isolates were performed
using standard methods.
3.5.5.4. DNA Extraction
All isolates grown in 1/10 liquid LB medium at 28±2oC with shaking overnight were
used for DNA extraction. Genomic DNA extraction from isolates and reference
strains was carried out by using Omega E.Z.N.A. Genomic DNA Extraction Kit
(Omega Bio-Tek, USA). Concentration and purity of isolated DNA were determined
using UV-Visible Spectrophotometer (NanoDrop 2000, Thermo Scientific, USA).
The results indicated pure DNA by the OD260/OD280 and the OD260/OD230
values of 1.8-1.85 and 2.0-2.2, respectively in all extractions. DNA samples were
also evaluated by 1.0 % agarose gel in TAE buffer with GelRed (Biotium Corp.,
USA) for 1 h at 80 V.
All DNA samples were stored at -20 oC for downstream applications.
87
3.5.5.5. Restriction Enzyme Analysis
Restriction enzymes (RE) are able to recognize specific sequences in DNA and cut
DNA from recognition sites to produce fragments. These restriction fragments form
a unique pattern and are separated by gel electrophoresis based on their size. In this
study, RE analysis was used to compare 16S DNA patterns of environmental isolates
following digestion with two different restriction enzymes called HaeIII and HhaI.
These enzymes are four-base hitters and cleave DNA within their recognition
sequences.
For RE analysis, genomic DNA of environmental isolates was extracted using
Omega E.Z.N.A. Genomic DNA extraction kit. Following extraction, the extracted
DNA was used as the template for PCR to amplify 1,500 bp of the 16S rRNA gene
sequence of environmental isolates using 27F and 1492R primer set (Lane 1991).
The PCR conditions were applied as mentioned in Section 3.5.2. PCR products were
run on a 1.2% agarose gel and visualized under UV light by staining with GelRed
dye. After amplification of 16S rDNA, PCR products were digested separately with
HaeIII and HhaI (Promega, USA) restriction enzymes. For each digestion, 5 µL of
PCR product was added into a digestion mixture containing 1 µL restriction enzyme
(10u/µL), 2 µL restriction buffer (10X Buffer C) and 12 µL water (nuclease-free) in
a 20-μL reaction volume. PCR products were digested at 37oC for 4 hrs in a water
bath. After digestion, 5 µL of digestion mixture was loaded onto 2% agarose gel in
TAE buffer and restriction fragments were separated by gel electrophoresis. The
restriction patterns unique to each isolate were visualized under UV light staining
GelRed dye. Depending on the results, the isolates with different restriction patterns
were named as Isolate FKM-3, FKM-6, FKM-7, FKM-9 and FKM-11.
3.5.5.6. Sequencing and Phylogenetic Analysis
Bacterial 16S rRNA genes have nine “hypervariable regions” (V1 – V9) that provide
essential sequence diversity among different bacteria (Chakravorty et al. 2007).
88
Therefore 16S rRNA gene sequence provides valuable information for identification
of environmental isolates at genus and species level. 16S rRNA gene is relatively
short (1,500 bp) and this makes it faster and cheaper to sequence.
Following DNA extraction from environmental isolates, the extracted DNA was
used for amplification of the 16S rRNA gene using 27F and 1492R primer set (Lane
1991). The details of reaction mixture and PCR conditions are given in Section 3.5.2.
The PCR products were evaluated by gel electorophoresis (1.2% agarose gel, TAE
buffer, 80 V). 16S rRNA gene fragments were excised from agarose gel and
recovered using a Clean-up kit (Macherey-Nagel, GmbH & Co. KG, Germany). By
using 27F and 1492R primers, sequencing of purified 16S rDNA was performed by
BGI-Tech (BGI-Tech Corp., Denmark) using an ABI 3700 XL automatic sequencer.
The sequences were submitted to the Finch TV program to view sequence and
identify errors. The sequences were edited and assembled using DNAStar program
(DNASTAR, Madison, WI). The sequence results of isolates were compared and
analyzed for similarity using NCBI Blast tool (GenBank, NCBI). The 16S rRNA
gene sequences of isolates and the reported nonylphenol degraders were aligned
using Clustal X 2.0 program (Thompson et al. 1997). The neighbour-joining
phylogenetic tree was generated using CLC Free Workbench 6.8.4 program.
The 16S rRNA gene sequences determined in this study have been deposited in the
GenBank database (NCBI) with following accession numbers: KP143085 (FKM-3),
KP143086 (FKM-6), KP143087 (FKM-7), KP143088 (FKM-9), and KP143089
(FKM-11).
3.5.5.7. Biodegradation Experiments
3.5.5.7.1. Synthesis of Single NP Isomers
Technical NP contains a variety of branched alkyl chain isomers different in terms of
the nonyl moiety (C9H19) attached to the phenol ring (Thiele et al. 1997, Wheeler et
89
al. 1997, Thiele et al. 2004). It was reported that tNP has more than 100 isomers and
4-NP constitutes 90% of tNP mixture (Ieda et al. 2005, Guenther et al. 2006).
Therefore, during biodegradation experiments with tNP, each isomer produces its
own intermediates bringing about a complex and hard-to-read chromatogram with
many peaks. Also, findings of recent studies indicated that NP isomers differ in
estrogenic activity and biodegradability in the environment (Gabriel et al. 2008,
Eganhouse et al. 2009, Shan et al. 2011, Lu and Gan 2014). In light of these
findings, biodegradation studies with single isomers of tNP can help better
understanding of fate of NP in the environment and make easier to monitor possible
degradation products produced during biodegradation assays. Therefore, in this
study, biodegradation experiments were carried out with two different single isomers
of tNP: 4(3’,5’-dimethyl-3’-heptyl)phenol (p353NP) and 4(3’,6’-dimethyl-3’heptyl)phenol (p363NP).
The p353NP and p363NP isomers were synthesized by Friedel-Crafts alkylation
described by Vinken et al. (Vinken et al. 2002). For synthesis of these isomers, 3.01
g of phenol (32 mmol) and 2.25 g of nonanol (15.6 mmol of 3’,5’-dimethyl-3’heptanol or 3’,6’-dimethyl-3’-heptanol) were added into a dry two-necked flask
containing 300 mL of anhydrous petroleum ether (boiling range from 60 to 90oC).
The two-necked flask was connected with a reflux condenser. The solution was
heated to 50oC under argon gas. At 50oC, 2.6 mL of BF3-ether complex was
introduced into solution slowly and then the solution was mixed for 30 min at the
same temperature. Then reaction was stopped and 300 mL of cold water (including
ice pieces) was added. After stirring for further 30 min, the organic phase was
separated using separation funnel and organic phase washed more 5 times with same
amount of water. After last washing step, organic phase was dried over Na2SO4.
With the help of rotary evaporator (Heidolph Laborota 4000 Efficient, Heidolph
Instruments, Germany) petroleum ether was removed and the single isomer of NP
(p353NP or p363NP) was obtained.
90
The products were purified by column chromatography using n-hexane/ethyl acetate
mix (v/v, 13:1) to remove ortho- and dialkylated by-products. Column
chromatography was performed using thick wall-glass column filled with silica gel
(Merck Silica gel 60, particle size of 63-200 μm). Eluents were analyzed by thin
layer chromatographic (TLC) plates coated with silica gel (Merck Silica Gel 60
F254). The plates were visualized under UV-light. The solvent was removed. As a
final product, p353NP or p363NP was obtained as colorless and viscous structure.
Yields were calculated as 61.4 and 70.3% for p353NP and p363NP, respectively.
To check chemical structure of synthesized isomers, GC/MS and nuclear magnetic
resonance (NMR) spectra analyses were carried out. GC/MS analyses are depicted in
Figures 3.13 and 3.14.
1
H-NMR and 13C-NMR analyses were carried out in deuterated chloroform (CDCl3,
MagniSolv, Merck) on Bruker Avance DPX 400 NMR System. The NMR results are
given in Appendix C.
91
92
Figure 3.13. Chromatographic illustration of p363NP with GC/MS analysis.
93
Figure 3.14. Chromatographic illustration of p353NP with GC/MS analysis.
3.5.5.7.2. Growth and Biodegradation Assays
Biodegradation experiments were conducted in 100-mL Pyrex bottles containing 20
mL of MSM and 20 mg of NP compound (as sole carbon source) in each of them.
For inoculation, 1% of dense culture of each bacterial isolate grown in MSM
medium with NP compound (1 mg/mL) was used. This set-up was used for tNP,
p353NP and p363NP chemicals.
Due to low solubility of NP (4.9±0.4 mg/L at 25oC) in aqueous environment, most
of the nonylphenol added into the medium remained as droplets on the liquid
medium or on the walls of glass bottles. Therefore, in order to monitor the
concentration of nonylphenol in biodegradation experiments, two series of identical
culture bottles were prepared. In other words, 20 glass bottles (100 mL) containing
20 mL MSM and 20 mg NP compound were set in duplicate. For each analysis day,
two bottles were sacrificed. The mean value was calculated and used for
representation of data obtained.
As mentioned earlier, NP is not stable at high temperatures, therefore, nonylphenol
was added into autoclaved empty bottle as non-sterile dissolved in ethyl acetate.
The solvent was evaporated under sterile laminar flow and then sterile 20 mL MSM
medium was introduced into bottle. Each bottle with MSM and NP compound (1
mg/mL of tNP, p353NP or p363NP) was inoculated with 1% dense culture of
isolate of interest
grown in MSM-NP medium. As parallel, control bottles
(addition of NP compound, no inoculation) were also prepared to ensure that
degradation happens only by microbial activity (no photo-degradation or loss of
compound due to evaporation) and to detect if contamination occurs.
After inoculation, the culture bottles were placed on a rotary shaker at 120 rpm and
28±2oC. Each analysis day, 2 culture bottles were chosen and used for following
analyses: growth, extraction and determination of NP compound and degradation
products.
94
For bacterial growth determination, 1 mL sample was taken from each bottle and
OD550 was measured. Then, the culture bottles were used for aforementioned
analyses.
3.5.5.7.3. Extraction of NP Compounds
For extraction of NP compound and degradation products from liquid culture, ethyl
acetate (EtOAc) was used as extraction solvent. Dichloromethane (DCM) and
diethyl ether (DEE) were also tested to use for extraction but EtOAc gave better
recovery as 95.9% following the spike of 100 ppm tNP into MSM culture medium
when compared to recovery values of 91% for DEE and 94.2% for DCM.
The amount of NP compound remained in culture medium was determined by
calibration curve prepared with five-points (r2 > 0.99). The calibration curves used
in these analyses are given in Appendix A.
The details of biodegradation experiments and extraction studies are presented in
Chapter 7.
95
96
CHAPTER 4
DETERMINATION OF NONYLPHENOL COMPOUNDS IN
SLUDGE SOLIDS AND LIQUIDS: METHOD DEVELOPMENT
AND APPLICATION
4.1. Introduction
Increases in population and level of industrial development over time have led to
increased burden on treatment plants. This increase has been accompanied by high
sewage sludge production and this has brought the need of new sludge disposal
options. In recent years, the land application of sewage sludge for agricultural
purposes has become appealing option due to being cost-effective. Also, sludge has
become a good candidate to use as soil conditioner or fertilizer to improve soil
properties by contribution of nutritional substances and elements in its structure. In
time some concerns related to land application of sewage sludge have come out
with the possibility of introduction of pathogens, organic pollutants and heavy
metals into environment (Banat et al. 2000, La Guardia et al. 2000).
Surfactants are common organic pollutants of aquatic and soil environment.
Especially, nonylphenol polyethoxylates (NPnEO) are widely used nonionic
surfactants in formulations of household, industrial and agricultural products
(Birkett and Lester 2002). The common use of NPnEO as surfactants have led to
release of these chemicals into environment at high concentrations through the
discharge of industrial effluents and wastewater treatment plants (Ahel et al. 1994,
Petrovic et al. 2002, Fries and Püttmann 2003). This makes the understanding of
fate of NPnEO in discharged environmental compartments critical.
97
Once NPnEO enter the environment, about 60% of these compounds end up in the
aquatic environment. Biodegradation starts from polyethoxylate chain by removal
of ethoxy groups (CH3CH2O-) which are shortened up to small persistent byproducts more resistant to microbial degradation (Jobling and Sumpter 1993).
These
by-products
are
nonylphenol
diethoxylate
(NP2EO),
nonylphenol
monoethoxylate (NP1EO), nonylphenol (NP), nonylphenoxy acetic acid (NP1EC)
and nonylphenoxyethoxy acetic acid (NP2EC) (Giger et al. 1984, Ahel et al. 1994,
Ahel et al. 1994, La Guardia et al. 2000, Jonkers et al. 2001). While NPnEC are
mainly the characteristic degradation products of aerobic biodegradation, NP and
short chain NP1-2EOs are dominant under anaerobic environments. Unfortunately,
these degradation products being more persistent, lipophilic and toxic than the
parent compounds, has increased concerns about use of NPnEO (Ekelund et al.
1993, Ahel et al. 1994, Lee et al. 1997).
It has been reported that especially NP is more toxic to aquatic species and higher
life forms (Comber et al. 1993, Staples et al. 1998). Soto et al. (1991) observed that
the presence of NP induces breast tumor cell proliferation (Soto et al. 1991).
Besides that, it has been proven that NP acts as endocrine disrupter (Jobling and
Sumpter 1993, Toppari et al. 1996). NP is capable of mimicking the 17-β-estradiol
hormone which plays important role in development of sex characteristics and
functioning of sex organs (Jobling and Sumpter 1993, Sharpe et al. 1995, Toppari
et al. 1996, Gray and Metcalfe 1997). The studies revealed that exposure to NP lead
to feminization of reproductive organs of male fishes (Jobling et al. 1996) (Aguayo
et al. 2004, Lavado et al. 2006) and interfere with the estrous cycle and pubertal
onset in rats (Laws et al. 2000, Kim et al. 2002).
With evidence for detrimental effect of NP compounds for environment and life
forms, NPnEO have been listed as hazardous substances by the OSPAR
Commission (Commission 1998). Then NP has been accepted as priority hazardous
substance in the Water Framework Directive (Directive 2000/60/EC, 2000) for
surface waters. Furthermore, the European Union passed an amendment to regulate
their uses (Directive 2003/53/EC, 2003). The awareness of adverse effects of NP
98
compounds has also led to increase in concerns about the land application of sludge
for agricultural purposes. Due to ability of NP compounds to reach environment
through waste streams and accumulation of persistent degradation products in
sewage sludge, European Union set a limit value for the sum of NP, NP1EO and
NP2EO (NPE = NP + NP1EO + NP2EO) as 50 mg/kg dry mass (dm) in Working
Document on Sludge, 3rd Draft (2000) for the agricultural use of sludge. In the EU
accession period of Turkey, the limit value for NPE in sludge was set as 50 mg/kgdm in Regulation on the Land Use of Domestic and Urban Sludges (2010). In the
USA, Environmental Protection Agency (EPA) have set a criteria which suggests
that nonylphenol concentration should not exceed 6.6 µg/L in fresh water and 1.7
µg/L in saltwater (Brooke and Thursby 2005, David et al. 2009). Lately, the EPA
proposed a Significant New Use Rule (SNUR) for 15 related chemical substances
commonly known as nonylphenols (NP) and nonylphenol ethoxylates (NPEOs) on
October 1, 2014 to regulate the use of these chemicals.
Setting up a limit value for NPEs prior to land application of sewage sludge and
the increase in concerns about the level of these compounds in surface waters have
brought a need for a reliable method for extraction and simultaneous detection of
these compounds in sewage sludge and surface waters. Some methods have been
reported in literature for extraction of NPEs from environmental samples. The
methods used for extraction of target NP compounds from solid phases (sludge,
sediment, soil) are reported as soxhlet extraction (Pryor et al. 2002) accelerated
solvent extraction (La Guardia et al. 2001, Meesters and Schröder 2002),
microwave assisted extraction (Fountoulakis et al. 2005), ultrasonication assisted
extraction
(Hashimoto
et
al.
2005,
Stasinakis
et
al.
2010),
ultrasonication+mechanical shaking (Abad et al. 2005, Aparicio et al. 2007),
pressurized liquid extraction (Andreu et al. 2007, Minamiyama et al. 2008) and
supercritical fluid extraction (Lin et al. 1999, Minamiyama et al. 2008). For
aqueous samples (water, wastewater, surface waters) liquid - liquid extraction
(Bennie et al. 1997, Conn et al. 2006) solid phase extraction (Fries and Püttmann
2003, Dıá z et al. 2004, Lian et al. 2009, Nie et al. 2009) (Fries and Puttmann 2003;
99
Diaz et al., 2004; Nie et al., 2009; Lian et al., 2009) and solid phase microextraction
(Kawaguchi et al. 2004, Basheer et al. 2005) methods have been used. For detection
and quantification of these compounds, GC/MS, GC/MS/MS, LC/MS and
LC/MS/MS have been used in these studies as an analytical technique. Variation in
extraction methods and analytical techniques are also accompanied by diverseness
in extraction procedure (extraction solvent, extraction time) and analytical
application (oven program, LOD, LOQ).
Due to the lack of a standard method suggested by authorities and variation in the
proposed methods in literature, this study has aimed to compare prominent methods
reported in literature and optimize a reproducible, reliable and easily applicable
method to identify and quantify NP, NP1EO, NP2EO and NP1EC compounds in
sewage sludge and effluent of waste activated sludge (WAS) samples. The
ultrasonic extraction and solid phase extraction methods applied for solid and
aqueous samples, respectively and then following derivatization procedure the
extracts were analyzed by GC/MS. Also, the proposed methods were applied for
the determination and seasonal monitoring of NP, NP1EO and NP2EO compounds
in dewatered sewage sludge and waste activated sludge samples taken from the
central wastewater treatment plant in Ankara in 2012. The proposed methods could
be used for quality monitoring of sewage sludge prior to land application to
compensate related regulations of European Directive and Turkish government.
4.2. Materials and Methods
4.2.1. Sludge samples
Dewatered sludge samples were supplied from Central Wastewater Treatment Plant
(WWTP), Ankara monthly between February 2012 and December 2012. The
treatment plant is the biggest treatment plant in Ankara and the second biggest in
Turkey with a current flow rate of 765,000 m3/day (ASKI, 2015). Following
sampling, sludge samples were freeze-dried to remove moisture and sieved through
1 mm sieve to obtain homogenous samples.
100
Pre-cleaned sludge samples (by sonication and soxhlet extraction to eliminate
matrix effects) were used for initial method trials and development. These sludge
samples were also taken from Central WWTP, Ankara.
4.2.2. Chemicals
The standard solutions of NP, NP1EO and NP2EO (analytical standard, 5 µg/mL in
acetone) were supplied from Fluka (Sigma Aldrich Co. LLC, USA). The standard
solution of NP1EC (10 ng/µL in acetone) was purchased from Dr. Ehrenstorfer
GmbH, Germany. The standard solution of 4-n-NP (10 ng/µL in cyclohexane) was
supplied from Dr. Ehrenstorfer (Augsburg, Germany) and used as surrogate in
extraction studies.
Derivatization agents MSTFA (N-methyl-N-trimethylsilyl) trifluoroacetamide and
BSTFA (N,O bis (trimethylsilyl)trifluoroacetamide) and the catalysts pyridine and
TMCS (trimethylchlorosilane) were purchased from Sigma Aldrich, LLC, USA.
BSTFA+ TMCS (99:1), kit was obtained from Supelco Analytical (Sigma Aldrich
Co. LLC, USA). The boron trifluoride (BF3) in methanol (20%) was supplied from
Merck KGaA, Germany.
The tC-18 cartridges (SEP-PAK Vac C18, 6 cc/500 mg) were supplied from Waters
Co, USA. Sodium sulfate, fine powder copper and solvents (acetone, methanol,
hexane, all GC-MS grade) were supplied from Merck KGaA, Germany. Petroleum
ether (40-60oC) was also purchased from Merck KGaA, Germany.
Standard solutions were prepared in acetone at analysis day and stored in amber
glass vials at -18°C.
4.2.3. Optimization of Derivatization Method for NP Compounds
For GC/MS analysis, compounds with functional groups (OH, –COOH, =NH, –
NH2, and –SH) are major concern due to their ability to form hydrogen bonds in
between compounds. This is accompanied by decrease in volatility and thermal
stability of the compounds resulting in low detectability (Schummer et al. 2009). In
101
order to overcome this problem, derivatization is applied to produce volatile and
thermally stable derivatives of compounds for GC/MS analysis. For NP, NP1EO
and NP2EO (NPE compounds), to derivatize of the hydroxyl groups in their
structure, silylation was applied as a derivatization method.
BSTFA and MSTFA have been used widely as silylation agents for derivatization
of organic compounds as reported in literature (Morley et al. 2005, Gatidou et al.
2007, Hao et al. 2007, Briciu et al. 2009, Adamusova et al. 2014). Therefore,
preliminary experiments were conducted in order to decide which silylation reagent
yields higher peak areas with better quality between BSTFA and MSTFA. In order
to develop a reliable derivatization procedure which enables the compounds to be
detected efficiently, BSTFA and MSTFA were tested by using two different
catalysts pyridine and TMCS (trimethylchlorosilane). In silylation process, active
hydrogens in NP compounds are replaced with TMS (trimethylsilyl) groups of
BSTFA (and MSTFA) and increase the reactivity of the derivatization agent.
BSTFA or MSTFA can also act as solvent so no solvent is needed to solubilize the
compounds during derivatization procedure.
For comparison studies for BSTFA and MSTFA, a 1 mL sample of NPE
compounds in the vials was dried under gentle stream of N2 gas and then 50 µL of
derivatization agent was added. Then the dry residues were derivatized for 20 min
at 65oC. With the cooling of the vials at room temperature, the derivatives of NP
compounds were injected to GC/MS after vortexing the vials for 1 min. Depending
on obtained results, BSTFA yielded better chromatographic peaks with higher
abundance and quality compared to MSTFA. After this point, derivatization
experiments continued with BSTFA for determination of optimal conditions for
derivatization procedure. Observation of some underivatized compound signals in
chromatograms made catalyst addition essential for complete derivatization.
Derivatization times for NPE compounds vary widely in studies reported in
literature. The time is necessary for dissolution of compounds in the derivatization
reagent and completion of reaction. Also, the compounds with poor solubility may
102
need heat for derivatization reaction to proceed. Due to poor solubilities (around 5
mg/L at 25oC) of NPEs, heating becomes necessary step during derivatization
process. In light of these requirements for NPE compounds, experimental trials
were performed with NP residues (evaporated to dryness from an acetone stock
solution of target NP compounds) at different temperature (65, 70, 75 and 90oC)
and time periods (30, 60, 90, 120 min and 5 hours) in the presence of 50 µL
BSTFA with a combination of pyridine or TMCS as a catalyst. With application of
these combinations to determine optimal conditions for derivatization reaction, the
results indicated that derivatization reactions yield better results at 70oC with 30
min-derivatization time. Also, the obtained results revealed that using of TMCS as
a catalyst improved the efficiency of derivatization reaction and yielded narrower
and higher peaks. On the other hand, pyridine led to a noisy background and
problems in peaks structure. It has been reported that the use of pyridine as a
catalyst promote the formation of secondary products and chromatographic
anomalies (Dalgliesh et al. 1966) so the observed problems in this study with the
use of pyridine as a catalyst coincide with the study of Dalgliesh and co-workers
(Dalgliesh et al. 1966). In the light of these experimental findings, TMCS was used
together with BSTFA in derivatization reactions of NPE compounds in following
studies.
The trials to determine optimal working volume ratio of BSTFA and TMCS for
effective derivatization revealed that the increase in percentage of TMCS lead to
decrease in abundance of NP peaks as presented in Figures 4.1 and 4.2. 1% addition
of TMCS into derivatization reaction including BSTFA was considered enough to
complete derivatization of all NPE compounds depending on obtained results.
When the results given in Figure 4.2 and 4.3 related to working volume
experiments were taken into consideration, addition of higher volume of
BSTFA+TMCS than 100 µL was accompanied by decrease in abundance of
chromatographic peak areas. With these results, to provide enough volume for
solubilization of NPE compounds, 100 µL working volume of BSTFA and TMCS
was chosen as reaction volume for derivatization process.
103
Figure 4.1. GC/MS chromatograms of NP silyl-derivatized in 100 μL of
BSTFA+10%TMCS and BSTFA+33%TMCS
Figure 4.2. GC/MS chromatograms of NP1EO silyl-derivatized in 100 and 200 μL
of BSTFA+1%TMCS and BSTFA+10%TMCS
104
Figure 4.3. GC/MS chromatogram of NP, NP1EO and NP2EO silyl-derivatized in
100 and 200 μL of BSTFA+1%TMCS
Considering the results of all these experimental studies conducted to optimize the
procedure, NPE compounds were derivatized in 100 µL of BSTFA + 1% TMCS at
70oC for 30 min.
Short-chain nonylphenoxy acetic acid (NP1EC) and nonylphenoxy ethoxy acetic
acids (NP2EC) are among the most frequently detected compounds in the aquatic
environment and estrogenic to wildlife at low concentrations (Argese et al. 1994,
Berryman et al. 2004, Williams et al. 2009). Therefore detection of these chemicals
in aqueous samples becomes important. For determination of NP1EC, a
derivatization method has been optimized by comparison of three different
derivatization procedures reported in literature. These procedures rely on
conversion of carboxylic acids into esters by alkylation of functional groups. In first
procedure, NP1EC was derivatized by conversion of acidic carboxylic groups into
methyl esters in the presence of dimethyl sulfate/NaOH as reported in (Díaz et al.
2002). In another tested derivatization procedure reported by (Ding and Tzing
1998), carboxylic acids of NP1EC was converted into propyl esters with npropanol–acetyl chloride (9:1, v/v) reagent. In last applied procedure, NP1EC was
105
derivatized in the presence of 14% BF3 in methanol solution for formation of
methyl esters as suggested by Lee et al. (Lee et al. 1997).
The derivatization experiments for NP1EC revealed that the derivatization method
suggested by Lee et al. (1997) yielded in satisfactory results when NP1EC
derivatives were subjected to GC/MS analysis (Figure 4.6) (Lee et al. 1997). The
application of procedure proposed by Ding and Tzing also yielded chromatograms
similar to their results but with much lower abundance and some impurities (Figure
4.5) (Ding and Tzing 1998). Methylation of NP1EC compounds as a way suggested
by Diaz et al. (2002) did not produce any remarkable chromatogram peaks
indicating methyl esters of NP1EC (Figure 4.4).
Figure 4.4. a) A GC/MS chromatogram of methylated NP1EC given in the study of
Diaz et al. (Díaz et al. 2002) b) A GC/MS chromatogram of methylated NP1EC
following derivatization in our lab using the method of Diaz et al. (Díaz et al.
2002).
106
Figure 4.5. a) A GC/MS chromatogram of propyl esters of NP1EC given in the
study of Ding and Tzing (Ding and Tzing 1998) b) A GC/MS chromatogram of
propyl esters of NP1EC following derivatization in our lab using the method of
Ding and Tzing (Ding and Tzing 1998).
In order to improve derivatization method proposed by Lee et al. (1997), some
modifications were introduced into the procedure. 20% BF3 (in methanol) reagent
was used instead of 14% BF3. Also, in order to generate and apply the same mixing
conditions for extraction of NP1EC derivatives into petroleum ether in all
applications, a mechanical shaking step was introduced. This mixing was achieved
by shaking of tubes at 400 rpm for 1 min instead of hand-shaking. The NP1EC
derivatives were reconstituted in 1 mL of hexane instead of isooctane as was used
in the study of Lee et al. (2002) for GC/MS analysis (Lee and Peart 2002).
107
Figure 4.6. a) A GC/MS chromatogram of methyl esters of NP1EC given in the
study of Lee et al. (Lee et al. 1997) b) GC/MS chromatogram of methyl esters of
NP1EC obtained in our study following derivatization method suggested by Lee et
al. (Lee et al. 1997).
After determination of derivatization reagents and ideal conditions to be applied on
extracts obtained from the sludge samples, the developed procedure for NPE
compounds was applied as follows: 1 mL extract was taken and evaporated under
gentle nitrogen gas stream until dryness. 100 µl BSTFA-1%TMCS reagent was
added and vials were capped. Following vortex-stirring for 1 min, the vials were
placed into oven at 70oC for 30 min for derivatization reaction. After 30 min, vials
were cooled at room temperature and then vortexed for 1 min. The derivatives were
transferred into injection vial and then injected to GC/MS.
108
For NP1EC, the modified derivatization method suggested by Lee et.al (1997) was
applied as follows; 1 mL NP1EC solution was evaporated to 200 μL using N2 gas
stream and then 2 mL 20% Boron trifluoride (BF3) solution was added. The vials
were kept at 85oC for 30 min in oven. At the end of 30-min, the vials were taken
and kept at room temperature for cooling. Then their contents were evaporated to
300 μL using N2 gas stream and placed into mechanical shaker at 400 rpm for 1
min. Following shaking step, 2.5 mL of double distilled water (ddH2O) was added
and mixed. Then 2 mL petroleum ether was added to each vial to extract the
methylated derivatives. This step was repeated for 3 times and then the eluate was
passed through Na2SO4 column. The eluate was evaporated to complete dryness by
N2 gas. Then, 1 mL of hexane was added to the vials and they were vortexed for 1
min. Each sample was transferred to injection vial and then sample was injected
into the GC/MS.
4.2.4. Optimization of GC/MS Method for detection of NP compounds
In order to identify and quantify NP compounds, different methods at different
oven programs were tried on a 7890A Agilent gas chromatograph coupled to a
5975C Agilent mass spectrometer with Triple-Axis. As a start point, methods used
in literature for NPE compounds given in Table 4.1 were tested to obtain a method
with higher peak quality, reproducibility and repeatability.
109
110
Dimethylsulfate/NaOH
BSTFA
NP, NP 1-2EO
NP
No derivatization
No derivatization
NP
NP
BSTFA+10%TMCS
Derivatization
NP, NP 1-4EO
Compound
60°C for 1 min,
60°C to 280°C at 10°C/min,
280°C held for 7 min
MS transfer line temp.: 280oC
Ion source temp.: 230oC
50°C for 2 min,
50°C to 100 C at 20°C/min,
100°C to 200°C at 10°C/min,
200°C to 290°C at 20°C/min,
290°C hold for 2 min
MS transfer line temp.: nm
Ion source temp.: nm
70°C for 3 min,
70°C to 160°C at 20°C/min,
160°C to 285°C at 10°C/min,
285°C held for 7 min
MS transfer line temp.: 290oC
Ion source temp.: 200oC
70 °C for 2 min,
70 to 180°C at 30°C/min,
180°C to 200°C at 2°C/min,
200°C to 310°C at 30°C/min,
310°C hold for 10 min
MS transfer line temp.: 310oC
Ion source temp.: nm
40°C for 1 min,
40 to 300oC at 6oC/min,
300oC hold for 5 min
MS transfer line temp.: 280oC
Ion source temp.:250oC
Oven Program
DB-1
(30mx0.25µmx.0.32mm)
DB5MS
(30mx0.25µmx0.25mm)
Shimadzu GC-17A/MS QP-5000
HP GC 5890/MS 5973
DB5MS
(30mx0.25µmx0.25mm)
DB5MS
(30mx0.25µmx0.25mm)
HP GC 5890/MS 5972A
Fisons GC 8060/MS MD 800
HP5MS
(25mx0.33µmx0.2mm)
Column
HP GC 5890/MS 5970
Instrument
Table 4.1. Details of GC/MS and oven programs of studies applied in preliminary experiments
(Gibson et al. 2005)
(Li et al. 2004)
(Díaz et al. 2002)
(Isobe et al. 2001)
(Barber et al. 2000)
Reference
111
No derivatization
nm: not mentioned in the study
NP
BSTFA+TMCS
No derivatization
NP, NP 1-2EO
1-2EO,
BSTFA+pyridine
NP, NP 1-2EO
NP,
NP
NPnEOs
BSTFA and Sigma-Sil A
Derivatization
NP, NP 1-2EO
Compound
HP GC 6890/MS 5975i
HP5890 series II/Fisons MD800
50oC for 1 min,
50 to 200oC at 20oC/min,
200oC held for 2 min,
200 to 235oC at 5oC/min,
235oC held for 5 min,
235 to 280oC at 25oC/min,
280oC held for 5 min
MS transfer line temp.: nm
Ion source temp.: nm
100oC for 1 min,
100 to 280oC at 10oC/min,
280oC held for 3 min
MS transfer line temp.: 280oC
Ion source temp.: nm
Shimadzu, GC-MS-QP2010
HP GC 5890 Series II/MS 5971
80oC for 1 min,
80 to 220oC at 15oC/min,
220 to 280oC at 5oC/min
MS transfer line temp.: 280oC
Ion source temp.:180oC
130°C for 0 min,
130°C to 280°C at 5oC/min,
280°C held for 20 min
MS transfer line temp.: nm
Ion source temp.:200oC
HP GC 6890/MS 5973
Instrument
50°C for 0 min,
50°C to 120°C at 110°C/min,
120°C to 300°C at 30°C/min,
300 °C held for 5 min
MS transfer line temp.: 250oC
Ion source temp.: 250oC
Oven Program
ZB-5
(30mx0.25µmx0.25mm)
nm
DB5MS
(30 mx0.25µmx0.32mm)
DB5MS
(60 mx0.25µmx0.32mm)
RTX5
(20mx0.1µmx0.1mm)
Column
Reference
(Richter et al. 2009)
(Lian et al. 2009)
(Lu et al. 2008)
(Gatidou et al. 2007)
(Benanou et al. 2002)
Table 4.1. Details of GC/MS and oven programs of studies applied in preliminary experiments (continued)
Our application of GC/MS programs revealed different results when chromatogram
peaks of NP compounds were compared to the ones reported in literature.
The obtained total ion chromatograms following application of GC/MS oven
programs reported in studies are presented in Figure 4.7 to Figure 4.13. The oven
program of Isobe et al. did not result in any distinguisable chromatography peaks
for NP compounds (Isobe et al. 2001). On the other hand, the other methods tested
yielded peaks for NP compounds. The results are illustrated in Figure 4.9-4.12 for
NP.
Figure 4.7. Comparison of total ion chromatograms of NP obtained from
application of oven programs given in studies of Lian et al. (Lian et al. 2009),
Gatidou et al. (Gatidou et al. 2007), and Gibson et al. (Gibson et al. 2005).
112
Figure 4.8. Comparison of total ion chromatograms of NP obtained from
application of oven programs given in studies of Richter et al. (Richter et al. 2009)
and Lu et al. (Lu et al. 2008).
Figure 4.9. Comparison of total ion chromatograms of NP obtained from
application of oven programs given in studies of Benanou et al. (2007), Donghao et
al. (2004) and Barber et al. (2000).
113
Figure 4.10. Total ion chromatogram of NP obtained from application of oven
program given in the study of Diaz et al. (2000)
With initial trials for NP, the oven programs yielding higher abundance and peak
quality were determined as Gatidou et al. (2007) and Diaz et al. (2002) (Figures 4.7
and 4.10). The difference in the performance of these oven programs could be
caused by the variation in parent compound, derivatization procedure applied prior
to GC/MS analysis, instrument feature, column type (dimension, stationary phase)
and operational conditions. These programs were applied for identification of
NP1EO and NP2EO. The total ion chromatograms obtained for NP1EO are
demonstrated in Figure 4.11 and Figure 4.12.
114
Figure 4.11. Total ion chromatogram of NP1EO obtained from application of oven
program given in the study of Diaz et al. (2000).
Figure 4.12. Total ion chromatogram of NP1EO obtained from application of oven
program given in the study of Gatidou et al. (2007).
115
Figure 4.13. Total ion chromatogram of NP2EO obtained from application of oven
program given in the study of Diaz et al. (2002)
When Figures 4.7 and 4.12 are examined, retention times of NP and NP1EO
obtained with Gatidou et al. (2007) method were very close each other. So the use
of GC/MS oven program of this study was not reasonable for further analysis. The
application of GC/MS program in the study of Diaz et al. (2002) yielded lower
abundance and problem in peak quality of NP2EO compound (Figure 4.13). Based
on these two studies, seven different oven programs were developed and tested in
order to optimize a GC/MS method for simultaneous detection of NPE compounds
with higher peak abundance, resolution and quality. The tested oven programs for
the simultaneous detection of NP compounds and operation conditions are given in
Table 4.2.
116
Table 4.2. GC/MS methods tested with different oven programs for NPE
compounds
Oven Program
Method
Carrier Gas and Flow
MS-Source/MSInterface/Quadrupole Temperature
He, 0.9 mL/min
230°C/280°C/150°C
He, 0.9 mL/min
230°C/280°C/150°C
He, 0.9 mL/min
230°C/280°C/150°C
He, 0.9 mL/min
230°C/280°C/150°C
He, 1 mL/min
230°C/280°C/150°C
He, 1.2 mL/min
230°C/280°C/150°C
He, 1 mL/min
230°C/280°C/150°C
100°C for 1 min,
100°C to 220oC at 10oC/min,
1
220oC to 280oC at 5oC/min,
280oC held for 5 min
40°C for 1 min,
2
100°C to 220oC at 15oC/min,
o
o
o
220 C to 280 C at 5 C/min
80°C for 2 min,
80°C to 220oC at 10oC/min,
3
220oC held for 1 min
220oC to 280oC at 5oC/min
280oC held for 5 min
80°C for 1 min,
80°C to 150oC at 10oC/min,
4
150oC to 200oC at 5oC/min,
o
o
o
200 C to 280 C at 10 C/min
280oC held for 4 min
100°C for 1 min,
100°C to 220oC at 10oC/min,
5
220oC to 280oC at 5oC/min,
280oC held for 5 min
100°C for 1 min,
100°C to 220oC at 10oC/min,
6
220oC to 280oC at 5oC/min,
280oC held for 5 min
100oC for 5 min,
100oC to 160oC at 25oC/min,
160oC to 260oC at 10oC/min,
7
260oC held for 5 min,
260oC to 285oC at 35oC/min,
285oC held for 7 min
When the GC/MS methods were evaluated based on yielded peak area and quality
of chromatograms, Method-1, Method-5 and Method-6 gave similar results due to
the application of same oven program. For these methods, gas flow was the variable
parameter and it can be said that change in gas flow did not result in any
117
remarkable change in the quality and abundance of peaks except for retention time.
As can be seen from Figure 4.14, this oven program gave better results compared to
Method-2, Method-3 and Method-4. Method-2 yielded peaks with lowest quality
when total ion chromatograms in Figure 4.14 were examined. Although Method-3
and Method-4 presented good-shaped peaks, they produced lower abundances
compared to Method-1. When Method-1 was compared with Method-7, it was seen
that the application of Method-7 gave satisfactory and better results with higher
abundance and symmetrical-narrow shaped peaks for the identification of NP
(Figure 4.15). Method 1 and Method 7 were further tested for detection of NP1EO
and NP2EO compounds. Due to inability for detection of low levels of target
compounds and obtaining of lower abundances with Method-1, Method-7 was
applied as the GC/MS program for identification and quantification of NPE
compounds for further analysis in this study.
Figure 4.14. Comparison of total ion chromatograms of NP obtained from the
application of different GC/MS methods
118
Figure 4.15. Comparison of total ion chromatograms of NP obtained from
application of GC/MS oven program of Method 1 and Method 7
Several GC/MS methods were also tested for derivatized NP1EC. The details of
applied methods are presented in Table 4.3. The derivatization method proposed by
Lee et al. (1997) was applied for NP1EC compounds with minor changes. In order
to compare results with the study of Lee and co-workers (1997), same GC/MS
program was followed for GC/MS analysis.
The GC/MS oven program used for NP1EC analysis in the study of Diaz et al.
(2002) did not yield any peak indicating the presence of NP1EC. When GC/MS
oven program used in the study of Ding and Tzing (1998) was tested, a cluster of
peaks was detected around the same retention time represented in this study,
however with much lower yield. Application of GC/MS oven program proposed by
Lee et al. (1997) resulted in better peak quality and higher abundance (Figure 4.16).
With these results, the GC/MS oven program given in Table 4.3 developed by Lee
and co-workers (1997) was applied for GC/MS analysis of NP1EC in this study.
119
120
1 µL
Splitless
250oC
280oC
230oC
150oC
SIM
100 C for 5 min,
8.5oC/min to 280oC,
280oC held for 15 min
1 µL
Splitless
250oC
280oC
230oC
150oC
SIM
70 C for 3 min,
20oC/min to 160oC
10oC/min to 285oC
285oC held for 7 min
27 min
Injection volume
Injection Mode
Injection
Temperature
MS Interface
Temperature
MS Source
Temperature
MS Quadrupole
Temperature
MS Mode
Duration
Oven Program
Helium (1 mL/min)
Helium (1.2 mL/min)
Carrier Gas (Flow)
41.18 min
o
HP-5MS
HP-5MS
Column
o
Ding and Tzing, 1998
Diaz et al., 2002
Table 4.3. GC/MS methods experimented for NP1EC analysis
SIM
70 C for 1 min,
30oC/min to 160oC,
5oC/min to 290oC,
290oC held for 5 min
35 min
o
150oC
230oC
280oC
250oC
Splitless
1 µL
Helium (1.2 mL/min)
HP-5MS
Lee et al., 1997
Figure 4.16. Total ion chromatogram of NP1EC obtained from application of oven
program given in the study of Lee et al. (1997)
As overall, for identification and quantification of NP, NP1EO and NP2EO a sample
volume of 1 µl was injected in splitless mode to GC/MS. An inlet temperature was
kept at 280oC. The oven program for NPE compounds, after preliminary studies, was
applied as follows: 100 °C for 5 min, 25°C/min to 160°C, 10°C/min to 260°C,
260°C for 5 min, 35°C/min to 285°C and 285°C for 7 min. For NP1EC, after trials,
the oven program was chosen as follows: the initial temperature was 70°C and
increased to 160°C at 30°C/min and then reached a final temperature of 290°C at a
rate of 5°C/min. For all compounds, MS source and quadrupole temperatures were
kept at 230°C and 150°C, respectively. The mass spectrometer was operated in the
electron impact ionization (EI) mode and the energy of the electrons was kept at 70
eV.
For qualitative analysis, GC/MS SCAN mode was used by monitoring the mass
range from 50 to 600. After identification of m/z values for each compound,
quantitative analyses were carried out using selective ion monitoring (SIM) mode in
GC/MS. The retention times (Rt) and the ions used for identification and
quantification of each NP compound were determined. The m/z values used for
121
derivatives of NP compounds are listed in Table 4.4. Being composed of different
isomers, NP, NP1EO, NP2EO and NP1EC yielded abundant peaks in
chromatograms after GC/MS analysis. Therefore, sum of the peak areas was used in
quantitative analysis to represent each compound’s abundance (Figures 4.17 and
4.18).
Table 4.4. Retention times and target ions of the NP compounds
Compound
Name
Retention time
(min)
Target ions
(m/z)
Quantification
4-n-NP
14.36
179, 292
Single peak
NP
12.23-12.93
179, 193, 207, 235, 277
Sum of 9 peaks
NP1EO
15.39-16.10
251, 265, 279, 293, 307
Sum of 11 peaks
NP2EO
17.91-18.79
281, 295, 309, 323
Sum of 9 peaks
NP1EC
16.40-17.20
193, 207, 221, 292
Sum of 7 peaks
Figure 4.17. Example for chromatographic view of GC/MS analysis for NP
compounds with application of optimized GC/MS oven program
122
Figure 4.18. Example for chromatographic view of GC/MS analysis of NP1EC
4.2.5. Optimization of Extraction Methods for NP compounds
4.2.5.1. Extraction from Solid Samples
For efficient and accurate extraction of NP compounds from solid samples, two
different extraction methods, sonication-assisted extraction and extraction by
mechanical shaking were tested. Sonication-assisted extraction method is generally
reported with high extraction efficiency by achieving a higher yield of target
compounds in shorter time with less solvent consumption and lower temperatures
compared to other alternatives (Castillo et al. 2000, Petrovic and Barceló 2000,
Fountoulakis et al. 2005, Aparicio et al. 2007). Extraction by mechanical shaking is
also a commonly used extraction method due to ease and simplicity (Spack et al.
1998, Schwab et al. 1999, Lu et al. 2008).
For each method, preliminary studies were carried out to decide on an ideal solvent
to use during extraction from solid samples. For this purpose, acetone, methanol,
acetone:methanol (1:1) and hexane were examined for their extraction efficiency. As
can be seen in Figure 4.19, hexane yielded lowest peak abundance when used for the
123
extraction of NPE compounds. On the other hand, acetone gave better
chromatograms with higher abundance and better peak quality. In the highlight of
these preliminary results, hexane was eliminated from further extraction studies.
Then extraction of NPE compounds from sludge samples following an 8 minsonication-assisted extraction with three different solvents (acetone, methanol,
acetone:methanol (1:1)) revealed that acetone was the most appropriate solvent for
the extraction of NP compounds with satisfactory recovery values. The average
recoveries (%) obtained for each compound are presented in Table 4.5.
Figure 4.19. Total ion chromatograms obtained for the extraction of target
compounds from solid samples using different solvents
124
Table 4.5. Effect of extraction solvents on the mean recoveries (%) of target
compounds in sludge samples obtained from 8-min sonication-assisted extraction
NP
Acetone
82.3±16.7
Methanol
184.9±3.0
Acetone:Methanol (1:1)
86.9±6.8
NP1EO
125±5.9
283.4±2.7
138.7±3.2
NP2EO
127.7±24.6
273.7±1.3
153.1±2.0
With the selection of proper extraction solvent, sonication-assisted extraction
method and extraction by shaking were evaluated with regards to their extraction
efficiency. Five different sonication times (3, 5, 10, 20, and 30 min) were studied to
determine the effect of time on extraction efficiency. For shaking-extraction method,
mechanical shaker was used and three different (16 h, 24 h, and 48 h) shaking times
were tested based on some previous studies (Kaya et al. 2012). Also two different
combination of these two methods (1 min-sonication+16 h-shaking and 1 minsonication+24 h-shaking) were experimented. In these extraction experiments,
acetone was used as extraction solvent.
As can be seen in Figure 4.20, better recoveries were obtained with 5-min sonication
of sludge samples. Three-minute sonication also yielded good recovery efficiencies
except for NP2EO. The recovery for NP2EO was determined as 174.1±3.8 following
3-min sonication. Despite the achievement of satisfactory recoveries for the other
three compounds, this sonication time was eliminated. With an increase in sonication
time beyond 5 min, a deterioration was observed in extraction efficiencies. It is
possible that excessive sonication may degrade the NP compounds in the sample and
lead to the worsening of recoveries. It has been reported that longer sonication times
lead to the decomposition of organic compounds due to heating caused by the effects
of ultrasound waves (Stephens et al. 1994, Santos et al. 2009, Lau et al. 2010).
Following 30-min sonication, NP2EO was not detected and NP1EO was detected
with much lower abundance. This observation supported the decomposition of NP
compounds by losing their ethoxylate groups leading to no detection or lower
detection of higher ethoxylated compounds.
125
250
4-n-NP
NP
NP1EO
NP2EO
(%)(%)
Recovery
Recovery
200
150
100
50
0
3 min
5 min
10 min
20 min
30 min
Sonication Time
Figure 4.20. Mean recoveries (%) of NPE compounds obtained for sonicationassisted extraction method at different sonication times with acetone
When mechanical shaking was applied for extraction of NP compounds from sludge
solids, it was seen that 16h-time period did not yield sufficient recovery for NP2EO
(Figure 4.21). When extraction time was increased to 24 h, satisfactory rcoveries of
each compound was observed except for NP2EO. The recovery values following
application of 24h-mechanical shaking for all compounds were obtained as follows:
118.20±24.90 for 4-n-NP, 109.90±17.30 for NP, 125.20±30.80 for NP1EO and
183.10±30.00 for NP2EO. The further increase of extraction time to 48h led to
deterioration in extraction of NP. When the recoveries of NP1EO and NP2EO
compared to the values of 24h-extraction, an improvement was observed for these
compounds. Due to lack of consistency between recoveries and large variations, it
was difficult to claim any extraction time as ideal for this method. The trials with
combination of 1 min-sonication with 16h and 24h mechanical shaking process
126
indicated that no consistency can be obtained for the extraction of different NP
compounds. As a result, it was dedicated that mechanical shaking extraction method
did not yield consistent and satisfactory recovery values when all the NP compounds
considered together compared to sonication-assisted extraction method. Also
mechanical shaking is time-consuming method over sonication.
300
200
Recovery (%)
Recovery (%)
250
4-n-NP
NP
NP1EO
NP2EO
150
100
50
0
16 h
24 h
48 h
1 min+16 h 1 min+24 h
Time
Figure 4.21. Mean recoveries (%) of NPE compounds for extraction by mechanical
shaking extraction method at different extraction times and combination of
extraction methods in the presence of acetone
As discussed earlier, sonication of sludge samples beyond 5-min caused
decomposition of NP compounds especially NP2EO. Therefore when the average
recoveries (%) obtained for each compound following 5-min-sonication were
examined (Table 4.6; 101.48±3.21 for 4-n-NP, 92.88±3.43 for NP, 134.62±5.78 for
127
NP1EO, 135.25±2.71 for NP2EO and 97.5±2.45 for NP1EC) it yielded the most
repeatable and reasonable results. USEPA suggested a recovery range of 70-130%
for most organic compounds in Method 8000C (USEPA; 2003), whereas other
studies accepted different recoveries such as 90-140% (Arditsoglou and Voutsa
2008) and 60-150% (Lian et al. 2009) for NP compounds. Based on the recoveries
obtained for NP compounds in literature, the recoveries obtained in this study can be
considered to be within reasonable range. Based on these findings, 5-minute
sonication extraction was determined as the extraction method to be used for NP
compounds from sludge solids.
Based on obtained recovery results, sonication assisted extraction method was
applied for the determination of the target compounds in solid phases of sludge
samples. For this purpose, freeze-dried and sieved dewatered sludge samples (0.5 g)
were placed into 12 mL amber vials and then 10 mL of acetone was added into each
vial. The vials were placed into ultrasonication bath and sonicated for 5 min.
Following sonication, the vials were centrifuged at 2500 rpm for 10 min. The
obtained extract (in acetone) was passed through sodium sulfate column for removal
of moisture. If necessary, after doing required dilutions, the extract was injected into
GC/MS following derivatization.
4.2.5.2. Extraction from Aqueous Samples
For extraction of NP compounds from aqueous samples (water, wastewater, lake,
seawater etc.) liquid-liquid extraction (Bennie et al. 1997, Helaleh et al. 2001, Liu et
al. 2004, Conn et al. 2006), and steam distillation (Ahel and Giger 1985) methods
were applied in early on studies. Due to need for large amount of samples, high
volume of solvents and long extraction times, these methods started to lose their
attractiveness. Instead of these methods, solid phase extraction (SPE) method has
become appealing recently for extraction of many organic compounds including NP
compounds from aqueous samples. The application of SPE is simple and requires
less solvent and short extraction time. In order to provide simultaneous extraction of
NP compounds in an efficient way, SPE method has been applied based on the
128
studies in literature (Ding and Tzing 1998, Isobe et al. 2001, Hao et al. 2006,
Gatidou et al. 2007, Guerreiro et al. 2008, Lian et al. 2009, Henriques et al. 2010).
Solid phase extraction relies on partitioning of the target compound between a liquid
(either sample or solvent) and a solid (sorbent) phase. The solid phase is a cartridge
packed with a suitable adsorbent. The selection of suitable adsorbent is the first key
step for extraction of target compounds. In literature, different SPE cartridges and
conditioning procedures have been reported for extraction of hydrophobic
compounds (Jonkers et al. 2001, Petrovic et al. 2001, Jeannot et al. 2002, Dıá z et al.
2004, Isobe and Takada 2004, Martnez et al. 2004, Koh et al. 2005, Motegi et al.
2006, Stasinakis et al. 2008). When chemical properties of NP compounds were
taken into consideration, reversed phase C18 cartridges were suggested for their
extraction.
The conditioning of cartridges prior to loading of liquid samples was an essential
step to ensure strong interaction with the target compound. Conditioning step results
in a wetting of the adsorbent and thus provides an environment which is suitable for
adsorption of the compound. During conditioning and extraction a certain flow rate
was employed using a vacuum manifold (VAC Elut 12, Agilent Technologies Inc.,
USA) for elution of liquid samples and solvents through cartridges. The cartridges
were placed on a vacuum manifold (VAC Elut 12, Agilent Technologies Inc., USA)
and conditioned by 3 x 4 mL of acetone, 3 x 4 mL of methanol and 2 x 3 mL double
distilled water, sequentially at a constant flow rate. In order to optimize SPE method,
different solvents like acetone, methanol, hexane and methanol:acetone (1:1) were
tested to determine appropriate extraction solvent for elution of NP compounds
from cartridges. For this purpose, aqueous phases of sludge samples (effluent part)
were spiked with NP, NP1EO, NP2EO and NP1EC at 50 µg/L concentration and
then samples were percolated through conditioned tC18 cartridges. After that, NP
compounds were eluted with the aforementioned organic solvents. The recovery
studies with hexane did not give satisfactory results so hexane was eliminated from
further recovery studies. The average recoveries for NP compounds employing
different elution solvents are presented in Table 4.6.
129
Depending on the results presented in Table 4.6, acetone:methanol (1:1, v/v) yielded
satisfactory recoveries which were in acceptable range reported in Method 8000C by
USEPA (2003). With these findings, the solid phase extraction of NP compounds
from aqueous samples was carried out by using acetone:methanol (1:1, v/v) as
elution solvent.
Table 4.6. The mean recoveries (%) of NP compounds in spiked water samples after
solid phase extraction with different solvents
Acetone
Methanol
Acetone:Methanol
4-n-NP
70.0±8.6
94.0±5.6
114.3±9.15
NP
60.3±6.1
198.6±10.4
96.3±3.70
NP1EO
71.7±5.8
72.3±6.2
81.6±6.20
NP2EO
96.8±8.6
85.8±7.6
83.1±4.98
NP1EC
117.0±9.4
99.5±10.9
105.1±4.70
The trials for different solvents yielded the methanol:acetone (1:1) mixture as the
most effective solvent in eluting the target chemicals. For the extraction of NP
compounds from aqueous samples from digesters, all steps mentioned in previous
paragraph were applied by percolation of liquid samples through tC18 cartridges and
elution of NP compounds with methanol:acetone (1:1) mixture as the extraction
solvent.
4.2.6. Recovery
The average recoveries (%) and relative standard deviations (RSD%) obtained are
summarized in Table 4.7 and Table 4.8 for sonication-assisted and solid phase
extraction methods, respectively. Acceptable recovery values range from 70 to 130%
for most analytes in GC/MS based methods (USEPA, 2003). On the other hand, the
recoveries obtained for NP compounds vary from study to study and the range
reported for these compounds changes between 60 to 150% (Arditsoglou and Voutsa
130
2008, Lian et al. 2009). In the light of these findings, the recovery values acquired in
this study can be accepted as in satisfactory and reasonable range (Tables 4.7 and
4.8). The RSDs of all recovery analyses are less than 10%. So it indicates the
precision and accuracy of the method is very well-established.
Table 4.7. Recoveries and RSDs obtained from extraction of spiked NP compounds
from solid phases
Compound
Spiked concentration
Recovery
RSD
4-n-NP
NP
(µg/mL)
50
50
(%)
101.48±3.21
92.88±3.43
(%)
3.69
3.16
NP1EO
50
134.62±5.78
4.29
NP2EO
50
135.25±2.71
5.10
NP1EC
50
97.5±2.45
4.80
Table 4.8. Recoveries and RSDs obtained from extraction of spiked NP compounds
from aqueous phases
Compound
Spiked concentration
Recovery
RSD
4-n-NP
NP
(µg/mL)
50
50
(%)
114.27±9.15
96.27±3.70
(%)
8.01
3.84
NP1EO
50
81.60±6.20
7.86
NP2EO
50
83.10±4.98
6.10
NP1EC
50
105.1±4.70
4.50
The recoveries obtained were quantitative in both methods with satisfactory RSD
values and indicates the suitability of the developed extraction methods for the
analysis of NP compounds in solid and aqueous samples.
131
4.2.7. Quality Assurance/Quality Control (QA/QC)
4.2.7.1. Linearity
The linearity of the response of NP compounds was determined by analyzing a series
of 8 standards in the concentration range of 0-1000 µg/mL during preparation of
calibration curves. The peak area of each compound and internal standard (4-n-NP)
was plotted against the concentration range of the standards which exhibited a linear
relationship with the correlation coefficients (r2 >0.99). In order to apply the linear
regression analysis for quantitative purposes, r2 must be ≥0.99 according to USEPA
Method 8000C (USEPA, 2003). The obtained data for target NP compounds are
summarized in Table 4.9.
Table 4.9. Quantitative calibration data for NP compounds
Compound
Concentration range
Equation
Correlation coefficient
4-n-NP
(µg/mL)
10-1000
y=303413x-3E+06
(r2)
0.999
NP
10-1000
y=216895x-4E+06
0.996
NP1EO
10-1000
y=16724x-80723
0.999
NP2EO
10-1000
y=12502x-123871
0.997
NP1EC
50-1000
y=60281x+282810
0.992
Stock solutions of NP compounds were prepared in acetone as 1000 µg/mL at
analysis day. Required dilutions to prepare standards at different concentrations were
conducted by using stock standard solutions. Standards were prepared as three
replicates and duplicate injections of each standard were performed in calibration
curve preparation.
4.2.7.2. Repeatability
To test repeatability of developed method, known concentrations of standards for
each compound (20 µg/mL and 500 µg/mL) were injected to GC/MS (n=6) and
132
obtained average concentration values and RSDs were evaluated. The results
obtained from repeatability studies are demonstrated in Table 4.10. The overall
method repeatability based on the relative standard deviations of the replicate (n=6)
analysis of two different standards was satisfactory, with RSD values varied between
1.10 and 3.18 with less than 10%.
Table 4.10. Repeatability studies at two different concentrations of NP compounds
Compound
4-n-NP
NP
NP1EO
NP2EO
NP1EC
Standard concentration
Average Concentration
RSD
(µg/mL)
20
500
(n=6)
21.15±0.47
498.26±7.42
(%)
2.22
1.49
20
20.92±0.35
1.66
500
501.06±5.52
1.10
20
21.45±0.68
3.18
500
498.46±5.86
1.21
20
19.90±0.26
1.32
500
505.49±9.93
1.96
50
47.53±2.92
2.85
500
513.24±3.36
3.07
4.2.7.3. Limit of Detection and Quantification (LOD and LOQ)
The limit of detection (LOD) and limit of quantification (LOQ) parameters were
determined for both extraction methods. The LOD can be defined as the minimum
concentration of a substance which could be detected in a sample with high level of
confidence (99%). LOQ is the minimum concentration of a substance in a sample
which can be quantitatively determined with satisfactory precision and accuracy.
In this study, LOD and LOQ values were determined using signal to noise (S/N)
ratio (ICH, 1996). S/N ratio is a measure of signal (S) strength created by the
substance relative to noise (N) created by background. The ratio represents the
quality of the signal obtained from the sample.
133
LOD and LOQ studies were performed by spike of known low concentrations of
standard solutions on sludge samples (ICH, 1996). Injections were carried out (n=6)
for each compound with blanks and following equations were used for calculation of
LOD and LOQ;
LOD = 3 ( )
(4.1)
LOQ = 10 ( )
(4.2)
For each compound, the calculated LOD and LOQ values based on Equation 4.1 and
4.2 are presented in Table 4.11.
Table 4.11. Limits of detection (LOD) and limits of quantification (LOQ) values for
the developed extraction methods
Compound
Spiked concentration
LOD
LOQ
4-n-NP
NP
(µg/g)
0.1
0.1
(µg/g)
0.00003
0.006
(µg/g)
0.0001
0.02
NP1EO
0.1
0.006
0.02
NP2EO
0.1
0.012
0.04
NP1EC
0.1
0.030
0.10
4.3. Method Application Studies
The optimized methods described in the previous parts were applied to dewatered
sewage sludge samples and solid and liquid fraction of digester sludge samples
successfully in this study. In this part experiments done on samples collected from
WWTP is discussed. In Chapter 5, results from digesters operated are discussed.
In order to monitor the levels of NP, NP1EO and NP2EO which are strictly regulated
for land application of sewage sludge in Europe (including Turkey), dewatered
134
sludge samples were taken monthly for eleven months from Ankara Central WWTP
during the year of 2012.
As can be seen in Figure 4.22, the results indicates that these compounds exist in
dewatered sludge samples collected from WWTP in Ankara. From the three
compounds tested, NP1EO or NP2EO had the highest concentration in most of the
sampling times. Results showed that the concentrations were not high. In none of the
months that the monitoring was conducted, the regulatory level of 50 mg/kg of NPE
was exceeded. On the other hand, the results also indicated a seasonal variation of
the sum NPE concentrations exist. In February, NPE concentration was determined
as 6.99±0.17 mg/kg-dm and during spring time it reached 8.58±0.34 mg/kg-dm (in
May). Results showed that in spring months, only a slight change was observed in
level of NPEs in dewatered sewage sludges. With the start of summer season, an
increase in concentration of NPE started and continued over summer months. In June
NPE concentration was determined as 8.69±0.17 mg/kg-dm and then with a
noteworthy increase it reached 12.36 ±0.31 mg/kg-dm in July. In August, it came as
high as 19.53±1.33 mg/kg-dm as the highest level of NPEs measured in all year.
With the month of September, a decrease was observed and levels went down to
about 10 mg/kg-dm or lower for NPE. In winter time, NPE concentrations slightly
fluctuated over time but exhibited much lower concentrations compared to summer
season.
135
25
NP
NP1EO
NP2EO
sum of NPE
Concentration (mg/kg-dm)
Concentration (mg/kg-dm)
20
15
10
5
ER
ER
MB
CE
DE
VE
MB
BE
NO
TO
OC
EM
BE
R
R
T
US
SE
PT
LY
AU
G
JU
NE
JU
MA
Y
IL
AP
R
H
MA
RC
FE
BR
UA
RY
0
Figure 4.22. Concentrations of NP, NP1EO, NP2EO and sum of NPE in dewatered
sludge samples
High values during summer months are likely due to the combination of dry weather
and increased washing and bathing frequencies. These two conditions likely lead to a
rise in detergent use and introduction of more nonylphenol compounds into sewage
system which at the same time is receiving a lower volume of water input due to dry
season. Conversely, in winter and spring, the decrease in NPE compound
concentrations can be attributed to snow and heavy rain regime which contribute to
dilution of NP compounds present in the waste stream. During all months, total
concentrations were consistently well below the 50 mg/kg-dm, the regulatory limit.
136
The application of proposed methods on solid and liquid phases of digester sludge
samples is going to be demonstrated in Chapter 5.
4.4. Conclusion
A reproducible and practical method for the simultaneous determination and
quantification of 4-n-NP, NP, NP1EO, NP2EO and NP1EC in sewage sludge and
aqueous samples was developed. During method development and optimization
process, key parameters for extraction and determination of these compounds were
studied in details.
Different derivatization agents and catalysts, time-periods and temperature ranges
were tested in this study. Based on the results, BSTFA+TMCS (1%) mixture yielded
better peaks. Optimum temperature and time for derivatization reaction were
determined as 70°C and 30 min, respectively. The derivatized forms of the NP
compounds were analyzed using GC-MS.
For sludge samples, sonication, mechanical shaking and a combination of both
methods were tested using different extraction times and solvents. It was found that
5-min-sonication-assisted extraction in acetone gave better results in terms of
recovery and reproducibility. For aqueous samples, application of solid phase
extraction using reversed phase C18 cartridges and elution of NP compounds with
acetone:methanol (1:1) resulted in satisfactory recoveries in the acceptable range
(USEPA, 2003).
In 2010, for the EU harmonization process, Turkey set a limit value for NPEs
(NP+NP1EO+NP2EO) as 50 mg/kg dry solids. Although a limit value has been set,
any method or procedure has not been announced by authorities for detection and
quantification of these compounds. Also, no study has been reported indicating the
level of these compounds in Turkish dewatered sludge samples.
With this study, it has been also aimed to draw attention of authorites and
researchers to accumulation of NP compounds in Turkish dewatered sludge samples
137
prior to land application. Ankara is a capital city with high population and less
industrialization. NPE compounds concentration reached nearly 20 mg/kg-dm (in
August) as the highest level measured in collected sludge samples. In highlight of
these findings it can be pointed out that environmental compartments of highly
populated and industrilazed cities can be polluted heavily with NPE compounds and
it can lead to serious environmental and public health problems. Therefore, Turkey
should start monitoring NPE concentrations in different environmental systems like
lake, river, sediment, sludge (for land application) etc. of highly populated and
industrialized cities and if necessary, the use of NPnEO as surfactant in industrial
applications should be limited or banned with regulations.
138
CHAPTER 5
BIODEGRADATION OF NONYLPHENOL DIETHOXYLATE IN
LAB-SCALE ANAEROBIC DIGESTERS
5.1. Introduction
Land application of sewage sludge is considered to be a sensible and appealing
solution for the ultimate disposal of municipal wastes. Although sludge contains
essential nutrients for plant life, agricultural use of sewage sludge has become
questionable due to its potential to introduce trace contaminants and their
metabolites into the environment (Petrovic and Barcelo, 2004). Surfactants are one
of the most widely discharged organic compounds into sewer systems. In particular,
nonylphenol
polyethoxylates
(NPnEOs)
reach
the
environment
in
high
concentrations through wastewater treatment plants (WWTPs) due to their
widespread household, industrial and agricultural applications (Talmage 1994).
Biological degradation of NPnEOs by loss of ethoxy groups leads to formation of
more hydrophobic and lipophilic degradation products with shorter ethoxy chains
including nonylphenol diethoxylate (NP2EO), nonylphenol monoethoxylate
(NP1EO) and nonylphenol (NP) by further de-ethoxylation in anaerobic
environments (Giger et al. 1984, Ahel et al. 1994, Ahel et al. 1994, Ahel et al. 1994).
Therefore, NP becomes the main break-down product of NPnEOs following an
anaerobic digestion process. On the other hand, nonylphenol carboxylates (NP2EC
and NP1EC) are major degradation products under aerobic and anoxic conditions.
The chemical structures of some example compounds of the group are depicted in
Figure 5.1
139
Figure 5.1. Structure of NPnEOs and potential degradation products
Studies have revealed that the degradation products, particularly nonylphenol, are
more persistant, toxic and bioaccumulative than the parent compound. Due to
mimicry of natural oestrogens by interacting with oestrogen receptors, NP has been
defined as an endocrine disrupting compound (EDC) (Jobling et al. 1996).
Nonylphenols can cause estrogenic effects on fish and higher life-forms, which can
lead to the feminization of male reproductive organs at low concentrations (Folmar
et al. 2002). Some studies also claim a cause and effect relation between
nonylphenols and some health problems such as hormone-related cancers and
decreases in the count or quality of sperm in humans (Toppari et al. 1996, Swan et
al. 1997).
140
Due to environmental and health concerns associated with these compounds, NP
compounds have been listed as a priority substance in the European Water
Framework Directive (EU Directive 2000/60/EC (WFD 2000)). Also, due to
controversy surrounding the application of sewage sludge on land, the European
Union proposed a limit value in “Working Document on Sludge-3rd Draft for NPEs
(NP+NP1EO+NP2EO) as 50 mg/kg dry solids. In some European countries,
NPnEOs have begun to be substituted by ethoxylated linear alcohols to avoid the
toxicity and poor degradability of NPnEOs. However, alcohol ethoxylates are
inferior detergents. Because of lower production costs and easier handling, NPnEOs
are still widely used, especially for industrial applications in USA, southern Europe
and developing Asian countries (Chiu et al. 2010). Therefore, it is important to
monitor their fate in environmental systems and to understand the mechanisms of
their transformation.
For sewage sludge stabilization, anaerobic digestion is more attractive process with
regard to energy considerations (Patureau et al., 2008). It is known that anaerobic
digestion is also commonly preferred option for treatment of industrial wastes
especially those originated from industries contributing notable amount of NPEs into
environment (Puteh et al.). It has been widely reported that NPnEOs are further
degraded in anaerobically stabilized sewage sludge to fully de-ethoxylated 4-NP, a
more lipophilic and toxic compound resistant to further microbial degradation (Ahel
et al. 1994, Ejlertsson et al. 1999, Bruno et al. 2002, Johnson et al. 2005, Lu et al.
2008, Patureau et al. 2008). Contrary to the popular understading, Chang et al.
reported nonylphenol degradation under anaerobic conditions with the order of
degradation rates: sulphate-reducing conditions>methanogenic conditions > nitrate
reducing conditions (Chang et al. 2004, Chang et al. 2005). In contrast to these
results obtained under anaerobic conditions, degradation of NP compounds has
largely been restricted to aerobic conditions (Ekelund et al. 1993, Staples et al. 2001,
Corvini et al. 2004, Yuan et al. 2004, Gabriel et al. 2005, Soares et al. 2006).
Although field studies are important for determination of these compounds in
environmental systems, operation of laboratory scale aerobic and anaerobic reactors
141
is crucial to get information about their degradation products and to understand the
step by step degradation mechanisms. The use of pure cultures does not represent the
real situation in environmental systems because mixed culture microorganisms
participate in degradation of compounds in real systems and they contribute
discrepant properties to the systems due to differences in degradation ability and
dominance. Besides, they do not necessarily provide information about what
happens in a sludge digester, which is a complex environment and is the last
“bioreactor” before the land application of sludges.
It is critical to monitor and evaluate the behavior of these compounds in the absence
of oxygen to decrease pollution in discharged environments. In this context, the aim
of this part of the study was to evaluate the fate of NP2EO in semi-continuous
anaerobic digesters and monitor anaerobic transformation of NP2EO into its
degradation products by anaerobic microbial community. The identification and
quantification of NP2EO and its degradation products, NP1EO, NP and NP1EC were
performed by gas chromatography/mass spectrometry (GC/MS). The effects of
realistic concentrations of NP compounds on digester performance were monitored
by pH, total solids (TS), total suspended solids (TSS), volatile solids (VS), volatile
suspended solids (VSS), chemical oxygen demand (COD) and biogas production
analyses.
5.2. Materials and Methods
5.2.1. Sludge Samples
In order to set laboratory scale semi-continuous anaerobic digesters, waste activated
sludge (WAS) and anaerobically digested sludge (ADS) were supplied from Ankara
Central Wastewater Treatment Plant (Ankara, Turkey). The plant has a current flow
rate as 765,000 m3/day (ASKI, 2015). Weekly WAS samples taken from the plant
were used to provide daily fresh feed for the digesters.
142
5.2.2. Chemicals
The standard solutions containing 5 µg/mL for NP, NP1EO and NP2EO in acetone
(OEKANAL) were purchased from Fluka (Sigma Aldrich Co. LLC, USA).
Surrogate, 4-n-NP (10 ng/µL in Cyclohexane), was supplied from Dr. Ehrenstorfer
(Augsburg, Germany). NP2EO (10 mg, 99% purity) was used as parent compound to
spike the digesters and was supplied from Dr. Ehrenstorfer (Augsburg, Germany).
Prior to GC/MS injection, a derivatization procedure was applied to the samples by
using BSTFA+TMCS (99:1), Sylon BFT kit purchased from Supelco Analytical
(Sigma Aldrich Co. LLC, USA). For water extraction, Sep-Pak Vac (6cc, 500 mg)
certified reversed C18 (tC18) cartridges supplied from Waters (Milford, MA, USA)
were used. Sodium sulfate (anhydrous), fine powder copper and all organic solvents
with GC grade (hexane, acetone, and methanol) were purchased from Merck KGaA,
Germany.
Stock solutions for calibration were prepared in acetone as 1000 mg/L and kept at 18oC in amber vials.
5.2.3. Digester Set-up
Duplicate lab-scale semi-continuous anaerobic digesters were operated with 2Lsludge-volume as 3 sets (spiked, abiotic control, and biotic control digesters).
Digesters mixed continuously by magnetic stirrers were operated at 35oC in a dark
hot-room.
The three sets of the lab-scale anaerobic semi-continuous digesters were operated to
monitor the biodegradation of NP2EO into its degradation products such as NP,
NP1EO and NP1EC to understand degradation pattern of these compounds. Each set
served a certain purpose as follows:
Set 1: Abiotic Control Digesters: This digester set was composed of two replicate
digesters. These digesters were denominated as “AD-1” and “AD-2” in the text.
143
These digesters were spiked with 3000 µg/L NP2EO (in acetone). Microbial
activities in these digesters were terminated by autoclaving. These digesters were fed
the same way like spiked digesters but feed was autoclaved prior to daily feeding
process to phase out microbial activity in the feed. The abiotic control digesters
enable us to monitor abiotic degradation of target compounds, if any; to evaluate
methane production efficiency of other two-sets of digesters and to ensure for
prevailing of anaerobic environment all throughout the process without intrusion.
Besides, these digesters were operated to assess digester performance with respect to
some measured parameters.
Set 2: Biotic Control Digesters: The two digesters in Set-2 were identical to each
other as were the abiotic control digesters. These digesters were named as “BD-1
and BD-2” and were not spiked with NP2EO. Only 3 mL acetone without NP2EO
was added to these control digesters since NP2EO was solubilized in acetone and
added to the other operated digesters. These digesters had microbial activities and
were operated to see the result when NP2EO was not added to the digesters. The
effects of NP2EO addition to the digesters can be understood by comparing these
digesters with the spiked ones. Also, it provides observation of NP compounds
introduced by feeding into digester system and gives idea about how feeding affects
digester stability. Results from these digesters form a baseline for anaerobic
digesters for methane production.
Set 3: Spiked Digesters: Set-3 was operated as two replicate digesters similar to the
other sets. These digesters were named as “SD-1 and SD-2”. These digesters were
spiked with 3000 µg/L NP2EO (in acetone). The spiked digesters operated make us
observe the biodegradation of NP2EO and the formation of biodegradation products
such as NP and NP1EO. Asides from that the digesters also provide information to
understand the effects of NP2EO on digester performance.
The digesters were operated until steady state which ensured the adaptation of
microorganisms and removal of the background NP compounds in WAS. Steady
state was considered to be achieved with the stability in concentrations of MLSS,
144
MLVSS and NP compounds defined by variability less than 10%. When it was
thought that the digesters had reached steady state conditions, 3000 µg/L NP2EO
was spiked into two sets of digesters (except for biotic control digesters) at 63rd day
following the steady state achievement and then degradation products were
monitored during operation time.
Certain amount of sludge (133.4 mL) was wasted from the digesters and the same
amount of fresh activated sludge (feed) was added to the digesters daily to maintain
semi-continuous operation and to keep sludge retention time (SRT) of 15 days. All
digesters were connected to graduated gas collection cylinders (4 L) to monitor gas
production during operation. After 147 days of operation (in total), digesters were
terminated.
5.2.4. Instrumentation
In order to identify and quantify NP compounds in sludge samples obtained from
semi-continuous anaerobic digesters, a 7890A Agilent gas chromatograph coupled to
a 5975C Agilent mass spectrometer with Triple-Axis was used (Santa Clara, CA,
USA). An amount of 1 µl of the silylated sample was automatically injected in the
splitless mode at 250oC into a 5% phenyl methyl siloxane capillary column (Agilent,
19091S-433UIE, 30m×0.25mm×0.25µm HP-5MS).
The oven program developed as explained in Chapter 4 and used for NP, NP1EO
and NP2EO was as follows: initial temperature 100oC for 5 min, increase from
100oC to 160oC at a rate of 25°C/min, then to 260°C at 10°C/min (held for 5 min)
and finally to 285oC at 35oC/min (held for 7 min). On the other hand, the oven
program for NP1EC was followed as 70oC for 1 min, 30°C/min to 160oC, 5°C/min to
290oC, 290oC for 5 min.
The GC/MS was operated at Selective Ion Mode (SIM). The mass spectrometer was
operated in the electron impact ionization (EI) mode and the energy of the electrons
was kept at 70 eV. The interface was kept at 280oC and the MS ionization source and
145
the MS quadrupole were kept at 230 and 150oC, respectively. The carrier gas was
high-purity helium (99.99%) with a constant flow of 1 mL/min.
The LOD and LOQ values were calculated based on signal to noise (S/N) ratio as
decribed in details in Chapter 4 by the spike of known low concentrations of
standard solutions on sludge samples (ICH, 1996). The LOD and LOQ values were
determined as 6 µg/kg and 20 µg/kg for NP and NP1EO, 12 µg/kg and 40 µg/kg for
NP2EO and 15 µg/kg and 50 µg/kg for NP1EC, respectively. Recoveries of these
compounds in spiked sludge samples (0.1 ng mg-1, 4n-NP as surrogate) were ranged
in acceptable values suggested by USEPA and Arditsoglou and Voutsa (2008) for
NP compounds. Details of recovery and relative standard deviation calculations can
be seen in Chapter 4.
5.2.5. Derivatization
NP compounds are polar chemicals with low volatility so a derivatization process is
necessary prior to injection of these chemicals to GC/MS to make them more volatile
and thermally stable compounds. For derivatization of NP, NP1EO and NP2EO, 1
mL sample extract (in acetone) was evaporated under gentle nitrogen stream, then
100 µl BSTFA-TMCS mixture was added. The vials were vortexed for 1 min and
heated at 70oC for 30 min. After cooling to room temperature and vortexing for
another 1 min, silyl-derivatized extracts were injected to GC/MS for analysis. The
details of method development are given in Chapter 4.
For NP1EC, the derivatization method suggested by Lee et al. (Lee and Peart 1999)
was modified and applied as follows; 1 mL NP1EC solution was evaporated to 200
μL using N2 gas stream and then 2 mL 20% Boron trifluoride (BF3) solution was
added. The vials were kept at 850C for 30 min in oven. At the end of 30-min, the
vials were taken and kept at room temperature for cooling. Then the contents were
evaporated to 300 μL using N2 gas stream and placed into mechanical shaker at 400
rpm for 1 min. Following shaking step, 2.5 mL of double distilled water (ddH2O)
were added and mixed. Then 2 mL petroleum ether were added to each vial to
146
extract the methylated derivatives. This step was repeated for 3 times and then the
eluate was passed through Na2SO4 column. The eluate was evaporated to complete
dryness by N2 gas. Then, 1 mL of hexane was added to the vials and they were
vortexed for 1 min. Each sample was transferred into injection vial and then sample
was injected into the GC/MS.
5.2.6. Extraction
5.2.6.1. Extraction from solid phase
Extraction of NP compounds from sludge solids were carried out by 5 minsonication-assisted extraction with acetone. The sludge samples were centrifuged at
2500 rpm for 10 min to provide separation of solids and liquids. After separation by
centrifugation, the extraction vials containing solids were put into freezer at -18oC.
Next day, the vials were placed into freeze-dryer for 24 hr to remove water. Then, 10
mL acetone was added to freeze-dried samples and the vials were placed into
sonication bath (Ultrasonic, Falc, USA) for 5 min. Following sonication, the vials
were centrifuged at 2500 rpm for 10 min. The extract (in acetone) was passed
through sodium sulfate column for moisture removal, evaporated under nitrogen
stream and injected into GC/MS following derivatization.
5.2.6.2. Extraction from liquid phase
Solid phase extraction (SPE) method using tC18 cartridges (Sep-Pak, Waters) was
used for extraction of NP compounds from liquid phases of sludge samples. All SPE
cartridges were conditioned by 3x4 mL of hexane, 3x4 mL of methanol, 3x4 mL of
acetone and 2x3 mL of Milli-Q water at a flow rate of 1 mL/min. After conditioning
process, liquid phase of sludge samples was passed through the cartridge at the same
flow rate. The retained substances were eluted with 10 mL of acetone:methanol (1:1)
mix at the same flow rate. The extract was evaporated to dryness with a gentle
stream of nitrogen and then injected to GC/MS following the derivatization.
147
5.2.7. Analytical Methods
The gas composition was determined with a Gas Chromatograph (GC)/Thermal
Conductivity Detector (TCD) (Agilent Technologies 6890N). Nitrogen (N2),
methane (CH4) and carbon dioxide (CO2) composition of the produced gas in the
digesters were monitored. The column equipped with the instrument was a 30.0
mX530μmX40.0μm nominal HP-Plot Q capillary column. The oven program was as
follows: column temperature was held at 45°C for a minute and then increased to
65oC at a rate of 10oC/min. Helium was the carrier gas with a flow rate of 3 mL/min.
On each analysis day, the GC/TCD was calibrated by using two different standard
gas mixtures.
The pH measurements were conducted by a pH meter of Oakton Waterproof 300
series (pH/mV/oC) according to Standard Method 4500H (Federation and
Association 2005). TS, TSS and VS-VSS of sludge samples were determined
according to Standard Methods 2540B, 2540D and 2540E, respectively (Federation
and Association 2005). These analyses were carried out as replica during operation
period and the values were represented as average.
To determine tCOD of the sludge samples, Hach Method 8000 named “Reactor
Digestion Method” approved by United States Environmental Protection Agency
(USEPA) was used (Jirka and Carter 1975). COD readings were conducted using
Hach DR 2400 spectrophotometer at 620 nm wavelength. Triplicate analyses were
carried out for each sludge sample and then the average value was reported as tCOD.
5.3. Results and Discussion
5.3.1. pH
pH is one of the critical parameters monitored during the operation of digester
systems due its effect on the activity of microbial population. In the semi-continuous
anaerobic digesters, for methane production pH is one of the important operational
148
parameters, therefore the digesters should be operated in the pH range which is
optimum for methanogens responsible for methane production. It has been reported
that the optimum pH range for methanogens is pH 7-8 and the operational pH range
for anaerobic systems is 6.5-8.5 (Hobson and Wheatley 1993). The other factor to be
considered about pH is that NP compounds are easily degraded at neutral pH values
(Chang et al. 2005).
In light of all these considerations, when the measured pH values over 147 days are
examined (Figure 5.2), it is seen that the pH of spiked digesters (SD-1/SD-2) and
biotic control digesters (BD-1/BD-2) were very close to each other, ranging between
7.0 and 7.6 throughout 147 days. These pH values were consistent with the required
pH values for methanogens and the degradation of NP compounds, so it can be said
that these digesters were operated in an acceptable pH range. For abiotic control
digesters (AD-1/AD-2), the initial pH was close to 8 but then decreased to 6.9 in five
days and the pH ranged within 6.1 and 6.9 in the following days. The common
feature of BD and SD digesters compared to AD control digesters was the presence
of microbial activity. The digesters with microbial activity exhibited the same pH
pattern all throughout the digester operation. The spike of NP2EO to the spiked
digesters at the 63rd day did not lead to any remarkable change in the pH. Also, after
the 70th day, the pH of the spiked and biotic control digesters remained stable. It can
be concluded that the addition of NP2EO and the formation of degradation products
demonstrated no considerable effect on the pH. Also, no significant fluctuation was
observed in the pH due to the daily feeding to these digesters. Therefore no buffer
was added to the digesters to adjust the pH during operation period.
149
BD-1
BD-2
SD-1
SD-2
AD-1
AD-2
Feed
8.5
8.0
pH
7.5
7.0
6.5
6.0
5.5
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.2. pH distribution vs. time for feed and operated anaerobic digesters
On the other hand, due to lack of microbial activity in AD digesters, the fluctuation
in the pH can be explained by the daily feeding. As can be seen from Figure 5.2, the
pH range of the feed was 6.1-6.5 so the addition of 133.4 mL eed to the system every
day and taking out the same amount of sludge from the digester started diluting and
changing the composition of the abiotic digesters. Since there were no biotic
reactions in these digesters, the pH value here was simulated the pH value of the
feed pretty well. Similar to the BD and SD digesters, no pH adjustment was made for
AD digesters during the operation.
5.3.2. Biogas Production and Composition
Total gas production and gas composition of each digester were monitored with
respect to time during operation process. The cumulative methane production levels
150
of each digester are given in Figure 5.3. As illustrated in Figure 5.3, the cumulative
methane production exhibited the same pattern for the biotic control and spiked
digesters. The BD and SD digesters produced gas daily and demonstrated an
increasing trend in the cumulative methane data without any problem throughout the
operation period. The BD-1/BD-2 and SD-1/SD-2 digesters followed the same
pattern which meant that microbial community grew in these digesters and that
microbial activity was consistent between these digesters all throughout the
operation time. It can be deduced from the graph that there was no problems such as
air intrusion, or toxicity etc. to upset the gas production in these digesters. In some
studies in literature, acclimatization of microorganisms to most organic pollutants
during digestion process was deemed unnecessary since the microbial community
was already exposed to these chemicals (Samaras et al. 2013). Similarly, spike of
NP2EO caused no negative effect on the digester such as induction of a lag period or
hampering of methane production.
Cumulative methane production (mL)
1.2e+5
BD-1
BD-2
SD-1
SD-2
AD-1
AD-2
1.0e+5
8.0e+4
6.0e+4
4.0e+4
2.0e+4
0.0
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.3. Cumulative methane production vs. time for operated anaerobic digesters
151
The replica digesters for each set behaved in the same manner which showed the
operation of the digesters were in constant and stable conditions. Also, the spike of
NP2EO in acetone or acetone alone to the digesters did not lead to any noteworthy
change in the gas production. Similar results were reported by Ejlertsson and coworkers (Ejlertsson et al. 1999). They used anaerobic digester sludge, landfilled
sludge and landfilled municipal solid waste as inoculum to study degradation of a
mix of NP1EO-NP2EO at 2 mg/L concentration under anaerobic conditions (at 37
°C) in 123 mL-serum bottles. They reported that the addition of NPE1O-NP2EO mix
did not result in an increase in the total amount of methane produced in all bottles
with three inocula compared to control ones (no addition of NPE1O-NP2EO mix).
These results were obtained for small-volume-bottled systems operated in batch
mode. With our study, it is revealed that addition of NP2EO into semi-continuous
anaerobic digesters operated in large-volume (2 L) did not contribute any remarkable
increase in methane production, either. This situation can be explained by remaining
of the phenol ring in intact structure during digester operation depending on findings
obtained from degradation experiments with labeled NP2EO (Ejlertsson et al. 1999).
The scissored ethoxylate groups during transformation of NP2EO into NP1EO and
NP could have been used during methanogenesis. The reason why there is no
contribution of ethoxylate groups on methane production may be explained by being
at too low concentration to make a remarkable change in methane production as
mentioned by Ejlertsson et al. (Ejlertsson et al. 1999).
For abiotic digesters, no methane production was observed except for AD-2 in the
last 30 days of the operation period. As can be seen in Figure 5.3, no indication of
gas production was exhibited for all operation time for AD-1 digester. On the other
hand, very small amount of gas production (which is not noticeable in Figure 5.3)
started in AD-2 (on 118th day). Silver nitrate (AgNO3) was added to this digester to
suppress the microbial activity. AgNO3 has been used to prevent microbial growth in
the literature (Rodrigues et al. 2006) and was added into the digesters to eliminate
microbial activity. With the addition of AgNO3, the gas production levels were kept
very low for AD-2 digester as can be seen from the Figure 5.3. The reason of
152
microbial activity starting in the last 30 days can be the contamination during the
feeding or sampling. The intervention of digesters for feeding and sampling every
day made them vulnerable for contamination. Therefore more attention was given to
the addition of daily feed to the abiotic digesters (which was autoclaved every day
and fed to the digester using a sterile syringe). The AD-2 digester produced
1851±mL 33.83 mL methane in total during all operation time. When this data is
compared to cumulative methane data of biotic control and spiked digesters (as
average of replica digesters, BDcumulative CH4 = 110408.26±489.98 mL and SDcumulative
CH4
= 113874.34±442.28 mL) it can be definitely said that it is not noteworthy
amount to change the operational purpose of this digester. Also, when the data
obtained for AD-2 were examined for the measured parameters given in the
following sections, no remarkable effect of this gas production between the days of
118-147 on other parameters was observed.
5.3.3. Chemical Oxygen Demand
The changes in COD throughout the operation period are presented in Figure 5.4.
For abiotic digesters (AD-1/AD-2), the COD values remained roughly the same for
35 days from the digester start-up, and then started to increase and remained stable
with little fluctuations up to spike day. As mentioned earlier, due to lack of microbial
activity, this increase can be caused by the daily feeding. When the COD values of
the feed with respect to time are taken into consideration, the increase in the COD of
the abiotic digesters was correlated to the increase in the COD of the feed. As
explained earlier, the aim was to keep the feed as consistent as possible over 147
days but as can be seen from Figure 5.4, at the beginning, the COD concentrations of
the feed for several weeks fluctuated due to seasonal variation (from summer to
autumn) or maintenance in treatment plant (operational problems, rain etc.). At the
spike day (63rd), a sudden increase was observed for these digesters. The feed has a
stable pattern around the spike day so the spike of NP2EO in acetone led to the
increase in the COD. Following the spike, the COD value remained high for 7 days
and then started to decrease to the COD level of the feed. It can be seen that the feed
153
COD concentrations presented small fluctuations over time so the observed regular
decrease can be attributed to a dilution effect caused by daily feeding in and removal
of sludge from the system. The feeding manner introduced a dilution effect and
when the COD concentration is compared to spiked digesters, abiotic digesters had
much higher COD values. If the abiotic digesters had microbial activity, similar
COD values would have been observed with those of the spiked digesters (SD-1/SD2). Also, the gas analyses performed for A-1/A-2 digesters proved the lack of
microbial activity.
BD-1
BD-2
SD-1
SD-2
AD-1
AD-2
Feed
60000
50000
COD (mg/L)
40000
30000
20000
10000
0
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.4. COD change with respect to time for operated anaerobic digesters.
As can be seen from the Figure 5.4, the biotic control (BD-1/BD-2) and NP2EOspiked digesters (SD-1/SD-2) presented the same trend for COD. After some
fluctuations in the COD values for 40 days, the digesters became stable up until the
spike day. The behavior of the two types of digesters was completely the same after
154
the addition of NP2EO in acetone to the SD-1/SD-2 and only acetone to the BD1/BD-2. This same pattern reveals that the sudden increment in the COD was
resulted from the addition of acetone, not the addition of NP2EO. Therefore, it can
be deduced that NP2EO alone had no COD contribution to the system for these
digesters. The COD remained high for the following 7 days and then a steady
decrease was observed till the 108th day. After that point, COD remained stable until
the termination of the digesters. The COD removals were calculated as given in
Table 5.1. The overall COD removal ranged within 55.71-59.06% for the digesters.
As can be seen from the results, no remarkable difference was present between the
BD-1/BD-2 and SD-1/SD-2 digesters. For typical digester performance, COD
reduction efficiency was reported as between 40 and 60% under mesophilic
conditions (Speece 1996, Appels et al. 2008). With obtained results, it can be said
that the performance of the digesters operated under mesophilic conditions are
comparable with literature by being in a range of 40-60% of COD reduction
efficiency.
Table 5.1. Percent COD removals for biotic control and spiked digesters
BD-1
BD-2
SD-1
SD-2
Initial COD (mg/L)
24740±953.94
24960±1020.79
24890±817.98
24320±932.59
Final COD (mg/L)
10348.15±491.2
10218.01±447.30
10706.05±425.33
10771±449.10
COD Removal (%)
8
58.17
59.06
56.99
55.71
5.3.4. Total Solids and Volatile Solids
TS and VS are the parameters monitored during the digester operation. In anaerobic
digester systems, TS represents the amount of total solids in the system and is
performed on sludge samples to monitor stability of the anaerobic digester systems.
On the other hand, VS represents the digestible biomass (organic matter) in
anaerobic digesters, in other words, biologically treatable part of the total solids and
155
is monitored to assess digester efficiency. Also, gas production generally is reported
based on mass of VS in literature.
The TS and VS measurements performed for the digesters with respect to operation
time are shown in Figure 5.5.
BD-1
BD-2
SD-1
SD-2
AD-1
AD-2
28000
TS (mg/L)
26000
24000
22000
20000
18000
16000
14000
16000
VS (mg/L)
14000
12000
10000
8000
6000
0
10
20
30 40
50
60
70
80
90 100 110 120 130 140 150
Time (days)
Figure 5.5. Change in TS and VS concentrations vs. time for operated digesters
156
The biotic control and NP2EO spiked digesters exhibited same trend for the
reduction of TS and VS over the operation time. As can be demonstrated in Figure
5.5, there was a regular decrease in TS for these digesters up to the 30th day and then
a steady pattern is observed up to the spike-day (63rd day). The time elapsed between
the digester-start up and the spike was committed for the achievement of steady
state. In literature, it was reported that the analyses should be performed after the
systems are operated for at least 2-3 SRTs for adaptation and to give a response to
the new conditions (Forster and Dallas-Newton 1980). It can be deduced that these
digesters reached steady state at about the 30th day (2 SRTs) and they were operated
for another 30 days at steady state based on Figure 5.5. After ensuring the digesters
reached steady state, NP2EO was dosed to the spiked digesters (SD-1/SD-2) at 63rd
day. As can be illustrated in Figure 5.5, these digesters followed the same steady
pattern after the spike for TS. No remarkable decrease in TS was observed in the
following days after spike. It can be seen that the addition of NP2EO into the spiked
digesters did not lead to any change in the reduction of TS compared to the biotic
control digesters (BD-1/BD-2); in other words, the degradation of TS in the spiked
digesters was not affected by the spiking of NP2EO. On the other hand, for VS
measurements, a noteworthy decrease in VS for these digesters continued after the
spike with no difference observed in the patterns of control and spiked digesters
(Figure 5.5). Therefore, it can be said again that the NP2EO spike had no effect on
removal efficiencies of TS and VS in the digesters.
When the TS and VS removal efficiencies (% removal) were calculated for BD and
SD digesters, similar efficiencies for each parameter were obtained for all the
digesters. As can be seen from the Table 5.2, TS and VS removal efficiencies varied
between 36.8 and 39.8% and 52.8 and 53.5% for the digesters, respectively.
157
Table 5.2. % TS and VS removals for biotic control and spiked digesters
BD-1
BD-2
SD-1
SD-2
Initial TS
Final TS
25650
16210
25980
16140
25840
15560
25545
15685
TS Removal
36.80
37.88
39.78
38.60
Initial VS
15080
15260
15220
15140
Final VS
7120
7180
7090
7040
VS Removal
52.78
52.95
53.42
53.50
The organic matter removal efficiency (%VS removal) for anaerobic digesters has
been reported as 40-60% (Speece 1996) in literature. So the findings for VS removal
efficiency are consistent with the literature. Also, McNamara et al. (McNamara et al.
2012) dosed Igepal CO-210 (mixture of NP1EO and NP2EO) into a lab-scale
mesophilic anaerobic digester and 50% VS reduction was reported at the end of
operation time. As can be seen, removal efficiency of anaerobic digesters operated in
this study also exhibited similar reduction efficiency in VS in the presence of
nonylphenol compounds.
5.3.5. Total Suspended Solids and Volatile Suspended Solids
TSS and VSS were the other solids parameters monitored during the digester
operation. TSS measurements indicates the amount of the non-soluble solids and
VSS is another indicator of organic matter in the digesters and generally it is
considered as representing the microbial biomass of the system. As can be seen from
Figure 5.6, TSS and VSS measurements followed the same pattern with the TS and
VS for all the digesters. The explanations offered for TS and VS in the previous
section are valid for TSS and VSS in this section, as well.
158
24000
TSS (mg/L)
22000
20000
18000
16000
14000
12000
16000
VSS (mg/L)
14000
12000
10000
8000
6000
4000
0
10
20
30
40
50
60
70
80
90 100 110 120 130 140 150
Time (days)
Figure 5.6. Change in TS and VS concentrations vs. time for operated digesters
When the TSS and VSS removal efficiencies (% removal) were calculated for biotic
control and NP2EO spiked digesters as given in Table 5.3, TSS removal efficiencies
of the digesters are seen to vary between 36.85% and 39.76%, on the other hand,
VSS removal efficiencies changed between 52.05 and 52.83%. The TSS and VSS
reduction efficiencies of the digesters lie in a reasonable range for anaerobic
systems.
159
Table 5.3. % TSS and VSS removals for biotic control and spiked digesters
BD-1
BD-2
SD-1
SD-2
Initial TSS (mg/L)
22320
22790
22460
22020
Final TSS (mg/L)
14095.65
14034.78
13530.43
13639.13
TSS Removal (%)
36.85
38.42
39.76
38.01
Initial VSS (mg/L)
13140
13470
13280
12960
Final VSS (mg/L)
6300.88
6353.98
6274.34
6230.09
VSS Removal (%)
52.05
52.83
52.75
51.93
5.3.6. Digester Performance
In order to evaluate overall digester performance, the methane yield as a function of
VS and COD were calculated. Also, average methane production rate for biotic
control and NP2EO-spiked digesters was determined to compare digester
performance in the absence and presence of NP2EO (Table 5.4). Besides, the overall
reduction efficiencies of digesters for measured parameters are also summarized in
Table 5.4 to represent overall performance of operated digesters.
The biotic control and NP2EO spiked digesters were operated for 147 days under
mesophilic conditions. These digesters exhibited the same pattern with regard to
reduction of solids contents (TS, VS, TSS and VSS). In a similar way, no remarkable
difference was observed in COD reduction abilities of BD and SD digesters. When
methane compositions of digesters given in Table 5.4 are compared, the methane
content (%) of total produced biogas in digesters are nearly same for all operated
digesters. It has been reported that anaerobic digestion of organic wastes produce
biogas which is composed of 55-70% methane (CH4) and 30-45% of carbon dioxide
(CO2) on volume basis (Speece 1996, Korres and Nizami 2013). In light of this
information, the amount of produced methane (as percent by volume) in all digesters
are compatible with literature.
160
Table 5.4. Evaluation of digester performance of biotic control and spiked digesters
BD-1
BD-2
SD-1
SD-2
TS reduction (%)
36.80
37.88
39.78
38.60
TSS reduction (%)
36.85
38.42
39.76
38.01
VS reduction (%)
52.78
52.95
53.42
53.50
VSS reduction (%)
52.05
52.83
52.75
51.93
COD reduction (%)
58.17
59.06
56.99
55.71
Mean methane composition, %
57.23
59.06
57.48
57.84
Methane productivity (L-CH4/L·day)
0.378
0.374
0.388
0.387
Specific methane yield (Y, L-CH4/g-VSadded)
0.375
0.367
0.382
0.383
Specific methane yield (Y, L-CH4/g-CODremoved)
0.368
0.372
0.383
0.385
Methane productivity (production rate) was determined for each digester.
Productivity varied between 0.374 and 0.388 L-CH4/L reactor-volume·day as given
in Table 5.4. The methane productivities are very close to each other. This is due to
the operational conditions (temperature, mixing speed etc.) which were kept the
same and stable during digester-operation period and the fact that NP2EO addition
did not result in any negative effect on methane productivity for spiked digesters.
Specific methane yield (Y) which is a more meaningful parameter for sludge
digesters was calculated as the volume of methane produced per unit mass of VS
added or per unit mass of COD removed. In some studies methane yield was
expressed based on VS added and a typical range was reported as 0.370-0.506 L/gVSadded (Benefield and Randall 1980). When methane yields of operated digesters
given in Table 5.4 are examined, all values are seen to be in the given range of
literature. When a comparison is made based on specific yield of methane, spiked
digesters exhibited a slightly better yield per g of VS-added than that of BD
digesters. With the help of this information, it can be said that addition of NP2EO
into digesters did not upset microbial flora and their activities overall. On the other
hand, in some studies, methane yield was stated in terms of methane production per
unit amount of COD removed. It is reported that theoretically 395 mL of methane
161
produced is equivalent to 1 g of COD removed from the waste (Speece 1996). When
the yield (L-CH4/g-CODremoved) values from Table 5.4 are considered in the light of
this information, the reported specific methane yield values based on g-COD
removed for this study are in acceptable range. The values are a bit lower than the
theoretical value and this is plausible because the specific methane yield values are
always reported lower than the theoretical methane yield due to use of some of
organic matter for synthesis of biomass, loss of some organic matter in effluent of
sample, limited degradation of sludge compared to soluble COD or toxicity of some
compounds (Müller et al. 2005, Goswami and Kreith 2007).
5.3.7. Nonylphenol Compounds
The semi-continuous anaerobic digesters were operated for 147 days to monitor
degradation of NP2EO and determine and quantify the degradation products like NP,
NP1EO and NP1EC. Therefore, the concentrations of NP1EC, NP, NP1EO and
NP2EO were measured in the solid and liquid portions of the sludge samples taken
from the digesters at certain intervals during the operation of the digesters. The
graphs are given separately for solid and liquid phases of each digester.
5.3.8. Biotic Control Digesters
The biotic control digesters were operated throughout the operation time and no
NP2EO was added to these digesters, only 3 mL acetone was added at the day of
spike. The digesters were operated for 147 days in overall. The concentrations of NP
compounds in set-up of these digesters and prior to spike-day are given in Table 5.5.
162
Table 5.5. Concentrations of NP compounds in sludge samples measured at digester
set-up and prior to spike for biotic control digesters
BD-1
BD-2
411.55±16.95
416.44±12.04
158.31±11.12
158.12±13.79
NP (µg/L)
228.21±9.55
236.65±10.68
NP1EC (µg/L)
nd
nd
128.50±12.07
130.69±14.11
181.35±10.68
184.95±12.75
NP (µg/L)
373.88±10.60
372.91±9.14
NP1EC (µg/L)
nd
nd
At digester set-up
NP2EO (µg/L)
NP1EO (µg/L)
Typeequationhere.
Prior to spike
NP2EO (µg/L)
NP1EO (µg/L)
Typeequationhere.
nd: not detected
5.3.8.1. Solid Phase of Biotic Control Digesters
The measured concentrations of NP compounds in solid phases of biotic control
digesters throughout operation time are shown in Figure 5.7 and Figure 5.8 for BD-1
and BD-2, respectively.
As can be seen from Figures 5.7 and 5.8, following the digester start-up, NP2EO
concentration of SD-1 digester started to decrease steadily till the 60th day. Then the
descent in concentration slowed down in the following 20 days. BD-2 digester also
behaved similarly and NP2EO concentration remained almost the same between 65th
and 80th days. Up to the day of spike (63rd day), it can be said that the level of the
NP2EO and NP1EO or other NP compounds attained a stable level. Up until this
day, the higher ethoxylates present in the digester were all possibly converted into
NP2EO, NP1EO and NP in both of the biotic control digesters. As can be seen in
figures, the decrease in NP2EO was accompanied by a slight increase in NP1EO
followed by a remarkable rise in NP concentration. On the other hand, no NP1EC
163
formation was observed as a by-product of biodegradation of NP compounds during
this period.
NP2EO
500
NP1EO
NP
450
450
400
350
NP2EO (µg/L)
350
300
300
250
250
200
200
NP&NP1EO (µg/L)
400
150
150
100
100
50
50
0
0
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.7. Change in NP compounds concentration (solid phase) vs. time for BD-1
(biotic control) digester
164
NP2EO
500
NP1EO
NP
450
450
400
400
NP2EO (µg/L)
350
300
300
250
250
200
200
150
150
NP&NP1EO (µg/L)
350
100
100
50
50
0
0
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.8. Change in NP compounds concentration (solid phase) vs. time for BD-2
(biotic control) digester
The absence of NP1EC and NP2EC in anaerobic systems were also reported in
literature by (Patureau et al. 2008). In this study, lab-scale continuous stirred tank
digesters (CSTR) were run as a mesophilic anaerobic digester with 20 d HRT at
35oC. It was reported that the carboxylated forms of nonylphenols, NP1EC and
NP2EC, were not detected in the liquid and solid fraction of the sludge samples
obtained from digester. On the other hand, some researchers have reported formation
of NPECs like NP2EC and NP1EC as intermediates during degradation of NPnEOs
under anoxic and anaerobic conditions (Field and Reed 1999, Lee and Peart 1999,
Schröder 2001, Ferguson and Brownawell 2003). Lee and co-workers (Lee and Peart
1999) collected digested sludge samples from seven different municipalities of
Canada and measured NP1EC and NP2EC concentrations in these samples. They
detected NP1EC and NP2EC only in three digested sludge samples out of seven and
165
their concentrations were reported as varying between 4 and 38 µg/g. These studies
were carried out by collection of sludge samples from municipalities and then
measurements were performed. Therefore it is hard to say that these studies represent
real situation due to being based on collected sludge not operated reactor or digester
systems. Also, there is a possibility of formation of these NPnECs during
transportation of sludge samples if aerobic or anoxic conditions prevail. As a result,
in this study, NP1EC was not detected in all operation period in any operated
anaerobic digesters.
For BD-1 digester, NP2EO concentration dropped from 411.55±16.95 µg/L to
128.50±12.07 µg/L in first 60 days. Similarly, NP2EO concentration decreased from
416.44±12.04 µg/L to 130.69±14.11 µg/L for BD-2 digester during the same time.
The trend in increase of NP and NP1EO concentrations followed a similar trend with
the decrease of NP2EO for 32 days and then their concentration remained more or
less the same with small fluctuations caused by daily feeding.
Based on the prevailing steady state conditions, 3000 µg/L NP2EO (in acetone) was
spiked externally into SD digesters at 63rd day. For both biotic control digesters,
where no addition of NP2EO was made, no increase in NP2EO was observed at 63rd
day (Figures 5.7 and 5.8) in solid phases. In order to demonstrate whether addition of
acetone affects biodegradation process or not, same volume of acetone without
NP2EO was dosed into biotic control digesters on the 63rd day.
The slight decrease in NP2EO concentration following acetone spike was
accompanied by the increase in NP1EO and NP concentration. The concentrations
remained nearly as 60 µg/L for NP2EO, 160 µg/L for NP1EO, 360 µg/L for NP. As
can be seen from Figures 5.7 and 5.8, no further degradation of NP1EO and NP was
observed beyond this point and digester exhibited the same pattern till the
termination day. The digesters were terminated at following concentrations:
62.98±8.57 µg/L NP2EO, 161.49±12.89 µg/L NP1EO and 364.87±11.82 µg/L NP
for SD-1 and 61.79±9.73 µg/L NP2EO,
162.58±13.99 µg/L NP1EO and
366.22±10.09 µg/L NP for SD-2. It can be said that daily feeding and acetone
166
addition did not contribute any remarkable change in concentrations of NP
compounds in the following days after spike in biotic control digesters.
5.3.8.2. Liquid Phase of Biotic Control Digesters
As can be seen from Figures 5.9 and 5.10, all NP compounds were measured at very
low concentrations in liquid phases of biotic control digesters. Due to their low
concentrations, the data fluctuated more when compared to the data of solid phases
60
100
50
80
40
60
30
40
20
20
10
NP&NP1EO (µg/L)
NP2EO (µg/L)
(Figures 5.7 and 5.8).
0
0
10
20
30
40
50
60
70
80
90 100 110 120 130 140 150
Time (days)
Figure 5.9. Change in NP compounds concentration (liquid phase) vs. time for BD-1
(biotic control) digester
167
NP2EO
60
NP1EO
NP
100
80
40
60
30
40
20
NP&NP1EO (µg/L)
NP2EO (µg/L)
50
20
10
0
0
0
10
20
30
40
50
60
70
80
90 100 110 120 130 140 150
Time (days)
Figure 5.10. Change in NP compounds concentration (liquid phase) vs. time for BD2 (biotic control) digester
At the day of spike just before spiking acetone into the biotic control digesters,
NP2EO was measured as 38.57±2.83 µg/L and 32.39±2.75 µg/L for BD-1 and BD-2
digesters, respectively. Then, following the acetone spike, due to the absence of
NP2EO into the system, slight fluctuations were observed in concentrations of NP
compounds. This fluctuation can be attributed to daily feeding or measurement of
low concentrations.
Relative comparison of Figures 5.7 and 5.8 with 5.9 and 5.10 shows that the
accumulation pattern of NP compounds exhibit a great difference between the solid
surfaces and in liquid media. The concentrations in the solid phase is nearly 10 times
higher than those in the liquid phase. This is due to the hydrophobic nature of NP
compounds, as a result they showed accumulation on sludge surfaces and major
portion of their concentration was present in solid phase. These results are consistent
168
with the reported findings in literature (Giger et al. 1984, Ahel et al. 1994, Ahel et al.
1994).
5.3.9. Spiked Digesters
In this digester set, the semi-continuous anaerobic digesters were operated to
monitor the degradation of externally dosed NP2EO and determine and quantify the
degradation products for 147 days of operation time. The concentrations of NP1EC,
NP, NP1EO and NP2EO were measured in the solid and liquid phases of the sludge
samples taken from the spiked digesters at certain intervals during the operation
time. NP, NP1EO and NP2EO concentrations were determined in the sludge samples
at digester set-up (Table 5.6) and prior to spike but NP1EC could not be detected in
the digester set-up sludges and did not form as a product during the digester
operation as explained for biotic control digesters.
Table 5.6. Concentrations of NP compounds measured in sludge samples at digester
set-up and prior to spike for spiked digesters
SD-1
SD-2
445.45±25.79
441.27±19.44
156.98±16.86
154.66±14.82
NP (µg/L)
232.57±10.24
235.79±9.48
NP1EC (µg/L)
nd
nd
164.77±16.18
155.15±17.39
133.23±12.32
131.93±14.20
NP (µg/L)
385.14±22.52
377.07±26.51
NP1EC (µg/L)
nd
nd
At digester set-up
NP2EO (µg/L)
NP1EO (µg/L)
Typeequationhere.
Prior to spike
NP2EO (µg/L)
NP1EO (µg/L)
Typeequationhere.
nd: not detected
169
5.3.9.1. Solid Phase of Spiked Digesters
The NP2EO spiked digesters (SD-1 and SD-2) were operated for 63 days without
any spike for the achievement of steady state. On the 63rd day, the digesters were
dosed with 3000 µg/L NP2EO in acetone. The concentrations of NP compounds in
the solid phases of these digesters for SD-1 and SD-2 are given in Figure 5.11 and
Figure 5.12 and in the liquid phases of these digesters are illustrated in Figure 5.13
and Figure 5.14.
NP2EO
NP1EO
NP
1600
3000
1400
2500
2000
1000
800
1500
600
1000
NP&NP1EO (µg/L)
NP2EO (µg/L)
1200
400
500
200
0
0
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.11. Change in NP compounds concentration (solid phase) vs. time for SD-1
(spiked digester)
170
NP2EO
NP1EO
NP
3000
1600
1400
2500
2000
1000
1500
800
600
1000
NP&NP1EO (µg/L)
NP2EO (µg/L)
1200
400
500
200
0
0
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.12. Change in NP compounds concentration (solid phase) vs. time for SD-2
(spiked digester)
As can be seen from Figure 5.11 and Figure 5.12, following the digester set-up,
NP2EO concentration decreased (except for several points) during the first 40 days
of operation time. Parallel to this, NP1EO and especially NP concentrations
increased in both of the replica digesters. So it can be stated that the higher
ethoxylated NP compounds including NP2EO entering the digesters were converted
into lower ethoxylated ones, including NP1EO and NP during anaerobic
degradation. Following this behavior all compounds held their concentrations nearly
the same from the 40th day onwards up until the day of the spike. The pattern of the
replica spiked digesters for 63 days of operation is mostly similar and also the
measured concentrations of the NP compounds are very close to each other. When
the spiked digesters are compared with biotic control digesters based on the data
obtained up to the day of spike, both sets of digesters exhibited similar behavior
171
(Figures 5.7-5.8 and 5.11-12). Also, when Table 5.4 and 5.5 are taken into
consideration, concentrations of NP compounds in biotic control and spiked
digesters at digester set-up are notably close to each other. So it can be claimed that
all digesters were started identical and there is no doubt about that the NP content
available in digesters were approximately the same following the set-up. The
digesters were operated under the same conditions so the expectation was that these
identical digesters to exhibit the same behavior in for all parameters measured up to
the spike day. As mentioned earlier, biotic control and spiked digesters presented
similar digester performance in terms of methane production, COD and removal of
solids contents. Besides when concentrations of NP compounds prior to spike given
in Table 5.5 and 5.6 are considered, the values with slight fluctuations resemble each
other. In the light of these findings, it can be asserted that the digesters were set
identical and behaved in similar way for biodegradation of NP compounds up to
spike day.
The digesters were operated for about four SRTs to ensure steady state and then they
were dosed with 3000 µg/L NP2EO in acetone. With the spike of 3000 µg/L NP2EO
at the 63rd day, NP2EO concentration increased sharply and reached 2815.91±77.15
µg/L for SD-1 and 2852.83±86.12 µg/L for SD-2. After the spike day, NP2EO
concentrations in solid phase of the spiked digesters started to decrease gradually
and reached 600.07±57.94 µg/L at the 83rd day and 269.80±20.53 µg/L at the 115th
day for SD-1. When SD-2 digester data was examined, it came down to
618.89±60.41 µg/L at the 83rd day and 246.45±28.62 µg/L at the 115th day. Figures
5.11 and 5.12 showed that the same pattern was observed for the biodegradation of
NP2EO in both replica digesters. The spiked digesters were terminated with
210.59±28.46 µg/L and 200.73±24.67 µg/L NP2EO concentrations for SD-1 and
SD-2, respectively. The obtained values were very close to each other so it can be
concluded that replica digesters were operated successfully in same operational
conditions without any disturbance.
When the NP1EO and NP concentrations presented in Figure 5.11 and 5.12 are
examined, it can be seen that on the 63rd day, an increase was observed in the
172
NP1EO for both replica digesters. The increase continued up to the 77th day for SD-1
and SD-2 and the peak concentrations were determined as 1012.73±64.27 µg/L for
SD-1 and 1157.68±65.66 for SD-2. After this day, NP1EO concentration started to
decrease in a pattern similar to that of NP2EO for these digesters. The increase in
NP1EO on the spike day is well-correlated to the increase in NP2EO but the
increment is not at the same level. It means that some of the NP2EO was
transformed into NP1EO and this transformation led to increase in the concentration
of NP1EO at the spike day and in the following 20-25 days. Also, at the same time
some of NP2EO or NP1EO was converted into NP. The consequent decrease of
NP1EO observed after the 77th day is due to the possible transformation into NP. As
can be seen from the graphs, NP concentration increased following NP2EO spike
into digesters and it reached the highest concentration at the 81st day as
769.93±64.40 µg/L for SD-1 and 774.79±55.50 µg/L for SD-2. After that, with
slight fluctuations, it nearly remained constant for 20 days and accumulated in the
system without further degradation. In a similar way, Hernandez Raquet et al.
operated a lab-scale batch methanogenic reactor treating urban sewage sludge at
30oC (Hernandez-Raquet et al. 2007). Their results showed that no degradation of
NP occurred in anaerobic environment. Also, they reported the estrogenic activity of
NP generated in anaerobic methanogenic reactor. In another study, Zhang et al.
studied anaerobic degradation of NP9EO mixture (15 mg/L) in upflow anaerobic
sludge blanket (UASB) system and reported the accumulation of significant amount
of NP in the anaerobic sludge (Zhang et al. 2008). They also emphasized the
presence of NP1-3EOs in anaerobic sludge at a considerable amount.
The slight fluctuations may also have resulted from daily feeding because the
concentrations of these compounds (background) change in the feed depending on
the sampling frequency from the WWTP. The samples for feeding were taken
weekly so the NP compounds introduced to the system as a background may differ
from week to week depending on WWTP operation conditions or weather
conditions. Therefore, biotic control digesters were set to monitor these
circumstances and compared with spiked digesters to understand the effects on
173
biodegradation of NP2EO. As general, it can be said that daily feeding did not
contribute any remarkable change in concentrations of NP compounds in following
days after spike in biotic control digesters so depending on this fact it can be claimed
that the noteworthy change in concentrations of NP compounds in spiked digesters
was caused by biodegradation of NP2EO externally added into digesters.
5.3.9.2. Liquid Phase of Spiked Digesters
As can be seen in Figures 5.13 and 5.14, all NP compounds were observed in the
liquid samples taken from the digesters, but were measured comparatively (to solids)
at very low concentrations in both digesters. The concentrations of all target
compounds fluctuated up until the spike day. At the spike day, following the spike, a
sudden increase from about 30 to 90 µg/L was observed in the NP2EO
concentration. Then a further gradual increase, from about 90 µg/L to about 130
µg/L followed between day 65 and 81. The concentration of NP2EO reached the
highest value at 81th day as 129.46±5.96 µg/L for SD-1 and 134.61±8.58 µg/L for
SD-2. Then NP2EO concentration started to decrease steadily in a similar way
exhibited in solid phases of these digesters. Again similarly, NP and NP1EO
concentrations increased following NP2EO spike. NP1EO concentration varied
between 50-90 µg/L for both digesters and kept a concentration of little above 50
µg/L at reactor termination. After minor fluctuations, NP remained constant at about
100 µg/L until the digester termination time.
174
160
200
180
140
160
NP2EO (µg/L)
140
100
120
80
100
80
60
60
NP&NP1EO (µg/L)
120
40
40
20
20
0
0
0
10
20
30
40
50
60
70
80
90 100 110 120 130 140 150
Time (days)
Figure 5.13. Change in NP compounds concentration (liquid phase) vs. time for SD1 (spiked digester)
175
NP2EO
160
NP1EO
NP
200
180
140
160
NP2EO (µg/L)
140
100
120
80
100
80
60
60
NP&NP1EO (µg/L)
120
40
40
20
20
0
0
0
10
20
30
40
50
60
70
80
90 100 110 120 130 140 150
Time (days)
Figure 5.14. Change in NP compounds concentration (liquid phase) vs. time for SD2 (spiked digester)
Results from the spiked digesters confirmed that the NP compounds concentrate on
solid surfaces due to hydrophobic properties as also observed in some previous
studies (Ejlertsson et al. 1999, Schröder 2001, Scrimshaw and Lester 2002).
When biotic control and spiked digesters are compared for the presence of NP
compounds in liquid phases, one can see that following spike, NP2EO concentration
increased to 92.52±7.48 and 98.97±8.93 µg/L for spiked digesters, whereas it stayed
as 35.95±2.41 and 34.38±2.98 µg/L for biotic control digesters. For spiked digester,
NP2EO continued to increase till the 81st day, then started to decrease and reached
59.48±4.83 and 62.59±5.94 µg/L concentrations at digester termination day. On the
other hand, the control digester showed fluctuations during the operation time and
exhibited a concentration of about 45 µg/L when the digesters were terminated.
Because of this difference, NP1EO and NP concentrations took different levels but
showed the same trends in the biotic control and spiked digesters.
176
5.3.10. Abiotic Control Digesters
While Figure 5.15 and Figure 5.16 represent the measured concentrations of the NP
compounds for solid phases of AD-1 and AD-2, Figures 5.17 and 5.18 demonstrate
concentrations of the NP compounds for liquid phases of AD-1 and AD-2,
respectively. The AD-1 and AD-2 digesters were autoclaved to remove the microbial
activity at the digester set-up and they were fed every day with the autoclaved WAS
(as a feed).
When Figures 5.15 and 5.16 are analyzed, nearly the same pattern was observed for
solid phases of both AD digesters. For 30 days following the set-up, a decrease was
observed with some fluctuations for all NP compounds. Due to lack of microbial
activity, this situation can be attributed to the daily feeding-sampling. The up-down
fluctuations were caused by the feed because the characteristics of the WAS used for
feeding were not the same all the time due to operational and seasonal changes.
These fluctuations can be deemed as acceptable operational problems. Also, the
changes were not drastic so it can be said that the taken sludge from the digesters
and given sludge as feed have the same concentrations of these compounds.
177
NP2EO
NP1EO
NP
400
3600
3300
350
3000
NP2EO (µg/L)
2400
250
2100
200
1800
1500
150
1200
NP&NP1EO (µg/L)
300
2700
100
900
600
50
300
0
0
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.15. Change in NP compounds concentration (solid phase) vs. time for AD-1
(abiotic control digester)
On the 63rd day, the abiotic digesters were spiked with 3000 µg/L NP2EO in
acetone. As can be seen from the Figure 5.15 and 5.16, NP2EO concentration
increased to 2849.14±165.12 µg/L for AD-1 and 2972.22±177.15 µg/L for AD-2. In
the following 10 days after spike, an increase was observed in the NP2EO
concentration (except for 67th day for AD-1 and 65th day for AD-2).
The NP
compounds available in the feed may have led to an increase in the concentration of
NP2EO. It can be said that NP2EO remained in the system nearly 2-SRT-period and
then with respect to time digesters began to be diluted due to the sampling-feeding
regime. After this point, NP2EO decreased gradually due to dilution effect up to
termination day. The digesters were terminated with concentrations of 710.40±88.46
µg/L and 659±62.37 µg/L NP2EO for AD-1 and AD-2.
178
NP2EO
NP1EO
NP
3600
400
3300
350
3000
300
2400
250
2100
1800
200
1500
150
1200
NP&NP1EO (µg/L)
NP2EO (µg/L)
2700
100
900
600
50
300
0
0
0
10
20
30
40
50
60
70
80
90
100 110 120 130 140 150
Time (days)
Figure 5.16. Change in NP compounds concentration (solid phase) vs. time for AD-2
(abiotic control digester)
In light of these findings, it is obvious that degradation of NP2EO was caused by the
microbial flora in SD digesters. NP2EO concentration for SD-1 and SD-2 digesters
decreased to 600.08±57.94 µg/L and 618.87±60.41, respectively within 20 days
following the spike, whereas for AD digesters only a small amount was removed,
likely the result of dilution and measured as 2359.48±184.12 µg/L for AD-1 and
2464.84±147.94 µg/L for AD-2.
As can be seen from Figures 5.17 and 5.18 belonging to AD digesters, all NP
compounds were measured at very low concentrations in liquid phases. At the spike
day, there was a small increase in the concentration of NP2EO due to the spike. The
increase in NP and NP1EO resulted from the WAS used for feeding. As stated
179
earlier, NP compounds concentrate on the solid surfaces due to their physicochemical properties so the level of the NP compounds is very low in liquid phases.
80
100
80
60
40
40
NP&NP1EO (µg/L)
NP2EO (µg/L)
60
20
20
0
0
0
10
20
30
40
50
60
70
80
90 100 110 120 130 140 150
Time (days)
Figure 5.17. Change in NP compounds concentration (liquid phase) vs. time for AD1 (abiotic control digester)
180
100
80
80
60
40
40
NP&NP1EO (µg/L)
NP2EO (µg/L)
60
20
20
0
0
0
10
20
30
40
50
60
70
80
90 100 110 120 130 140 150
Time (days)
Figure 5.18. Change in NP compounds concentration (liquid phase) vs. time for AD2 (abiotic control digester)
5.4. Conclusion
The study has focused on the operation of NP2EO spiked laboratory scale anaerobic
semi-continuous digesters to observe the degradation pattern and products of
NP2EO. Also, biotic and abiotic control digesters were operated to compare the
effect of NP2EO spike on the digester performance and degradation mechanisms in
anaerobic digesters.
The digester performance was monitored by pH, methane production, COD, TS/VS
and TSS/VSS analyses. pH of spiked digesters (SD-1/SD-2) and biotic control
digesters (BD-1/BD-2) were ranged between 7.0 and 7.6 throughout operation and
these pH values provided the environment for activity of microbial flora and the
degradation of NP compounds. Cumulative methane production with respect to
181
operation period revealed that NP2EO spiked digesters followed the same pattern
with control digesters (no NP2EO spike). It meant that microbial community
responsible for methane production was not affected with the spike of NP2EO and
then the presence of biodegradation products of NP2EO (NP1EO and NP) in digester
system. The spike of NP2EO (in acetone) at 63rd day led to the increase in COD
concentration. The same level of increase was also observed in biotic and abiotic
control digesters. With this observation, the increase in COD concentration was
attributed to common compound added into both systems, i.e. acetone. For spiked
and biotic control digesters, approximately 60% COD removal was observed. With
this finding, it was clear that addition of NP2EO did not affect digester performance.
Control and spiked digesters followed similar patterns and exhibited similar removal
efficiencies (nearly 52-53%) for volatile solids and volatile suspended solids.
Therefore, it seemed that the NP2EO spike had no negative effect on removal
efficiencies of VS/VSS in the digesters. In a similar way, biotic control and NP2EO
spiked digesters exhibited same trend (between 36-40%) for the reduction of TS and
TSS over the operation time.
It was observed that NP2EO concentration decreased from nearly 3000 µg/L to 200
µg/L in 84 days following the spike for SD digesters. It was observed that NP1EO
first increased and decreased as an intermediate compound, but NP accumulated. For
BD digesters, a decrease in the concentration of NP2EO (introduced by the daily
feed) was observed throughout digester operation. This decrease was accompanied
by the increase in concentrations of NP and NP1EO. Following the spike, no
remarkable change was observed in concentrations of nonlyphenol compounds. The
data obtained for spiked digesters in relation to biotic control digesters present the
mechanism of NP compounds degradation clearly.
At the digester set-up and all throughout the operation, no NP1EC was detected in
any of the digesters. This can be explained by lack of specific microbial flora
responsible for the transformation of NPnEOs into NPnECs in operated digester
systems or insufficient operation time for their formation. It can also be thought that
182
NPnECs are degradation products of aerobic biodegradation, not anaerobic as stated
by some studies in literature.
Due to hydrophobic nature of NP, great majority of NP2EO and its degradation
products were found in the solid fraction of sludge samples obtained from digesters.
The target compounds were measured at low concentrations in the liquid fraction.
Based on these findings, it can be said that under anaerobic conditions, NP tends to
be formed as the final product and its accumulation occurs during anaerobic sludge
digestion. This is unfortunate since NP is a chemical of critical concern due to its
high toxicity and estrogenic properties.
Abiotic degradation is negligible based on the obtained data so NP2EO degradation
was achieved by anaerobic microbial community and transformed into NP1EO and
NP without any contribution by abiotic reasons.
183
184
CHAPTER 6
ASSESSMENT OF ANAEROBIC MICROBIAL COMMUNITY
STRUCTURE IN THE PRESENCE OF NONYLPHENOL
DIETHOXYLATE
6.1. Introduction
Nonylphenol polyethoxylates (NPnEOs) represent a large portion of alkylphenol
polyethoxylates which are commonly used as non-ionic surfactants in many
industrial sectors. NPnEOs are produced in high volume due to being cost-efficient
with higher surface activity for use in household, industrial and agricultural
applications and as personal-care products (Talmage 1994). Following their wideuse, NPnEOs are released into aquatic environment mostly via wastewater treatment
plants (WWTP) where they are only partially degraded into shorter-chain
hydrophobic and more stable metabolites including nonylphenol diethoxylate
(NP2EO),
nonylphenol
monoethoxylate
(NP1EO),
nonylphenol
(NP)
and
nonylphenol carboxylates (NP2EC and NP1EC).
The toxic effects of shorter-chain NP compounds on aquatic organisms including
fish, algae and invertebrates have been reported. Due to their toxicity, a group of NP
compounds have been listed among the priority substance in the European Union
(EU) Water Framework Directive (WFD 2000). Besides the toxicity effect,
endocrine disruptor ability of NP, by mimicking the natural hormone 17β-oestradiol
(White et al. 1994, Lee and Lee 1996) is another critical problem to be addressed
about this chemical. These observations have raised a significant concern about the
long-term impact of these compounds on wildlife and human health.
185
Due to low solubility and high log Koc values, these metabolites, especially NP, tend
to accumulate on sewage sludge following anaerobic treatment. This situation
increase the concern about application of sewage sludge for agricultural purpose
with the possibility of introduction toxic, persistent and endocrine disrupting NP
compounds into the soil environment. With increasing concerns about NP
compounds, European Union suggested a limit value for the sum of NP, NP1EO and
NP2EO (called as NPE) as 50 mg/kg dry mass (dm) in Working Document on
Sludge, 3rd Draft (2000) for agricultural use of sewage sludge.
With these developments, and considering that anaerobic digestion is the most
commonly used and promoted sludge stabilization system worldwide, determination
of anaerobic microbial community structure in the presence of NP compounds is
important. It is critical to understand the influence of NP compounds on population
dynamics in anaerobic digesters and to investigate the contribution of the bacterial
and
archaeal
populations
on
biodegradation
of
these
compounds.
For
characterization of microbial community structure, conventional methods (culturebased) have some limitations including requirement of specific growth conditions,
need for selective media, lack of detection of viable but non-cultivable (VBNC)
microorganisms (Riviere et al. 2009, Manyi-Loh et al. 2013). Therefore, for reliable
characterization
of microbial
community structure, DNA-based
molecular
identification and quantification methods like fluorescence in situ hybridization
(FISH), denaturing gradient gel electrophoresis (DGGE) and quantitative PCR
(Leclerc et al. 2001, Sousa et al. 2007, Malin and Illmer 2008) have become favored
over conventional methods.
Anaerobic digestion is a well-established technology for mass reduction of waste and
sustainable energy generation via production of methane. This strategy sees a very
wide variety of application in regards to types of waste treated and reaction
conditions. We do have some degree of understanding of the complex processes in
these systems such as the major phylogenetic groups. However, anaerobic digesters
are still somewhat regarded as a black box as regards to the details of microbial
communities and the metabolic processes involved. In recent years, efforts have
186
been underway to gain a deeper and more thorough understanding of these systems
as discussed below.
It has been reported that Firmicutes, Proteobacteria, Bacteroidetes and
Actinobacteria are the dominant phyla in the bacterial community of the anaerobic
digesters (Chouari et al. 2005, Riviere et al. 2009, Ito et al. 2011, Lee et al. 2012,
Regueiro et al. 2012, Yu et al. 2014, Guo et al. 2015). Their main distribution was
reported at phylum level in different anaerobic digesters at different abundances.
This variation from study to study can be associated with difference in biomass
origin, influent characteristics and operational conditions. In this study, typical
sewage digesting digesters were operated and therefore when literature was surveyed
for this type of anaerobic digesters Proteobacteria were reported as the dominant
bacterial phylum of the entire Bacteria domain. For example, Guo et al. (2015)
performed the metagenomic sequencing to characterize microbial community
structure of the anaerobic digestion sludge from a full-scale WWTP and reported
that Proteobacteria, Firmicutes, Bacteroidetes, and Actinobacteria accounted for
41.2%, 12.5%, 9.6%, and 5.2% of all bacterial community, respectively.
Proteobacteria represent a massive level of morphological, physiological and
metabolic diversity and contribute tremendous significance to carbon, nitrogen and
sulfur cycling (Kersters et al. 2006, Spain et al. 2009). Proteobacteria are also
crucial phylogenetic group in anaerobic digestion process because most of Alpha-,
Beta-, Gamma-, and Deltaproteobacteria are reported as well-known glucose,
propionate, butyrate, and acetate-consuming microbial communities (Ariesyady et al.
2007, Guo et al. 2015). In other words, it makes this phylum an important
component in the degradation of organic matter in the digesters.
Recent advances in genome sequencing revealed that five bacterial strains are
capable of anaerobic degradation of aromatic compounds using different electron
acceptors (Carmona et al. 2009). It has been reported that these strains belong to
Alpha-, Beta- and Delta sub-groups of Proteobacteria (Larimer et al. 2004,
Matsunaga and Yasuhara 2005, Rabus et al. 2005, Butler et al. 2007, McInerney et
al. 2007). Also, it was found that benzoyl-CoA reductase is a key enzyme for
187
dearomatization of the benzene ring in the anaerobic degradation of aromatic
compounds and 21 strains belong to Alpha, Beta- and Deltaproteobacteria with BCR
genes detected in anaerobic samples (Carmona et al. 2009). In light of these findings
it could be said that Proteobacteria is a basic phylum that plays an important role in
degradation of aromatic compounds and organic pollutants. Therefore, in the
purpose of this study is to investigate, the distribution and abundance of four subgroups of Proteobacteria (Alpha-, Beta-, Gamma- and Delta-) in overall microbial
community in sludge samples using FISH and qPCR assays and to understand the
response and time dependent behavior of these sub-groups during the degradation of
NP2EO in anaerobic digesters. No study has been reported about the change in
abundance and diversity of Proteobacteria sub-groups in the presence of any NP
compounds in anaerobic digesters. Additionally, in order to understand the effect of
NP2EO and degradation products on methanogenic activity of digesters,
Methanosarcina and Methanosaeta acetoclastic methanogenic groups were
monitored in operated anaerobic digesters.
6.2. Materials and Methods
6.2.1. Laboratory-scale Anaerobic Digesters
Duplicate lab-scale semi-continuous anaerobic digesters were operated with 2L
sludge-volume as 3 sets (abiotic control, biotic control and spiked digester). The
digesters mixed continuously by magnetic stirrers were operated at 35oC in a dark
hot-room. 3000 µg/L NP2EO was spiked into one set of digesters (spkied digesters)
at 63rd day following the stabilization period and then degradation products were
monitored during operation time. Certain amount of sludge (133.4 mL) was wasted
from the digesters and the same amount of fresh activated sludge (feed) was added to
the digesters daily to maintain semi-continuous operation and to keep sludge
retention time (SRT) of 15 days. All digesters were connected to graduated gas
collection cylinders (4 L) to monitor gas production during operation.
188
Gas production and composition, total and total suspended solids (TS and TSS),
volatile and volatile suspended solids (VS, VSS) and chemical oxygen demand
(COD) were monitored during the operation-period of digesters. After 147 days of
operation (in total), digesters were terminated.
6.2.2. Sampling of Sludge for Microbial Analysis
6.2.2.1. Fluorescence in situ Hybridization
Initial studies for FISH application focused the preparation and homogenization of
samples due to presence of large aggregates of cells and flocs in sludge samples and
optimization of an analysis technique. The aggregates present in sludge interfered
with hybridization step in FISH analysis and deteriorated visualization quality. Also,
detachment of cells is very important to carry out accurate counting using
epifluorescence microscopy. In order to disaggregate flocs and detach cells from
particles/flocs, different methods including vortexing, shearing with a syringe and
needle and sonication have been applied (Biegala et al. 2003, Ishii et al. 2004, Lam
et al. 2004, Sekar et al. 2004, Thayanukul et al. 2010).
Sonication has been used commonly to disintegrate microbial cell aggregates and to
detach cells from particles in FISH studies (Weiss et al. 1996, Biegala et al. 2003,
Ishii et al. 2004, Lam et al. 2004). Ishii et al. (Ishii et al. 2004) reported that
practicing of sonication on soil samples with a sonication probe (at minimum power
for 20 s) improved FISH analysis for identification of bacteria and archaea in
sediments. In another study, Biegala et al. (Biegala et al. 2003) vortexed samples in
order to disaggregate cell clumps but vortex did not work for this purpose. When
sludge samples were sonicated (6 W with an amplitude of vibration of 20%) by 2
mm sonication probe, an improvement in hybridization efficiency was reported by
Bieagala and friends (Biegala et al. 2003). In another study, Araya et al. studied
microbial community structure in stream waters and biofilms by FISH and for
dispersion of cells, sonication step was applied for 2 min (125 W, 400 kHz) in a
sonication bath (Araya et al. 2003).
189
In highlight of studies in literature, shearing with syringe and needle and two
different sonication procedures were tested for dispersion of cells, detachment of
microbial cells from particles and disaggregation of microbial flocs. Ultrasonic bath
(Falc Instruments, 180 W) and ultrasonic homogenizer (Sartorious, LabSonicP,
France) with 3 mm-sonication probe were used for sonication process to compare
efficiency of these sonication procedures. For shearing, a sterile syringe with needle
(1 cc, 13 mm, insulin injector) was used and fixed samples were syringed up-down
for 100 times.
Sonication time and intensity are important parameters have to be optimized to
ensure that cells are detached from particles and dispersed without losing their cell
integrity. Sonication intensity is controlled by amplitude setting on ultrasonic
homogenizer. To prevent disruption of cells, the lowest amplitude (20%) was
selected and applied on fixed sludge samples for ultrasonic-probe application. Also,
sonication was carried out pulse-mode to prevent heat build-up. Following
determination of sonication intensity to be applied, trials were carried out to
determine optimal sonication time for ultrasonication bath and homogenizer. The
sonication times applied as 5, 10 and 15 min for ultrasonic bath and 5, 10, 20, 40 and
60 sec for 3-mm-sonication probe application. The disintegration of flocs and
dispersion of cells were evaluated via bright-field microscopic analysis. Fixed sludge
samples were sonicated in ultrasonication bath for 5, 10 and 15 min for dispersion of
cells. Following application of mentioned sonication times, sonicated fixed sludge
samples were analyzed under microscope. The examples for obtained microscopic
views are illustrated in Figure 6.1.
190
Figure 6.1. Sonication procedure in ultrasonic bath at different sonication times.
Bright-field microscope view, 40X objective.
As can be seen from Figure 6.1, following 5-min sonication in the ultrasonication
bath large aggregates/flocs shown at time t=0 (no sonication) were disintegrated into
smaller cell clumps but application of 5-min sonication was not enough to break all
cell aggregates present in samples. The application of 10-min sonication on fixed
sludge samples led to the decrease in size of aggregates. Further increase of
sonication time to 15 min did not lead to any remarkable improvement in size of
aggregates. Due to these results obtained ultrasonic bath was found not satisfactory
for sludge disaggregation purposes.
At this stage, fixed sludge samples were sonicated with 3 mm-sonication probe using
ultrasonication homogenizer at different sonication times ranging from 5 s to 60 s.
After application of these sonication times, the acquired microscopic images are
depicted in Figure 6.2.
191
With increase in sonication time from time 0 (t=0) to t=10 s dispersion of large flocs
was achieved but still presence of some large cell clumps revealed the need for
extension of sonication time. The increase in sonication time from 5 s to 10 s
improved the unity of aggregate size and dispersion of cell clumps. Further increase
in applied sonication time up to 20 s led to disintegration of nearly all large
aggregates in samples causing a uniform distribution of them. When applied
sonication time was increased from 20 s to 40 s and 60 s, microbial cells were no
longer apparent in the micrographs, indicating that cell lysis may have occurred.
Additionally, deterioration in uniformity of particle size at longer sonication times
occurred (Figure 6.2). As can be seen from Figure 6.2, an approximate uniform size
of aggregates with minimum cell lysis was observed between 10 s and 20 s.
As can be seen from Figures 6.1 and 6.2, sonication helped for disaggregation of
flocs and detachment of cells from particles in sludge samples. When obtained
images for different sonication procedures were compared, sonication using 3-mmprobe gave better results in terms of detachment of cells and flocs. Prolonged
sonication time (higher than 20 sec) led to disruption of the cells. In light of obtained
results the tubes containing fixed sludge samples kept in ice were treated by a 3-mm
probe for 15 sec at minimum power (with amplitude of vibration of 20%). With
application of sonication, homogeneous distribution of aggregates was provided by
breaking cell debris, inorganic particles and disintegration of flocs. This step
improved quality of images and decreased the background noise.
192
Figure 6.2. Sonication procedure with sonication 3-mm-probe at different time
periods. Bright-field microscope view, 40X objective.
Following the selection of dispersion method, sludge samples taken from anaerobic
digesters at predetermined time intervals were fixed by 4% paraformaldehyde in
phosphate buffer saline (PBS) at 4oC for overnight. The fixed sludge samples were
193
washed twice with PBS (137 mM NaCl, 2.68 mM KCl, 1.47 mM KH2PO4, 8.1 mM
Na2HPO4.7H2O, pH 7.4) and then sludge samples were sonicated by using
sonication probe (3 mm diameter, LabSonic P, Sartorious Corp., France) to disperse
cell flocs. Sonication was carried out at an amplitude of vibration of 20% for 15 s.
After sonication step, the samples were suspended in PBS:ethanol (1:1, v:v) and
stored at -20oC for FISH analyses.
Fixed samples (10 µL) were immobilized on gelatin coated glass slides and dried at
46oC for 10 min. The dried-slides were consecutively immersed in 50%, 80%, and
96% (v/v) ethanol series for 3 min each for dehydration. The hybridization buffers
(900 mM NaCl, 20 mM Tris-HCl (pH 7.4), 0.01% SDS and X% formamide) with
different concentrations of formamide (X values are presented in Table 6.1) for each
phylogenetic group were prepared. The hybridization buffer containing 30 ng/µL of
probe was applied to each slide. The hybridization was carried out at 46oC for 3 hrs
in a moisture chamber (by folding a piece of tissue and soaking it with the
hybridization buffer). Unbound probes were washed off by keeping slides in
preheated washing buffer (Y mM NaCl (depending on percentage of formamide used
in hybridization buffer, 20 mM Tris-HCl (pH 7.4), 5 mM EDTA (pH 8), 0.01%
SDS) at 48oC in water bath for 7 min. This step was repeated again in fresh washing
buffer and then slides were rinsed with double-distilled water (ddH2O). Following
air-dry, the slides were stained with 2 µl of DAPI (5 µg/mL, 4’,6-diamidino-2phenylindole, Merck, Germany) for 5 min in the dark, rinsed with ddH2O and airdried. The slides were mounted with DABCO (1,4-diazabicyclo[2.2.2]octane, Sigma
Aldrich, USA) or Citifluor (Citifluor Ltd., Canterbury, UK) prior to visualization
under microscope.
194
195
16S, 825–847
16S, 821–844
MX825
MS821
Methanosaeta
Methanosarcina
d
c
b
Most members of Deltaproteobacter
Concentration of NaCl in washing buffer.
Formamide concentration in the hybridization buffer.
Escherichia coli rRNA numbering.
16S, 915-934
23S, 1027–1043
GAM42a
Gammaproteobacteria
ARC915
23S, 1027–1043
BET42a
Betaproteobacteria
Archaea
16S, 968–986
ALF968
Alphaaproteobacteria
16S, 385-402
16S, 338–335
NON338
Negative Control
SRB385
16S, 338–355
EUB338MIX
Most Bacteria
Deltaproteobacteriad
Positiona
Probe
Specificity
CGCCATGCCTGACACCTAGCGAGC
TCGCACCGTGGCCGACACCTAGC
GTGCTCCCCCGCCAATTCCT
CGGCGTCGCTGCGTCAGG
GCCTTCCCACATCGTTT
GCCTTCCCACTTCGTTT
GGTAAGGTTCTGCGCGTT
ACTCCTACGGGAGGCAGC
GCWGCCWCCCGTAGGWGT
Sequence (5´……3`)
40
50
35
30
35
35
35
20
35
FA (%)b
Table 6.1. Oligonucleotide probes used for FISH analysis
46
18
70
102
70
70
70
215
70
NaCl (mM)c
(Raskin et al. 1994)
(Raskin et al. 1994)
(Stahl 1991)
(Amann et al. 1992)
(Manz et al. 1992)
(Manz et al. 1992)
(Neef et al. 1999)
(Wallner et al. 1993)
(Daims et al. 1999)
Reference
Slides were examined with a Carl Zeiss Axio Scope.A1 epifluorescence microscope
equipped with a 100 W high pressure mercury lamp (HBO 100) using the set of
filters DAPI and Cyanine 3 (Cy3). EC Plan Neofluar oil-immersion objective
(100X, 1.30 N.A.) was used for examination of slides.
Digital images of the slides were taken with Zeiss AxioCam MRm microscope
camera. Images were analyzed by using the DAIME Program (Daims 2009).
DAIME is a scientific image analysis and visualization program providing
examination
of
microscopic
images
of
microorganisms
obtained
by
FISH. “Biovolume-fraction” was applied for FISH analysis of samples using
DAIME program. Images were exported in *.tif format from AxioVision and
imported as image stacks into DAIME. Each image stack consisted of all replicates
for each dye and sample. Image stacks were paired; each sample was represented
by an image stack for DAPI and an image stack for population-specific probe.
Images were first background subtracted using the local backgrounding method
under default settings. Manual thresholding was then performed to remove
background noise. Automatic RATS-L segmentation was performed using default
settings. Image stacks were overlaid and automated artifact removal was performed
using a 75% congruency threshold, the default setting, which removes population
probe objects which are not represented in the DAPI image. Objects in the
population-specific image over 400 pixels were also considered as artifacts and
were rejected from further analysis. The images were then manually examined for
remaining artifacts, which were also removed. The resulting pairs of image stacks
were subjected to biovolume analysis, which in this application calculates the area
of population-specific objects as a fraction of DAPI-stained objects. The result was
taken as the average of the replicates.
In order to represent data as a relative quantity (relative abundance), the biovolume
(total measured area) of the target group was divided to biovolume (total measured
area) of total microbial community. Probe-bound microbial cells were normalized
relative to total cells stained by DAPI prior to biovolume fraction assays. In this
196
study, biovolume fraction was applied successfully to determine relative
abundances of different microbial phylogenetic groups in sludge samples.
All reported FISH data have been corrected for any non-specific binding to nontargeted cells and also for any autofluorescence effect caused by compounds
available in sludge samples (mainly humic acids) under the same filter set, by
subtracting the positive signals of the negative control NON338 (reverse
complementary probe of EUB338) of the same fluorochrome (Cy3) used in the
oligonucleotide probes.
Lists of oligonucleotide probes, formamide concentrations in hybridization buffer
and NaCl concentrations in washing buffer applied are shown in Table 6.1. The
probes were synthesized and labeled with fluorochrome Cy3 at the 5´ ends of the
oligonucleotide probes by Alpha DNA, Canada.
6.2.2.2. Quantitative PCR
6.2.2.2.1. DNA Extraction
For application of qPCR, total DNA have to be isolated from sludge samples. This
step is the key process for the success of investigation of microbial communities in
digester systems by molecular approach. In order to isolate genomic DNA from
sludge samples with high quality and yield, different commercial DNA extraction
kits were tested. In highlight of findings (given in Chapter 3, 3.5.1. DNA Extraction
part), E.Z.N.A. Soil DNA Extraction Kit (Omega Bio-Tek, USA) was selected
with the application of pretreatment step prior to extraction of total genomic DNA
from sludge samples. A 0.5 g (wet weight) of sludge suspended in 1 mL lysis buffer
of extraction kit was added into bead-beating tubes (Omega Bio-Tek, USA) and
incubated at 70oC for 10 min. Then, the tubes were placed on ice for 5 min. These
steps were repeated one more time following a short vortex-step and then the
sludge samples were homogenized with glass beads by using vortex adapter at
maximum speed for 20 min. After this point, manufacturer’s instructions were
197
followed. Extracted genomic DNA samples were quantified using UV-Vis
Spectrophotometer (Thermo Scientific, DR 2000 NanoDrop). The OD260/OD280
and OD260/OD230 values obtained for all total sludge DNA varied between 1.82.0 and 2.0-2.2, respectively.
Different dilutions of sludge DNA extracts in water were further used as templates
in qPCR assays with the taxon-specific primer sets.
6.2.2.2.2. Optimization of Quantitative PCR Assays
Quantitative PCR have been successfully applied for the determination of microbial
communities in various wastewater treatment processes under anaerobic conditions
(Harms et al. 2003, Lee et al. 2009, Yang et al. 2012). Therefore, qPCR is a useful
tool for sensitive detection and quantitation of microbial populations in
environmental samples. In this study, SYBR Green qPCR was performed for
detection and quantification of specific microbial groups in sludge samples.
TaqMan probe-based qPCR have also used in studies for detection and
quantification of microbial populations, due to its high specificity. However, the
need for the synthesis of different TaqMan probes for different target sequences is
the main drawback of this application. Also it becomes a costly method when a
wide range of microbial groups are to be studied. On the other hand, SYBR Greenbased qPCR does not need the development and synthesis of any additional probes.
Also, SYBR Green-based qPCR method is inexpensive and easily-applicable
compared to TaqMan probe-based qPCR.
Instead of SYBR Green dye, Eva Green was also used as a dye in qPCR reactions.
It has been proven that Eva Green is a better DNA binding dye than SYBR Green
with respect to several features. High concentrations of SYBR Green can inhibit the
PCR reaction especially in early stages of qPCR. On the other hand, Eva Green can
be used at high concentrations to increase fluorescence signal without PCR
inhibition. Also, Eva Green dye is more stable and much safer than SYBR Green
dye. The other important superiority of Eva Green over SYBR Green is that Eva
198
Green produces significantly stronger and brighter fluorescence signals with less
non-specific amplification than SYBR Green. Due to these benefits, Eva Green dye
was used in qPCR reactions.
The all taxon-specific primer pairs were selected from studies in literature to
determine the major phylogenetic groups in biotic control and NP2EO-spiked
anaerobic digesters. The list of taxon-specific primers and reference studies are
presented in Table 6.2. The selected primer sets have been tested for specificity in
reference studies (Table 6.2) previously. Therefore, in this study, the specificity of
qPCR reactions and primer pairs were further tested with DNA extracted from
bacterial strains representative of the target phylogenetic groups.
In order to determine the optimal annealing temperature, gradient PCR was carried
out for all taxon-specific primer pairs. Following setting of optimal annealing
temperature for each primer set (Table 6.2), qPCR reactions were performed in 96well plates using ABI 7500 Real Time PCR System (Applied Biosystems, USA). In
these analyses, genomic DNA of bacterial strains was used as a DNA template to
check specificity of primer-pairs as mentioned earlier. For Deltaproteobacteria,
Archaea, Methanosaeta and Methanosarcina groups, positive control bacterial
strains could not be obtained from any research laboratory in METU. Therefore,
total DNA were isolated from sludge samples obtained from anaerobic reactors
operated under sulfate reducing conditions and then used as a DNA template for
specificity trials of primer pairs used for Deltaproteobacteria. In a similar way, for
Archaea and methanogenic groups, sludge samples (in order to use a different
source) from mesophilic anaerobic digester of Tatlar WWTP were supplied to
perform experimental trials. Total DNA was extracted from these sludge samples
and was used for specificity trials of mentioned taxonomic groups. For all other
phylogenetic groups the trials were carried out with genomic DNA obtained from
bacterial strains. Strains of Sphingomonas sp. TTNP3 (representative bacterial
strain of α-Proteobacteria), Pseudomonas putida (representative bacterial strain of
γ-Proteobacteria) were obtained from Dr. Anthony Hay at Cornell University,
USA.
199
To check primer pairs of Bacteria domain and β-Proteobacteria in qPCR assays, E.
coli K12 and Burkholderia cepacia strains available in research laboratory of Dr. F.
Dilek Sanin were used.
All qPCR assays related to specificity indicated that primer pairs and qPCR
conditions (annealing temperature and primer concentration) given in Table 6.2
were specific for the target phylogenetic groups. Non-specific amplification were
not observed when qPCR products were evaluated by 1.0% agarose gel in TAE
buffer with GelRed (Biotium Corp., USA). The gels were visualized by Quantum
ST-4 3000 Gel Image Acquisition System (Montreal Biotech Inc., Canada). The
expected size of target PCR products (Table 6.2) was obtained following qPCR
reaction with taxon-specific primer sets.
200
201
Arc806
Sae835
Sar835
MS1b
Mb1b
Methanosaeta
Methanosarcina
Gam1202
Gam1080
Gammaproteobacteria
Arc349
Bet680
Eub338
Betaproteobacteria
Archaea
Alf685
Eub338
Alphaaproteobacteria
Del685
Eub518
Eub338
Bacteria
Del361
Primer
Primer
Deltaproteobacteria
Reverse
Forward
Target Group
R: AGA CAC GGT CGC GCC ATG CCT
F: CGG TTT GGT CAG TCC TCC GG
R: GAC AAC GGT CGC ACC GTG GCC
F: CCG GCC GGA TAA GTC TCT TGA
R: GGA CTA CVS GGG TAT CTA AT
F: GYG CAS CAG KCG MGA AW
R: ATC TAC GGA TTT CAC TCC TAC A
F: AAG CCT GAC GCA SCA A
R: CGT AAG GGC CAT GAT C
F: TCG TCA GCT CGT GTY GTG A
R: TCA CTG CTA CAC GYG
F: ACT CCT ACG GGA GGC AGC AG
R: TCT ACG RAT TTC ACC YCT AC
F: ACT CCT ACG GGA GGC AGC AG
R: ATT ACC GCG GCT GCT GG
F: ACT CCT ACG GGA GGC AGC AG
(5´ to 3`)
Sequence of primer pairs
266
266
475
324
170
360
365
200
(bp)
Amplicon size
60
60
50
54
57
60
60
54
( C)
o
Annealing Temp.
Table 6.2. Taxon-specific primers used in qPCR assays.
600
600
900
600
60
500
500
500
(nM)
Primer Conc.
(Shigematsu et al. 2003)
(Shigematsu et al. 2003)
2000)
(Takai and Horikoshi
(Stults et al. 2001)
et al. 2011)
(Bacchetti De Gregoris
(Fierer et al. 2005)
(Fierer et al. 2005)
(Fierer et al. 2005)
Reference
Following
determination
of
optimal
annealing
temperature
and
primer
concentration for each target group, qPCR amplifications were performed using
ABI 7500 Real Time PCR System (Applied Biosystems, USA) in polypropylene
96-well plates. The qPCR mixture of 25 µL was prepared using 12.5 µL of
EvaGreen Ssofast Supermix with Low ROX (Biorad, USA), X µL of each primer
(depending on target group, given in Table 6.2), 5.0 µL of template DNA (1 ng/µl)
and 5.0 µL sterile dH2O. The thermal cycling protocol applied was as follows:
initial denaturation for 3 min at 98oC followed by 35 cycles of 15 s at 98oC and 1
min at annealing temperature determined for each target group (Table 6.2).
After the amplification, a melting curve analysis was carried out with a temperature
increment of 0.5oC/s from 60oC to 95oC to confirm that the signal obtained in qPCR
originated from specific target PCR products and not from artifacts like primer
dimers or contaminants. QPCR amplifications were performed triplicate per DNA
sample and standards. The sequences of primers, their target groups, relevant
annealing temperatures and primer concentrations are listed in Table 6.2.
The target PCR product of primers specific for each phylogenetic group was
amplified by conventional PCR using the primer sets given in Table 6.2. In
conventional PCR, genomic DNA extracted from the relevant positive control
strain was used as a DNA template. The 50 µL reaction mixture contained the
following: 1xPCR buffer, 4 mM MgCl2, 400 µM dNTPmix, 500 nM of each
primer, 2.5 U of Taq DNA polymerase (Thermo Scientific, USA) and 50 ng of
genomic DNA template. PCR reactions were performed with Thermo Scientific
thermal cycler (Thermo Scientific, USA) according to following program: 30
cycles of 95°C for 1 min, ToC for 1 min, and 72oC for 1 min followed by a final
extension step of 10 min at 72oC. T (oC) represents the annealing temperature of
each qPCR reaction and related T value for each one is presented in Table 6.2.
The PCR products were cleaned by Gel and PCR Clean-up kit (Macherey-Nagel,
Germany) and then purified 16S rRNA PCR products specific for each target group
were cloned into pGEM®-T Easy vector system as described in the manufacturer’s
202
protocol (Promega, USA). Following insertion of each target 16S rRNA gene
product into pGEM®-T Easy vectors, they were transformed into competent cells
of E.coli JM109 (Promega, USA). 100-200 μl of each transformation culture was
inoculated on LB/ampicillin/IPTG/X-Gal agar plates. The plates were incubated at
37oC for overnight and then blue-white screening was carried out. Positive colonies
(white
ones
on
LB/ampicillin/IPTG/X-Gal
agar)
were
inoculated
onto
LB/ampicillin plates and cultured overnight at 37oC. Next day, single colony was
picked from each plate and inoculated into 5 mL of LB broth with ampicillin (100
mg/mL). The grown cell cultures in LB broth were used for plasmid isolation and
preparation of glycerol stock of cultures.
The recombinant plasmids were isolated using a Plasmid Miniprep Kit (Promega).
Quality and quantity of isolated plasmid DNAs were determined with UV-Vis
Spectrophotometer. The recombinant plasmids were stored at -20oC.
To test the isolated plasmids for positive recombination, a colony PCR was carried
out using insert-specific primers and vector primers (T7F/M13R). Then amplified
PCR products were confirmed for expected product size by agarose gel
electrophoresis.
10-fold serial dilution series of the recombinant plasmids ranging from 103 to 109
copies/µL were constructed representing each phylogenetic group. A linear
relationship between the log of the plasmid DNA copy number and the threshold
cycle (CT) value was observed for all recombinant plasmid standard curves
(r2>0.996). Amplification efficiencies were calculated by using the slope value of
each standard curve based on following formula (Pfaffl 2001):
203
The amplification efficiencies for all qPCR amplifications varied between 1.7 and
2.1 (90-110% as percentage) reported as acceptable range by (Pfaffl 2001).Standard
curves are shown in Appendix B.
Relative abundance for each target sub-group of Proteobacteria in each sludge
sample taken at different sampling days was determined by normalizing its
abundance with the abundance of total Bacteria in that sludge sample. Also,
relative abundances of methanogens represented by two groups including
Methanosarcina and Methanosaeta in the same sludge samples were assessed
following normalization of data with the abundance of total Archaea in sample of
interest. Also, in order to determine diversity of Proteobacteria groups and
methanogens in total microbial community, relative abundance of each
phylogenetic group was determined by normalizing with the abundance of total
microbial flora (Bacteria + Archaea) for each sampling day. Percent relative
abundance was used to represent data and determined within six target phylogenetic
groups.
The relative abundance of each target sub-group of Proteobacteria in each sludge
sample taken at different sampling days was calculated as the ratio between the
measured copy number for each group-specific qPCR assay and the “Bacteria”
assay. The relative Methanosaeta/Archaea ratio is the ratio of copy numbers
measured for the “Methanosaeta” and “Archaea” qPCR assays by using specific
primer pairs for these groups. For Methanosarcina; the relaive abundance was
calculated as the ratio between the measured copy number for Methanosarcina
qPCR assay and the “Bacteria” assay. The expression of target copy number as a
fraction of the total bacterial population provides a more precise index of
abundance since PCR amplification efficiency can vary between samples (Fierer et
al., 2005).
204
6.2.3. Analytical Methods
Extraction of NP compounds from sludge solids were carried out by 5 minsonication-assisted extraction with acetone. Solid phase extraction (SPE) method
using tC18 cartridges (Sep-Pak, Waters) was used for extraction of NP compounds
from liquid phases of sludge samples. For identification and quantification of NP
compounds, the extracts were derivatized (with BSTFA+TMCS (99:1)) and then
subjected GC/MS analysis. An amount of 1 µl of the sample was injected in the
splitless mode to GC/MS. The oven program used was as follows: 100oC for 5 min,
25oC/min to 160oC, 10oC/min to 260oC, 260oC for 5 min, 35oC/min to 285oC and
285oC for 7 min.
Biogas composition of the produced gas in the digesters were monitored by a
GC/TCD. The program was as follows: holding the column temperature at 45oC for
a minute and increase the temperature to 65oC by 10oC/min.
Total COD analysis was carried by applying dichromate oxidation method
approved by United States Environmental Protection Agency (US, EPA). TS, VS,
TSS and VSS solid analyses were performed according to Standard Methods
(APHA, WEF, 2005).
All details of analytical methods used in this study have been given in Chapter 3
and Chapter 4.
6.3. Results and Discussion
After an adaptation/stabilization period of seven weeks, biotic control and spiked
digesters started to exhibit constant performance indicating steady state situation.
With steady state in all digesters, spiked digesters were dosed with 3000 µg/L of
NP2EO at 63rd day. Digesters were operated for 147 days in total.
Analysis of the microbial population structure of these digesters was performed by
application of FISH and quantitative PCR molecular methods. qPCR is a sensitive
205
and reliable method to determine the abundance of DNA sequences that can be
used to estimate the distribution of microbial populations in the community. FISH
method is also useful molecular tool to provide information on the microbial
community
structure
by
targeting
different
phylogenetic
groups
using
oligonucleotide probes.
No significant differences in probe hybridization for any of the probes used were
found among replicate digesters. The sludge samples taken from the parallel
digesters exhibited similar microbial patterns, so this result proved that replicate
digesters were operated concurrently and under the same conditions. Because of the
low variability between replicates (< 20% for RSD), data were assessed together
and expressed as average to reveal the effect of NP2EO spike on the microbial
community composition.
6.3.1. Examination of Microbial Community Structure by FISH and qPCR
During the operation period, the anaerobic digesters demonstrated a good
performance in terms of solids destruction and COD removal efficiencies. pH
measurements indicated that the anaerobic digestion system functioned efficiently
without any disturbance. The average methane composition accounted for about
57.25±0.81% (average value for both digesters) of the biogas. These findings
revealed that addition of NP2EO into spiked digesters and acetone into biotic
control digesters did not upset microbial flora and their activities overall based on
measured operational parameters.
Fluorescence in situ hybridization (FISH) with rRNA-targeted oligonucleotide
probes (Table 6.1) were used to characterize and determine the composition of
microbial community structure in semi-continuous anaerobic digesters at certain
time intervals. For this purpose, seven different sludge samples taken from both
digesters at different days were analyzed using FISH method. One of the sampling
days was chosen prior to NP2EO spike. The other sampling days were determined
to be after the spike day. Due to 90% of NP2EO was degraded in 60 days following
206
to NP2EO spike, six different sampling days were chosen within the mentioned
time interval.
Relative abundance for each target sub-group of Proteobacteria in each sludge
sample taken at different sampling days was determined by normalizing its
abundance with the abundance of total Bacteria (EUB338) in that sludge sample.
Also, relative abundances of methanogens represented by two groups including
Methanosarcina and Methanosaeta in the same sludge samples were assessed
following normalization of data with the abundance of total Archaea (ARC915) in
sample of interest. Also, in order to determine diversity of Proteobacteria groups
and methanogens in total microbial community, relative abundance of each
phylogenetic group was determined by normalizing with the abundance of total
microbial flora (Bacteria + Archaea) for each sampling day. Percent relative
abundance was used to represent data and determined within six target phylogenetic
groups.
Some examples of FISH images used in semi-quantitative image analysis are
presented in Figures 6.3-6.11. The given FISH images represent overlays of DAPI
and Cy3-specific probe micrographs. In order to illustrate the overall process,
DAPI, EUB338MIX-Cy3 and the overlay of DAPI-EUB338MIX are illustrated in
Figure 6.3. DAPI is a cell-permeable fluorescent dye which binds to any doublestranded DNA in fixed cells and thus stains all microbial cells in sludge samples. In
the presented FISH images, the blue color indicates DAPI-stained microbial cells.
The red colored microbial cells indicate cells hybridized with specific probe
(labeled with Cy3). When the DAPI and EUB338MIX labeled Cy3 images are
overlaid, pink-colored microbial cells represent the positive target cells. Following
processing of data, relative abundance of target microbial groups was determined.
207
Figure 6.3. Epifluorescence photographs showing in situ hybridization for Bacteria
domain. a) DAPI-staining b) EUB338MIX-Cy3 c) Overlay of DAPI and
EUB338MIX-Cy3. (66th day, spiked digester), Bar = 5 µm, 6300X
208
Figure 6.4. Epifluorescence photographs showing in situ hybridization for Archaea
domain a) DAPI-staining b)overlay of DAPI and ARC915-Cy3 c) processed image
of overlay of DAPI and ARC915-Cy3 d) ARC915-Cy3. (103rd day, biotic control
digester), Bar = 5 µm, 6300X
209
Figure 6.5. Epifluorescence photographs showing in situ hybridization for Alphasubgroup. a) DAPI-staining b) overlay of DAPI and ALF968-Cy3 c) processed
image of overlay of DAPI and ALF968-Cy3 d) ALF968-Cy3. (83rd day, spiked
digester). Bar = 5 µm, 6300X
210
Figure 6.6. Epifluorescence photographs showing in situ hybridization for Betasubgroup. a) DAPI-staining b) overlay of DAPI and BET42a-Cy3 c) processed
image of overlay of DAPI and BET42a-Cy3 d) BET42a-Cy3. (66thday, spiked
digester). Bar = 5 µm, 6300X
211
Figure 6.7. Epifluorescence photographs showing in situ hybridization for Gammasubgroup. a) DAPI-staining b) overlay of DAPI and Gam42a-Cy3 c) processed
image of overlay of DAPI and Gam42a-Cy3 d) Gam42a-Cy3. (71st day, spiked
digester). Bar = 5 µm, 6300X
212
Figure 6.8. Epifluorescence photographs showing in situ hybridization for Deltasubgroup. a) DAPI-staining b) overlay of DAPI and SRB385-Cy3 c) processed
image of overlay of DAPI and SRB385-Cy3 d) SRB385-Cy3. (83rd day, spiked
digester). Bar = 5 µm, 6300X
213
Figure 6.9. Epifluorescence photographs showing in situ hybridization for
Methanosaeta genus. a) DAPI-staining b) overlay of DAPI and MX825-Cy3 c)
processed image of overlay of DAPI and MX825-Cy3 d) MX825-Cy3. (124thday,
control digester). Bar = 5 µm, 6300X
214
Figure 6.10. Epifluorescence photographs showing in situ hybridization for
Methanosaeta genus. a) DAPI-staining b) overlay of DAPI and MX825-Cy3 c)
processed image of overlay of DAPI and MX825-Cy3 d) MX825-Cy3. (124thday,
control digester). Bar = 5 µm, 6300X
215
Figure 6.11. Epifluorescence photographs showing in situ hybridization for
Methanosarcina genus. a) DAPI-staining b) overlay of DAPI and MS821-Cy3 c)
processed image of overlay of DAPI and MS821-Cy3 d) MS821-Cy3. (103rd day,
spiked digester). Bar = 5 µm, 6300X
As can be seen from Figure 6.12, Bacteria were the dominant domain in all the
sampled sludges for both digester systems. The average relative abundance of
Bacteria domain within control digesters was determined as 72.6%±3.5%. On the
other hand, Archaea domain was accounted for 23.1%±0.8% of total microbial
community according to FISH results. Similarly, for spiked digesters average
216
relative abundances for Bacteria and Archaea were determined as 71.2±4.2% and
22.8%±1.8%, respectively. Note that the total microbial counts take standard
deviations into account, i.e. the averages plus the standard deviations total to 100%.
Quantitative PCR (qPCR) was also used to determine the total abundance of
bacterial and archaeal 16S rRNA genes in each sample by targeting each domain
(Figure 6.12). Similar to FISH results, Bacteria was determined as predominant
domain of anaerobic microbial community in digesters according qPCR assays. The
obtained average relative abundances for Bacteria and Archaea in microbial
community of biotic control digesters were 74.2%±3.4% and 21.8%±0.6%
respectively. For spiked digesters, 73.5%±4.6% and 20.7%±1.2% abundances were
observed for Bacteria and Archaea domains.
When the studies reported in literature is examined, it is seen that the relative
abundance of Bacteria was reported as a dominant domain which makes up 70–
80% of the total microbial population of typical anaerobic digesters (digesting
sewage sludge), on the other hand, Archaea was represented by 20-30% in overall
community structure (Chelliapan et al. 2011, Regueiro et al. 2012, Guo et al. 2015).
Our results are in accordance with these results reported in literature.
217
BD digesters
SD digesters
Archaea
Archaea
Bacteria
Bacteria
0
20
40
80
60
0
20
Percentage (%)
40
60
80
Percentage (%)
Figure 6.12. Overall distribution of Bacteria and Archaea domains in operated
semi-continuous anaerobic digesters determined by FISH.
BD digesters
SD digesters
Archaea
Archaea
Bacteria
Bacteria
0
20
40
60
80
100
Percentage (%)
0
20
40
60
80
100
Percentage (%)
Figure 6.13. Overall distribution of Bacteria and Archaea domains in operated
semi-continuous anaerobic digesters determined by qPCR assays.
218
The abundance of Alphaproteobacteria populations was determined in control and
spiked digesters as given in Figure 6.14 at different digester operation times. The
percent abundance of α-Proteobacteria in the control digesters was similar to that
in the spiked digesters throughout the study period following the spike at 63rd day.
After the introduction of NP2EO into spiked digesters, a decrease was observed in
the abundance of α-Proteobacteria in biotic control and spiked digesters when
compared to their abundance on the 61st day (prior to spike). This situation could be
explained by the introduction of acetone added into both digester systems or could
be explained by experimental error.
6
Biotic Control Digester
Spiked Digester
Relative Abundance (%)
5
4
3
2
1
0
61
66
71
76
83
103
124
Time (days)
Figure 6.14. Relative abundance of Alphaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization using ALF968 oligonucleotide probe.
219
In both digesters, amount of α-Proteobacteria varied between 2.4%±0.2% and
3.1%±0.5% in 60 days of operation period after the spike of NP2EO or acetone into
the digesters. Depending on these results it could be concluded that the addition of
NP2EO into spiked digesters did not lead to any noteworthy change in the
abundance of α-Proteobacteria when compared to control digesters. The minor
fluctuations could be attributed to daily feeding/sampling. The bacterial strains
belong to α-Proteobacteria did not dominate in the system during transformation of
NP2EO into NP in the 60 day-operation period following the spike.
Similar trend for α-Proteobacteria was observed with qPCR assays. As illustrated
in Figure 6.15, the relative abundance values of both digesters for α-Proteobacteria
were nearly close to each other. The addition of acetone or NP2EO into digesters
did not lead any remarkable change. Therefore, it could be concluded that the
abundance of α-Proteobacteria determined at 61st day by FISH analysis may be
overestimated. The relative abundance of α-Proteobacteria with respect to
sampling times did not indicate notable enrichment of α-Proteobacteria in spiked
digesters in the presence of NP2EO.
220
6
Biotic Control Digester
Spiked Digester
Relative Abundance (%)
5
4
3
2
1
0
61
66
71
76
83
103
124
Time (days)
Figure 6.15. Relative abundance of Alphaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by qPCR
assays.
Betaproteobacteria was one of the more abundant bacterial classes (with about
10% abundance) in both digesters prior to spike. With the introduction of NP2EO
into the digesters, the abundance of β-Proteobacteria increased very significantly in
spiked digesters compared to the biotic control digesters as illustrated in Figures
6.16 and 6.17. After the NP2EO spike, in 20 days, relative abundance of βProteobacteria reached high levels compared to control digesters and prevailed in
the system. Its abundance approached to 20% at the 71st day and then decreased to
18.6±1.4% and 17.4±1.6% at 76th and 83rd days, respectively. In next 20 days, the
relative abundance of β-Proteobacteria decreased further to 11.4%±1.6%. This
decrease may be caused by the diminution in concentration of NP2EO and
221
accumulation of NP in the spiked digesters. The NP2EO concentration was around
340 µg/L at the 103th day and nearly 90% of NP2EO was degraded by microbial
community. So it could be argued that the approach of relative abundance of βProteobacteria in spiked digesters to those of in control digesters was caused by the
decline in amount of parent compound, NP2EO.
25
Biotic Control Digester
Spiked Digester
Relative Abundance (%)
20
15
10
5
0
61
66
71
76
83
103
124
Time (days)
Figure 6.16. Relative abundance of Betaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization using BET42a oligonucleotide probe
Similar to FISH results, Betaproteobacteria abundance increased following the
NP2EO addition as indicated by qPCR assays. Although a bit higher abundances
were determined for β-Proteobacteria during analysis, it is clear that qPCR results
222
are consistent with FISH results. As can be seen from Figure 6.17, abundance of βProteobacteria in control digesters remained around 9.5-10% during the operation
period so it could be considered that the typical abundance of β-Proteobacteria in
anaerobic digesters
to be in this range. The abundance of β-Proteobacteria
increased to 18.6%±1.2% at 66th day in the spiked digesters by approximately
doubling its abundance (prior to spike). In following days, similar trend was
observed and then after 40 days of the spike, the abundance of β-Proteobacteria
came down to the same level that in control digesters.
30
Biotic Control Digester
Spiked Digester
Relative Abundance (%)
25
20
15
10
5
0
61
66
71
76
83
103
124
Time (days)
Figure 6.17. Relative abundance of Betaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by qPCR
assays
223
The increase in abundance of Betaproteobacteria could be explained by their better
ability to
adapt NP2EO-amended environment.
The domination
of β-
Proteobacteria in NP2EO-enriched digesters indicated that bacterial strains
belonging to this sub-group of Proteobacteria may play an important role in
NP2EO degradation. The degradation ability of bacterial strains could be an
inherited-ability or plasmid-originated by horizontal gene transfer. Lozada and coworkers evaluated the effect of NP10EO on the bacterial diversity in lab-scale
activated sludge reactors (aerobic) by feeding reactors with synthetic wastewater
(Lozada et al. 2004). Bacterial community composition was determined using FISH
and β-Proteobacteria was reported as dominant subclass of Proteobacteria. On the
other
hand,
the
well-known
nonylphenol
degraders
(Sphingomonas
xenophaga Bayram, Sphingomonas sp. strain TTNP3) isolated from aerobic
environments belong to Alphaproteobacteria class. In this study, in an anaerobic
environment, α-Proteobacteria did not show any enrichment during the degradation
of NP2EO, indicating that different phylogenetic groups are involved in NP
compound metabolism under anaerobic conditions.
Soares et al. operated an aerobic packed-bed bioreactor continuously treating
contaminated water with tNP (technical nonylphenol) at temperatures between 5.5
and 15°C and the abundance of Proteobacteria sub-groups were analyzed by FISH
(Soares et al. 2006). In their study, β-Proteobacteria was reported as the most
abundant class at all three operation temperatures (5.5, 10 and 15°C) in reactors. In
light of these findings they claimed that β-Proteobacteria may have a specific role
in the degradation of nonylphenolic compounds.
Another Proteobacteria sub-group observed at higher abundances in the operated
anaerobic digesters was Gammaproteobacteria as illustrated in Figure 6.18 and
Figure 6.19. Relative abundance of γ-Proteobacteria in digesters was close to 8%
in both control and spiked digesters prior to spike. Then the abundance of γProteobacteria in sludge samples of spiked digesters rose to 12.4%±0.9% at 66th
day according to FISH results. After 8 days following NP2EO addition into spiked
digesters (71st day), γ-Proteobacteria abundance reached to 16.6%±1.3%, on the
224
other hand, the level of γ-Proteobacteria remained around 7.4% in control
digesters. Up to 83rd day (20th day of the spike) γ-Proteobacteria predominated and
its abundance was about the double of that in control digesters. At the 103rd and
124th days, relative abundance of γ-Proteobacteria in spiked digesters came down
to the same level with control digesters and remained around 7.5%.
20
Biotic Control Digester
Spiked Digester
18
Relative Abundance (%)
16
14
12
10
8
6
4
2
0
61
66
71
76
83
103
124
Time (days)
Figure 6.18. Relative abundance of Gammaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization using GAM42a oligonucleotide probe
Based on qPCR results, pretty much similar trend was observed for γProteobacteria. The abundance reached to 18.2%±1.3% at 71st day according to
qPCR results and then remained around 15% in the following days with minor
fluctuations. Following 40 days after the spike, the abundance values became in
225
alignment with that of control digesters. During the operation time based on the
data obtained at different days following to spike, any notable change in the relative
abundance of γ-Proteobacteria was not observed in control digesters as presented
in Figure 6.18 and 6.19. This indicated that the increase in the level of γProteobacteria in spiked digesters was triggered by the addition of NP2EO.
20
Biotic Control Digester
Spiked Digester
18
Relative Abundance (%)
16
14
12
10
8
6
4
2
0
61
66
71
76
83
103
124
Time (days)
Figure 6.19. Relative abundance of Gammaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by qPCR
assays
In most studies, bacterial strains including Acinetobacter, Aeromonas, Shewanalla,
Proteus,
Stenotrophomonas
and
Pseudomonas
species
belonging
to
γ-
Proteobacteria were reported as NPEO degraders (Barberio and Fani 1998, John
226
and White 1998, Salvadori et al. 2006, Soares et al. 2006). Barberio et al. claimed
that some of the genes involved in biodegradation of NPEOs may be localized on
plasmid molecule(s) (Barberio and Fani 1998). Therefore, the growth ability of
strains especially belonging to γ-Proteobacteria with NPEOs as the sole carbon and
energy source could be related to the presence of plasmids. Another study by the
same group (Barberio et al. 2001) reported horizontal gene transfer by plasmid in
some strains of γ-Proteobacteria providing the growth ability in the presence of
NP6EO and NP9EO. On the other hand, Lozado and co-workers studied bacterial
community structure in NP10EO amended reactors and reported the relative
abundance of γ-Proteobacteria in low numbers in all sludge samples (Lozada et al.
2004). Di Gioia et al. operated biofilm reactors in batch and continuous mode in the
presence of Igepal CO-520 (industrial mixture of NPnEOs with an average 5
ethoxylate group) and determined microbial community structure by FISH (Di
Gioia et al. 2009). The results revealed that the shift from batch to continuous mode
led to increase in the abundance of γ-Proteobacteria, which is accompanied by a
noteworthy decrease in α-Proteobacteria, Interestingly, the shift in reactor
configuration from batch to continuous negatively affected the level of αProteobacteria in the bacterial community. These studies were carried out in
aerobic environment, so it could be the reason why α-Proteobacteria remained in
low abundance values in anaerobic continuous digesters operated in our study.
Relative abundance of Deltaproteobacteria within control digesters ranged between
3.2%±0.4% and 4.0±0.3% at specific sampling times during the digester operation
as presented in Figure 6.20. A small increase in abundance of δ-Proteobacteria was
observed between 71st and 83rd days in spiked digesters and its abundance reached
to 4.3±0.4%. It is hard to attribute this change to the presence of NP2EO due to
observation of similar fluctuations in abundance of δ-Proteobacteria for control
digesters. The daily feeding/sampling may lead to these minor fluctuations.
Therefore, it could be said that the distribution of bacterial strains belonging to δProteobacteria in the spiked digesters was similar to that in the control digesters
227
and addition of NP2EO did not lead any remarkable change in abundance of δProteobacteria in spiked digesters according to FISH results.
When qPCR results given in Figure 6.21 were examined, less fluctuation was
observed for abundance of δ-Proteobacteria between control and spiked digesters.
On the other hand, no noticeable increase was seen in abundance of δProteobacteria in spiked digesters attributed to NP2EO presence.
6
Biotic Control Digester
Spiked Digester
Relative Abundance(%)
5
4
3
2
1
0
61
66
71
76
83
103
124
Time (days)
Figure 6.20. Relative abundance of Deltaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by
fluorescence in situ hybridization using SRB385 oligonucleotide probe
228
6
Biotic Control Digester
Spiked Digester
Relative Abundance (%)
5
4
3
2
1
0
61
66
71
76
83
103
124
Time (days)
Figure 6.21. Relative abundance of Deltaproteobacteria in total microbial
community of biotic control and spiked anaerobic digesters analyzed by qPCR
assays
Chang et al. observed degradation of NP and NP1EO by anaerobic microorganisms
from NP-acclimated river sediments (Chang et al. 2004, Chang et al. 2005, Chang
et al. 2007). Degradation of NP occurred at 30°C in a time range of 46.2 to 69.3
days. The experiments were also carried out under different anaerobic conditions
and the results showed that the degradation rates followed an order of sulfatereducing condition > methanogenic conditions > nitrate-reducing conditions. It was
also
reported
by
same
research
group
that
Bacillus
niacini, Bacillus
cereus and Acinetobacter baumanni were strains responsible for degradation of NP.
Although this group claimed the degradation of NP under anaerobic conditions, no
mechanism has been proposed by this group or any other researchers.
229
Deltaproteobacteria represents sulfate-reducing bacteria and as can be seen from
Figures 6.20 and 6.21, no notable increase was observed in the abundance of δProteobacteria following the addition of NP2EO into spiked digesters.
It is known that hydrogenotrophic methanogenesis and acetoclastic methanogenesis
are the two main pathways of methane formation in anaerobic digestion process.
Acetoclastic methanogenesis is considered to account for the majority of methane
production (70%) in anaerobic digestion processes (Speece 1996). Therefore, in
order to understand the effect of NP2EO addition on methanogenesis, it is
important to monitor populations of Methanosaeta and Methanosarcina which are
the two main genera of acetoclastic methanogens in operated digesters.
When the distribution of methanogen populations in the biotic control and spiked
digesters were examined with FISH and qPCR assays (Figures 6.22-6.25),
Methanosaeta were determined as the dominant methanogen population throughout
the study period. The dominancy of Methanosaeta in both digesters is consistent
with the studies reported in literature (Raskin et al. 1994, Zheng and Raskin 2000,
McMahon et al. 2001, Ali Shah et al. 2014, van Haandel et al. 2014, Chen and He
2015). In typically operated anaerobic digesters under stable conditions,
Methanosaeta populations predominate over Methanosarcina populations due to
low acetate concentrations. This situation could be explained that Methanosaeta
members have lower maximum specific growth rates (µmax) and half saturation
concentrations (KS) for growth on acetate (Batstone et al. 2004, Conklin et al. 2006,
Demirel and Scherer 2008, Chen and He 2015). In this study, the digesters
representing typical anaerobic digestion of sewage sludge were operated so it could
be expected that both digesters had low acetate concentrations encouraging the
growth of Methanosaeta populations. As expected, Methanosaeta dominated in
both digester systems based on Archaea domain. (Figures 6.26 and 6.27).
230
20
Biotic Control Digester
Spiked Digester
18
16
Relative Abundance (%)
14
12
10
8
6
4
2
0
61
66
71
76
83
103
124
Time (days)
Figure 6.22. Relative abundance of Methanosaeta in total microbial community of
biotic control and spiked anaerobic digesters analyzed by fluorescence in situ
hybridization using MX825 oligonucleotide probe
231
20
Biotic Control Digester
Spiked Digester
18
Relative Abundance (%)
16
14
12
10
8
6
4
2
0
61
66
71
76
83
103
124
Time (days)
Figure 6.23. Relative abundance of Methanosaeta in total microbial community of
biotic control and spiked anaerobic digesters analyzed by qPCR assays
Dominancy of Methanosaeta in typical anaerobic digesters also indicates the
stability and sustainability of anaerobic digestion process because Methanosarcina
become more prevalent than Methanosaeta in unstable anaerobic digestion
processes with high acetate concentrations caused by organic acids accumulation
(Conklin et al. 2006, De Vrieze et al. 2012, De Vrieze 2014, Chen and He 2015).
So, the obtained data by FISH and qPCR assays also indicated that both digesters
were operated under stable conditions with dominancy of Methanosaeta
acetoclastic methanogens.
232
8
Biotic Control Digester
Spiked Digester
7
Relative Abundance (%)
6
5
4
3
2
1
0
61
66
71
76
83
103
124
Time (days)
Figure 6.24. Relative abundance of Methanosarcina in total microbial community
of biotic control and spiked anaerobic digesters analyzed by fluorescent in situ
hybridization using MS821 oligonucleotide probe.
233
7
Biotic Control Digester
Spiked Digester
Relative Abundance (%)
6
5
4
3
2
1
0
61
66
71
76
83
103
124
Time (days)
Figure 6.25. Relative abundance of Methanosarcina in total microbial community
of biotic control and spiked anaerobic digesters analyzed by qPCR assays
Methanosaeta became comparatively less abundant in spiked digesters than in
control digesters, despite remaining the dominant acetoclastic methanogen group
(Figures 6.22 and 6.23). Similar situation was also revealed when the distribution of
Methanosaeta and Methanosarcina in archaeal community was determined as
presented in Figure 6.26 and 6.27. As can be seen from Figure 6.26, as average
77% of the archaeal community was composed of Methanosaeta in control
digesters, on the other hand, this group represented the 65.1% of the archaeal
community of spiked digesters. Similarly, according to qPCR results, 75.7% of the
archaeal community was represented by Methanosaeta in control digesters, on the
other hand, Methanosaeta species formed 62.5% of the archaeal community of
spiked digesters. The obtained abundance results for FISH and qPCR are close to
234
each other and indicates more or less the same picture. The decrease in abundance
of Methanosaeta was accompanied by an increase in level of Methanosarcina.
As can be seen in Figures 6.24 and 6.25, following the spike, the relative
abundance of Methanosarcina increased from about 2.5-3% to about 5%. This
behavior was not observed in the control digesters. This situation could be caused
by the addition of NP2EO into these digesters. Symsaris et al. studied the effects of
triclosan, diclofenac and nonyphenol on the methanogenic communities and
reported IC50 values for when a mesophilic WWTP sludge was used as an inoculum
(Symsaris et al. 2015). The values were given as 546, 35, and 363 mg/L for DCF,
TCS, and NP, respectively. They also performed FISH analysis and showed that
hydrogenotrophic methanogens were more resistant to the inhibitory effect of these
compounds than acetoclastic methanogens. Bozkurt and Sanin investigated the
toxicity of NP2EO at a concentration range of 1-30 mg/L on anaerobic
microorganisms by performing anaerobic toxicity assays (ATA) and reported that
any toxic or significant inhibitory effect was not observed on methane production
activity for studied concentrations of NP2EO (Bozkurt and Sanin 2014). Similarly,
in this study, addition of NP2EO into spiked digesters did not lead any notable
effect on methane production compared to control digesters. It can be said that the
accumulation of NP in the spiked digesters following biodegradation of NP2EO
may lead to increase in abundances of Methanosarcina and hydrogenotrophic
methanogens (data represented by other Archaea) due to being more resistant to
toxicity.
235
BD digesters
Methanosaeta
Methanosarcina
Other Archaea
0
20
40
60
80
100
Percentage (%)
SD digesters
Methanosaeta
Methanosarcina
Other Archaea
0
20
40
60
80
100
Percentage (%)
Figure 6.26. Overall distribution of Methanosaeta and Methanosarcina in Archaea
domain of biotic control and spiked anaerobic digesters analyzed by fluorescence in
situ hybridization.
236
BD digesters
Methanosaeta
Methanosarcina
Other Archaea
0
20
40
60
80
100
Percentage (%)
SD digesters
Methanosaeta
Methanosarcina
Other Archaea
0
20
40
60
80
100
Percentage (%)
Figure 6.27. Overall distribution of Methanosaeta and Methanosarcina in Archaea
domain of biotic control and spiked anaerobic digesters analyzed by qPCR assays.
It has been reported that Methanosarcina are more resistant to high concentrations
of inhibitory compounds like ammonia, hydrogen sulfide and volatile fatty acids
(Demirel and Scherer 2008). As can be seen from Figures 6.26 and 6.27,
Methanosarcina abundance in spiked digesters was determined as 19.9% and
17.5%, respectively and when compared with control digesters, there is a notable
increase in the abundance of Methanosarcina species. As discussed in Chapter 5,
NP concentration in spiked digesters reached the highest concentration around 750
237
µg/L (at time period between 81 and 85 days) and then remained at around 600
µg/L up to end of the digester operation period. It is obvious that the NP was not at
high concentrations leading to toxicity on microbial community but it led to nearly
10% decrease in abundance of Methanosaeta without changing dominancy of this
genus and performance of spiked digesters.
6.4. Conclusions
This study demonstrated that molecular techniques, FISH and qPCR, are useful
tools in monitoring the effect of NP2EO addition on bacterial and archaeal
composition of anaerobic digesters. After NP2EO spike into anaerobic digesters,
FISH and qPCR assays revealed that NP2EO addition promoted a considerable
shift in certain sub-groups of Proteobacteria and archaeal community composition
of the anaerobic digester sludge.
The composition of the bacterial community structure in the NP2EO spiked
digesters
exhibited
a
high
abundance
of
Betaproteobacteria
and
Gammaproteobacteria compared to other sub-groups of Proteobacteria and control
digesters. It can be said that the members of Betaproteobacteria and
Gammaproteobacteria are better at adapting sudden changes in the environment
and may play a significant role in degradation of NP compounds.
All of the digesters were dominated by acetoclastic Methanosaeta species. With the
introduction of NP2EO into anaerobic digesters, NP2EO was degraded in 60 days
followed with some accumulation of NP1EO and total accumulation of NP in the
system. During this period, in spiked digesters, Methanosarcina populations
increased by 10% (relative increase) compared to control digesters. Although
Methanosaeta dominated throughout all operation period in all digesters, the
increase in the abundance of Methanosarcina in spiked digesters was attributed to
the addition of NP2EO. It can be said that NP2EO addition and accumulation of NP
with time as a last degradation product led to a noticeable shift in abundance of
Methanosarcina without leading any disturbance in digester performance.
238
Further research could be performed for detection of bacterial strains responsible
for degradation of NPEOs under anaerobic conditions at species level. Also,
biodegradation experiments can be carried out with these strains for degradation of
NP which is the recalcitrant compound of anaerobic degradation. The explanation
and understanding of catabolic pathways followed by these strains under anaerobic
environments enable us to provide enrichment of these microorganisms in
anaerobic digesters and NP-contaminated areas for solution.
239
240
CHAPTER 7
ISOLATION AND CHARACTERIZATION OF NOVEL
BACTERIAL STRAINS CAPABLE OF DEGRADATION OF
NONYLPHENOL ISOMERS
7.1. Introduction
Nonylphenol (NP) is a well-known pollutant primarily introduced into the
environment
through
the
biodegradation
of
nonylphenol
polyethoxylates
(NPnEOs). NPnEOs, which are produced by the ethoxylation of technical grade NP
(tNP), are commonly used as non-ionic surfactants in the formulations of paints,
food packages, personal care products, pesticides and many industrial applications
in large amounts worldwide (Talmage 1994, Guenther et al. 2002, Ying et al. 2002,
Montgomery-Brown and Reinhard 2003). Nonylphenol is a persistent end-product
of microbial degradation and accumulates in environmental compartments such as
sediment, soil, sludge and tissues of organisms. It has become known that NP is
toxic towards aquatic organisms and invertebrates (Jobling and Sumpter 1993,
Jobling et al. 1996, Toppari et al. 1996, Gray and Metcalfe 1997) and functions as
an endocrine disrupter, which leads to reproductive problems and disorders in fish,
animals and humans. The increasing concerns regarding NP have motivated
researchers to determine microorganisms that are responsible for degradation of
these toxic and estrogenic compounds.
241
Rapid biodegradation of NP with a linear alkyl chain (4-n-NP) by Candida
aquaetextoris has been reported (Vallini et al. 2001). However, the tNP used in
manufacturing of NPnEOs is composed of more than 100 isomers which differ in
the structure and the position of the alkyl chain attached to the phenol ring makes
the degradation complicated (Ieda et al. 2005). 85% of these isomers possess a
quaternary α-carbon on the branched alkyl chain which prevents β- and ω-oxidation
of nonyl chain (Van Ginkel 1996). Thus, biodegradation of branched isomers of NP
is different and more difficult than linear-chained molecule.
Given the ubiquity of branched NP following ultimate aerobic and/or anaerobic
degradation of NPnEOs by mixed microbial flora in sediments, soils, wastewaters,
activated sludge, studies have been focused on investigation of pure cultures
capable of using NP with a branched alkyl chain as the sole carbon and energy
source. Several studies described the isolation of bacterial strains able to grow on
and degrade branched NP in pure culture. The first such strain was Sphingomonas
sp. strain TTNP3, which is able to grow on tNP (as a sole carbon source) as
reported by Tanghe et al. (Tanghe et al. 1999). Tanghe and co-workers
demonstrated preferential degradation of para-NP isomers (p-NP, substituents at
the 1 and 4 positions on the benzene ring) over ortho-NP isomers (o-NP,
substituents at the 1 and 2 positions on the benzene ring). Then, Sphingomonas
cloacae (Fujii et al. 2001) and Sphingobium xenophagum strain Bayram (Gabriel et
al. 2005) were isolated from wastewater treatment plants by a similar method to
isolate Sphingomonas sp. strain TTNP3. Only one strain, Sphingobium amiense,
has been isolated from river sediments with an ability to degrade branched NP in
the presence of organic nutrients (diluted nutrient broth ingredients) (de Vries et al.
2001, Ushiba et al. 2003) Soares et al. reported isolation of three psychrotrophic
bacterial strains including Stenotrophomonas sp., Pseudomonas mandelii and
Pseudomonas veronii capable of degrading tNP from contaminated soils (Soares et
al. 2006). On the other hand, another study showed that Pseudomonas spp. only
degrade the poly-ethoxy chains of NP compounds to produce NP2EO and no
further degradation occurs (Van Ginkel 1996). Similarly, Stenotrophomonas strain
242
Bc6 was reported as capable of degrading low-ethoxylated NPEO mixtures
(Salvadori et al. 2006). Watanabe et al. announced Pseudomonas spp. and
Acidovorax sp. as NP degraders but the NP used as the substrate for degradation
assays and experimental conditions were not detailed enough to ascertain them as
branched NP degraders (Watanabe et al. 2011).
The initial biodegradation studies were performed using tNP as a sole carbon and
energy source to understand degradation mechanisms. These studies were
weakened by the fact that tNP is a complex isomeric mixture, leading to formation
of many degradation metabolites which makes difficult to monitor and understand
the biodegradation mechanisms. The development of a method to synthesize single
isomers of NP overcame the difficulty. Particular isomers (i.e. p353NP, p262NP
and p363NP) synthesized by Friedel-Crafts alkylation (Vinken et al. 2002) started
to be used as substrates for degradation studies (Corvini et al. 2004, Gabriel et al.
2005, Corvini et al. 2006, Gabriel et al. 2007, Gabriel et al. 2008).
The biodegradation studies with single isomers of NP helped to characterize the
metabolic pathway for degradation of NP. It was revealed that Sphingomonas sp.
strain TTNP3 and Sphingobium xenophagum Bayram are able to degrade several
branched isomers of nonylphenol by a type II ipso-substitution mechanism.
Hydroquinone has been reported as a key central metabolite during the
biodegradation of NP (Corvini et al. 2004, Gabriel et al. 2005, Corvini et al. 2006,
Gabriel et al. 2007, Gabriel et al. 2008, Kolvenbach and Corvini 2012).
Hydroquinone
is
further
degraded
to
4-hydroxymuconic
semialdehyde,
maleylacetic acid and 3-oxoadipic acid by Sphingomonas sp. strain TTNP3.
Hydroquinone dioxygenase (HQDO) has been isolated from Sphingomonas sp.
strain TTNP3. HQDO catalyzes the ring fission of hydroquinone to 4hydroxymuconic semialdehyde (Kolvenbach and Corvini 2012). The same research
group revealed a gene cluster in Sphingomonas sp. strain TTNP3 and demonstrated
the three gene products, HqdC, HqdD, and HqdE, responsible for 4hydroxymuconic semialdehyde dehydrogenase, maleylacetate reductase, and
intradiol dioxygenase activities, respectively (Kolvenbach and Corvini 2012).
243
All of the well-known bacterial strains with ability to degrade branched NP belong
to genus Sphingomonas of class Alphaproteobacteria. A solid environmental
system such as soil, sediment or sludge harbors a highly diverse bacterial
community
including
to
different
genera
of
Proteobacteria.
Although
Betaproteobacteria and Gammaproteobacteria have been presented as predominant
sub-groups in biodegradation studies with NPnEOs (Lozada et al. 2004, Soares et
al. 2006, Xin et al. 2008, Gu et al. 2010), only one study has reported isolation of
bacterial strains belonging to these groups, such Stenotrophomonas sp. and
Pseudomonas sp., with ability to degrade branched NP (Soares et al. 2006). Soares
et al. (2006) isolated these strains from contaminated soils and performed
biodegradation experiments at psychrotrophic temperatures using tNP as a
substrate. The isolation studies to date have been carried out from a variety of
different environments such as soil, sediment, wastewater and sludge samples or
from bioreactors operated at different conditions. Variation in location, type of
environment used for isolation, season, temperature and culture media used during
enrichment studies may lead to the isolation of different bacterial strains. Therefore,
the motivation of this study was to isolate phylogenetically novel bacterial strains
from activated sludge and investigate their degradation behaviors in the presence of
tNP and single isomers of p353NP and p363NP.
7.2. Materials and Methods
7.2.1. Chemicals
Technical grade nonylphenol (tNP, Pestanal) was purchased from Sigma-Aldrich
(Sigma Aldrich Co LLC, USA). 4-tert-butylphenol (≥99%) and 4-tert-octylphenol
(97%) were also supplied from Sigma-Aldrich. All solvents were high-performance
liquid chromatography (HPLC) grade and were supplied from Merck (Darmstadt,
Germany). All salts and medium components were also purchased from Merck
(Darmstadt, Germany). All restriction enzymes and buffers were supplied from
Promega (Promega, USA).
244
Two different NP isomers 4(3’,5’-dimethyl-3’-heptyl)phenol (p353NP) and 4(3’,6’dimethyl-3’-heptyl)phenol (p363NP) were synthesized by Friedel-Crafts alkylation
described by Vinken et al. (2002) and Ruβ et al. (2004). Chemical structures were
confirmed by GC-mass spectrometry (GC-MS) and by one- and two-dimensional
nuclear magnetic resonance (NMR) techniques. All details of isomer synthesis and
confirmation process are presented in Chapter 3 under Section 3.5.5.7.1.
7.2.2. Enrichment Studies
Mineral salts medium (MSM) was prepared according to the method of McCullar et
al. (1994) and used for enrichment and biodegradation assays of NP-degrading
isolates. Solid MSM plates were prepared by the addition of 16 g of noble agar into
a liter of MSM and then sterilized by autoclaving. Nonylphenol was added into the
media after autoclaving due to the heat-sensitivity of these compounds. Due to low
solubility of nonylphenol in the aqueous environment (4.9±0.4 mg/L at 25oC),
nonylphenol remained as droplets in the culture medium. During enrichment
studies, nonylphenol was added directly into MSM media using a sterile syringe.
Luria Bertani medium (LB, 10 g NaCl, 10 g tryptone and 5 g yeast extract per liter,
pH 7) was prepared to cultivate the isolates to high cell densities and to assess the
purity of the cultures. Bacterial isolates in LB medium were stored in 20% glycerol
stock at -80°C.
Activated-sludge samples (WAS) were obtained from the return sludge line of the
municipal WWTP in Ankara, Turkey to perform enrichment studies. 250 mL of
WAS sample was introduced into sterile 500 ml-Erlenmeyer flasks and then spiked
with 1 mg/mL of tNP. The flasks were placed into a dark shaking-incubator at
28±2oC and 120 rpm for 10 days. Following 10 days, 1 ml of enrichment culture
was transferred into 250 mL-Erlenmeyer flasks containing 100 ml of sterile MSM
with 1 mg/mL of tNP. From then on, the cultures were transferred successively
three more times into MSM medium containing 1 mg/mL tNP using identical
growth conditions at each transfer.
245
Growth was evaluated by measuring optical density (OD) at 550 nm using a HachLange DR 3900 Spectrophotometer (Hach Lange GmbH, Germany). The
enrichment culture of the fourth transfer was serially diluted (in sterile 0.85%
NaCl) and plated on tNP-MSM agar plates. Colonies exhibiting different
morphologies on agar plates were re-streaked on LB agar plates and colonies with
different morphologies were determined.
7.2.3. Identification and Characterization of Bacterial Isolates
To extract the genomic DNA of bacterial isolates for 16S rRNA gene amplification,
bacterial isolates were inoculated into 1/10 diluted LB medium and incubated at
28±2oC overnight in a shaking incubator. 1.5 mL of bacterial culture grown in LB
was centrifuged for 10 min at 4000g at room temperature. Following suspension of
bacterial cultures in TE (Tris-EDTA) buffer, genomic DNA of the isolates was
extracted using OMEGA E.Z.N.A. Bacteria DNA kit. The yield and purity of
genomic
DNA
of
bacterial
isolates
were
determined
using
UV-Vis
Spectrophotometer (Thermo Scientific, DR 2000 NanoDrop). Then the 16S rRNA
gene of the isolated bacteria was amplified from genomic DNA by using the
universal primers 27F and 1492R (Lane 1991). After amplification of 16S rDNA,
PCR products were digested separately with HaeIII and HhaI (Promega, USA)
restriction enzymes to compare 16S rDNA restriction patterns of isolated bacterial
strains. The procedure applied for amplification of genomic DNA and restriction
enzyme analysis are presented in Chapter 3 section 3.5.5.5.
Based on the restriction results, the isolates with different restriction patterns were
named as Isolate FKM-3, FKM-6, FKM-7, FKM-9 and FKM-11. 16S ribosomal
DNA sequence analyses of both strands for these isolates were carried out by BGITech (BGI-Tech Corp., Denmark) using an ABI 3700 XL automatic sequencer. The
obtained 16S rRNA gene sequences were subjected to BLAST (Basic Local
Alignment Search Tool) analysis (Altschul et al. 1990) and aligned with the
bacterial gene sequences in the GenBank (Benson et al. 2009). Phylogenetic trees
246
based on the 16S-rDNA sequences were generated by the Clustal W program using
a neighbor-joining method (Thompson et al. 1994).
The 16S rDNA sequences determined in this study have been deposited in the
GenBank database (NCBI) with following accession numbers: KP143085 (FKM3), KP143086 (FKM-6), KP143087 (FKM-7), KP143088 (FKM-9), and KP143089
(FKM-11).
Cell morphology of the isolates was determined by observing DAPI stained cells
under an epifluorescence microscope (Zeiss, Axioscope, Germany). Biochemical
properties of the isolates were determined by standard biochemical tests (Smibert et
al. 1994).
7.2.4. Biodegradation Assays
In biodegradation assays, unless specified otherwise, nonylphenol dissolved in
ethyl acetate was introduced as the sole carbon source into empty 100 mL-sterile
pyrex bottles. Ethyl acetate was evaporated in a sterile laminar flow cabinet.
Following removal of solvent from Pyrex bottles, sterile MSM medium (20 mL)
was added into each bottle. After inoculation with a bacterial isolate of interest, the
bottles were placed on a shaker at 28±2oC and 120 rpm. Light was blocked to
prevent photo-degradation of nonylphenol. After inoculation, the screw caps of the
bottles were kept loose to allow air intrusion into culture medium. Control bottles
were also set with and without nonylphenol in culture medium to ensure
disappearance of nonylphenol was caused by the bacterial isolates.
To monitor the disappearance of nonylphenol in biodegradation assays, identical
bottles were set as parallel and each analysis day, two bottles were sacrificed for
growth determination and NP analysis following extraction of the whole culture.
247
7.2.5. Extraction and Quantification of NP Compounds
At certain intervals, two culture bottles were sacrificed to monitor biodegradation
of nonylphenol by the isolate of interest. Up to analysis day, the bottles were frozen
and kept at -20oC. After thawing, the cultures were acidified by addition of 300 µL
of 1 N HCl. Then, 40 mL of ethyl acetate (extraction solvent) was added into
culture medium and vortexed at high speed vigorously for 5 min. Extraction
procedure was repeated with 20 mL of ethyl acetate at same conditions. The
obtained extract was passed through a sodium sulfate column to remove residual
water. The extract was concentrated under a gentle flow of N2 gas until dryness.
Then dry extracts were re-dissolved in 2 mL of ethyl acetate. Ethyl acetate extracts
were diluted properly with same solvent and then GC/MS analysis of extracts was
carried out.
4-tert-octylphenol was used as external standard in recovery studies. For
determination of extraction efficiencies, known concentrations of nonylphenol (1
mg, dissolved in ethyl acetate) and 4-tert-octylphenol (1 mg, dissolved in ethyl
acetate) were spiked into 20 mL of MSM medium prior to ethyl acetate-extraction.
The average recoveries for extraction of 4-tert-octylphenol and nonylphenol were
determined as 97.2±3.4% and 95.9%±2.2%, respectively.
4-tert-butylphenol was used as an internal standard during GC/MS analysis.
Internal standard solution of 4-tert-butylphenol was prepared with ethyl acetate as
(100 mg/L). The concentration of nonylphenol from cultures was calculated by
using internal standard method. Calibration curves were prepared in a range of 50200 mg/L for each compound with satisfactory correlation coefficients (r2 > 0.99).
The calibration curves are depicted in Appendix A.
The extracts were analyzed by GC-MS on a GC 6850 gas chromatograph (Agilent
Technologies, Santa Clara, USA) coupled to an MSD 5975C mass spectrometer
(Agilent Technologies). The injector was operated in splitless mode at 280°C.
Volume of 1µL was injected by a 7683B autosampler (Agilent Technologies).
Separations were performed on a HP-5MS capillary column (30 m long, 0.25-mm
248
internal diameter, and 0.25-µm film thickness; Agilent Technologies). The oven
program was: 50°C for 5 min, with 10°C/min to 280°C and held for 5 min. Helium
was used as a carrier gas, with a flow rate of 1.0 mL/min. The interface and source
temperatures were set at 230 and 150°C, respectively. Detection was performed in
the electron impact mode (70 eV). Following the determination of characteristic
m/z values for each compound in scan mode (m/z 50-400), GC/MS analyses were
carried out in SIM mode. The m/z values used for identification and quantification
of compounds used in this study are given in Table 7.1.
Table 7.1. m/z values used for identification and quantification of each compound
Compound Name
Target ions (m/z)
4-tert-octylphenol
41.1,77.1, 91.1, 96.0, 107.1, 121.1, 133.1, 206.2
4-tert-butylphenol
51.0, 65.0, 77.0, 95.0, 107.0, 119.0, 135.1, 150.1
tNP
107.0, 120.0, 135.0,149.1,163.1, 177.1, 193.1, 207.1, 221.1
p353NP
55.0, 77.0, 91.0, 107.0, 121.1, 135.1, 149.1, 191.1, 207.0, 220.2
p363NP
55.1, 77.0, 91.0, 107.0, 121.1, 135.1, 149.1, 191.1, 207.0, 220.2
7.3. Results and Discussion
7.3.1. Isolation and Characterization of Bacterial Strains
Enrichment studies were started with amendment of tNP (1mg/mL) into activated
sludge samples taken from the return sludge line of a municipal WWTP. After four
transfers into MSM-tNP (1mg/mL) medium with a time interval of 10 days, MSMtNP agar plates were inoculated from last-transferred MSM-tNP medium. In one
month, tiny colonies appeared on agar plates. Fifteen different looking colonies
from MSM-tNP agar plates were plated on LB agar plates. With repetitive plating
on LB agar, five different bacterial strains were distinguished based on difference
in colony morphologies. Restriction enzyme analyses carried out with HhaI and
249
HaeIII enzymes determined five different restriction patterns. In light of these
findings, five bacterial strains able to grow in the presence of tNP as the sole
carbon and energy source were isolated.
Control bottles with only MSM-tNP (no transfer from enrichment culture) and
without tNP (MSM and enrichment culture) did not exhibit any bacterial growth
during enrichment studies. Figure 7.1 demonstrates control and enrichment bottles
operated during enrichment studies with tNP.
Figure 7.1. Enrichment studies in the presence of tNP
250
With repeated plating on LB agar plates, individual colonies were picked and
cultured in LB medium. The bacterial isolates grown in LB medium were stored in
glycerol (20%) at -80oC.
The five bacterial strains capable of growing in the presence of tNP as only carbon
source were designated as isolate FKM-3, FKM-6, FKM-7, FKM-9 and FKM-11.
Based on 16S rDNA sequencing, these strains were determined to be Burkholderia
sp., Stenotrophomonas sp., Shinella sp., Ochrobactrum sp. and Citrobacter sp.,
respectively. The isolated bacterial strains on LB agar plates are presented in Figure
7.2 and their morphological and biochemical characteristics are presented in Table
7.2.
Figure 7.2. Isolated bacterial strains on LB agar plates
251
Table 7.2. Morphological and biochemical characteristics of the isolated bacterial
isolates
Parameter
FKM-3
FKM-6
FKM-7
FKM-9
FKM-11
smooth
smooth
smooth
smooth
smooth
Colony color
yellow
weak orange
cream
cream
cream
Gram reaction
negative
negative
negative
negative
negative
Cell morphology
rod-shaped
rod-shaped
rod-shaped
rod-shaped
rod-shaped
Motility
+
+
-
-
+
Catalase
+
+
+
+
+
Oxidase
+
-
+
+
-
Methyl-red
-
-
-
-
+
Citrate utilization
+
-
-
+
+
Colony
morphology
7.3.2. Biodegradation Tests
In order to show that the isolated strains were able to grow with tNP as the sole
carbon and energy source, MSM medium (50 ml) was supplemented with 1 mg of
tNP/mL and 500 µL of a week-old culture (OD550 = 0.45) grown in MSM-tNP
medium. The concentration of nonylphenol was monitored by GC/MS. Optical
density was monitored during biodegradation tests. The development of turbidy
indicating bacterial growth and the change of colorless tNP droplets to a yellow
color suggested that degradation was occurring. No turbidity or coloration was
observed in the abiotic controls. With time, yellowish-color disappeared and turned
into cream-colored droplets suspended in the MSM medium. 7th day culture flasks
are demonstrated in Figure 7.3. In the overall 21 day-operation time, remaining tNP
amount and optical density were determined at five different time points. As can be
seen from Figure 7.4, a decrease was observed in the amount of tNP following
addition into the medium. A remarkable diminution was noted in 2nd day when
252
compared to control data, but then no further consistent change in concentration of
tNP was observed.
Figure 7.3. Demonstration of MSM-tNP flasks inoculated with isolated bacterial
strains at 7th day of biodegradation assay
The same amount of tNP was extracted from control bottles and it did not change
with time. The concentration of tNP in control bottles in the overall experimental
period was 795.9±11.3 mg/L. The loss of 200 mg/L in the absence of bacterial
growth may have been caused by experimental error during addition of tNP or by
volatilization. Compared to control values, 18% of tNP was degraded by Isolate
FKM-3 in 21 days. 26%, 35%, 37% and 41% tNP was metabolized by isolates,
FKM-6, FKM-7, FKM-9, and FKM-11. The positive control strain Sphingomonas
253
sp. TTNP3 degraded 35% of tNP in 21 days. This low level activity shown by
TTNP3 is not consistent with what has been reported in literature; it was reported
that 98% of tNP was degraded by Sphingomonas sp. TTNP3 in 13 days (Tanghe et
al. 1999).
Figure 7.4. Degradation of tNP by bacterial isolates in MSM medium
The lack of further degradation of tNP after 2 days for all bacterial strains suggests
the formation of a toxic compound which prevents further degradation and growth
(Figure 7.4 and 7.5). Gabriel et al. indicated that although 96% of 4-NP112 (one of
single isomer of tNP) was degraded by S. xenophaga Bayram in 6 days, no colonies
formed on LB agar plates after the 4th day in 20 day-experiment (Gabriel et al.
2005). It is also difficult to monitor degradation of tNP by GC/MS due to the fact
that it yields a lawn of peaks (due to many isomers) and to the fact that each isomer
produces its own metabolic intermediates, which makes the chromatogram even
more complex and hard-to-read (Figure 7.6). Also, it has been revealed that each
isomer of tNP has different biodegradability and estrogenic activities (Gabriel et al.
2005, Corvini et al. 2006).
254
Figure 7.5. Degradation of tNP by bacterial isolates in MSM medium
Figure 7.6. Total ion chromatogram (TIC) belonging to the control sample at 7th
day. 4-tert-butylphenol: internal standard
255
In the literature, to avoid working with a complex mixture of NP (tNP) consisting
of more than 100 isomers (Ieda et al. 2005), single isomers such as p353NP,
p262NP and p363NP have been preferred as substrates for biodegradation
experiments (Corvini et al. 2004, Gabriel et al. 2005, Corvini et al. 2006, Gabriel et
al. 2008). In light of these findings and to carry out further biodegradation tests,
p353NP and p363NP were synthesized in the research laboratory of Prof. Dr.
Cihangir Tanyeli (Figure 7.7).
Figure 7.7. The synthesized and purified p353NP and p363NP
New biodegradation assays were prepared using the newly synthesized single
isomers as substrate rather than the technical mixture. For this purpose,
biodegradation experiments were performed in 100 mL-bottles with 10 mL sterile
MSM amended with 10 mg of target NP isomer. For each strain, two different sets
were organized with each single isomer. The MSM medium was inoculated with a
pre-culture of each strain grown in MSM supplemented with target isomer as the
sole carbon source. As mentioned earlier, on each analysis day, two identical
bottles were sacrificed to determine remaining amount of NP compound and optical
density.
256
In these analyses, degradation of single isomers was not observed by neither
isolated bacterial strains nor the positive controls Sphingomonas xenophaga
Bayram and Sphingomonas sp. TTNP3. This situation can be attributed to there
being not enough biomass for initiation of degradation. Therefore, after this point,
to save chemicals, the trials were performed using one of the isolated bacterial
strains. During growth studies, Isolate FKM-3 had yielded distinctive yellow-color
formation in the presence of p353NP and p363NP as illustrated in Figure 7.8;
therefore, this isolate was chosen for further degradation assays.
Figure 7.8. The synthesized and purified p353NP and p363NP
This time, biodegradation experiments were performed in 100 mL-bottles with 20
mL sterile MSM amended with 20 mg of p353NP and p363NP separately for each
set. Following 7-day operation, no remarkable disappearance of any of the NP
isomers was seen in the Isolate FKM-3 or Sphingomonas sp. TTNP3 cultures,
although growth was clearly observed during operation period. Total ion
chromatograms representing 1st, 3rd and 7th incubation-days for each NP isomer in
257
the presence of Isolate FKM-3 were overlaid and the obtained chromatograms are
illustrated in Figure 7.9. As can be seen from this figure, no notable decrease in the
amount of p353NP and p363NP during 7-day-operation time could be observed.
Corvini et al. reported that Sphingomonas sp. TTNP3 degraded 1000 mg/L p353NP
within 10 days (Corvini et al. 2004).
Figure 7.9. Overlaid total ion chromatograms obtained for p353NP and p363NP at
1st, 3rd, and 7th analysis days.
258
In light of these findings, it is obvious that the environment to stimulate the
degradation of these compounds could not be provided during the experiments. If it
were, positive control strains would have been able to degrade NP isomers as
reported in literature. Gabriel et al. (Gabriel et al. 2005, Gabriel et al. 2007, Gabriel
et al. 2008) reported the addition of very small amounts of yeast extract and
peptones (10 µg of each/ml) to provide essential vitamins and nitrogen into MSM
medium during biodegradation assays with Sphingomonas xenophaga Bayram. In
order to understand the reason why growth was not correlated with degradation,
yeast extract was added at minor level as described in the studies of Gabriel and coworkers.
The new set was prepared with 100 mL-bottles including 20 mL sterile MSM
amended with 10 mg of p363NP. All bacterial isolates and positive control strains
Sphingomonas xenophaga Bayram and Sphingomonas sp. TTNP3 were inoculated
to determine degradation of p363NP isomer. This time, all strains were induced in
the presence of tNP in MSM, then inoculated into MSM-p363NP (0.5 mg/mL)
medium. After reaching a certain optical density (≈ OD550 =0.5) 200 µL of
inoculum was introduced into each MSM- p363NP (0.5 mg/mL) medium.
This set was also operated for 7 days and as can be seen from Figure 7.10, the
bacterial isolates exhibited strong growth in the presence of p363NP as primary
carbon source and small amount of yeast extract (10 µg/mL).
259
Figure 7.10. Biodegradation test bottles belonging to positive control and bacterial
isolates (4th day of operation), p363NP as a sole carbon source
As can be seen from Figure 7.11, at the end of the 7-day operation, a small decrease
in p363NP concentration was observed for bacterial strains and no remarkable
change was observed for positive control strains. As illustrated in Figure 7.10,
Sphingomonas xenophaga Bayram and Sphingomonas sp. TTNP3 did not grow
well in the presence of p363NP. This could be attributed to inoculum introduced at
the beginning of assay. The main reason could be related to the low bacterial
biomass introduced into assays. To overcome these difficulties, in literature, resting
cell analyses are preferred. For both Sphingomonas strains, the data obtained in
later studies were carried out using resting cells rather than growing cultures
(Corvini et al. 2004, Corvini et al. 2006, Gabriel et al. 2007).
260
Figure 7.11. Biodegradation test bottles representing positive control and bacterial
isolates (4th day of operation), p363NP as a sole carbon source
7.4. Conclusions
Several presumptive NP-degrading isolates were obtained by enrichment from
WAS. These strains all exhibited what appeared to be growth in liquid NP MSM
medium, but there was no remarkable NP degradation as detected by GC/MS
analysis. Well documented NP-degrading strains Sphingomonas xenophaga
Bayram and Sphingomonas sp. TTNP3 used as positive controls also apparently
failed to degrade NP in these assays. The time spent on this part of the experimental
study was rather limited. Therefore, at this stage it is not reasonable to say that the
isolated bacterial strains in this study are not able to use NP isomers as the sole C
source. The apparent lack of NP degradation could be caused by the analytical
complexity of NP. For ex. GC/MS, which was used as the main analytical device in
our study, was not used in some of these monoculture studies. All studies with
Sphingomonas strains to date have utilized HPLC to monitor degradation of these
compounds, although the authors have not indicated the reason for their preference.
GC/MS was used only for identification of end-products only in these studies.
Additionally, in these studies, radioactively labeled-NP isomers (which can be
detected at very low concentrations) were used as a substrate to monitor
261
biodegradation. As mentioned earlier, resting cell and crude cell extract analyses
were used in these studies to overcome the problems of low biomass growth in
MSM leading to low activity levels. In future studies, the biodegradation assays
could be carried out using resting cells and crude cell extracts.
measurements should be performed using HPLC instead of GC/MS.
262
Additionally,
CHAPTER 8
CONCLUSIONS
This study was undertaken to accurately analyze and monitor the fate of toxic and
endocrine disrupting nonylphenol compounds in anaerobic digesters. Since many
limitations of the previous analytical methods suggested in the literature were
determined, a practical and reliable method for routine and simultaneous analysis of
NP, 4-n-NP, NP1EC, NP1EO and NP2EO in sludge and aqueous samples was
developed. After testing a number of different methods and method parameters, the
best extraction method for these compounds was found to be 5 min sonicationassisted extraction using acetone as the solvent. The recoveries achieved for the
target compounds were within the suggested values by USEPA and other studies.
The optimized extraction method showed good repeatability with relative standard
deviations (RSD) less than 6%. The limits of detection (LODs) were determined as
6 μg/kg for NP and NP1EO, 12 μg/kg for NP2EO, 0.03 μg/kg for 4-n-NP and 30
μg/kg for NP1EC. The developed method was successfully applied to dewatered
sewage sludge obtained from the Central WWTP in Ankara, Turkey for a year. This
part of the study also demonstrated the seasonal variability of NP compounds, such
that in winter, NPE concentrations slightly fluctuated over time but exhibited much
lower concentrations, whereas in summer the concentrations were observed to be
much higher. The sum NPE (NP+NP1EO+NP2EO) was found to be in between
5.5-19.5 μg/kg throughout the study, which is in compliance with Turkish and
proposed European regulations.
The main focus of the study was the operation of NP2EO (selected model
compound) spiked laboratory scale anaerobic semi-continuous digesters (SD) to
263
observe the degradation pattern and products of NP2EO. During operation period,
abiotic (AD) and biotic control digesters (BD) were also operated to compare the
effect of NP2EO-spike on the degradation mechanisms in the anaerobic digesters.
All digesters were operated for 147 days under mesophilic conditions. The BD and
SD exhibited the same pattern with regard to reduction of solids contents (TS, VS,
TSS and VSS). TS and TSS removal efficiencies of control digesters varied
between 36.80%-37.88% and 36.85%-38.42%, respectively. On the other hand, for
spiked digesters, removals varied between 38.01%-39.76% for TS and 38.6039.78% for TSS. When VS and VSS removal efficiencies were taken into
consideration, efficiencies changed between 51.93% and 53.50% for both digesters.
For COD removal, 58.6% reduction was achieved by control digesters (average of
BD-1 and BD-2) and 56.4% by spiked digesters (average of SD-1 and SD-2). All
solids and COD removal values are consistent with literature.
When methane production rates were taken into consideration, 0.376 and 0.388 LCH4/L·day were determined (as average) for BD and SD, respectively. The
methane yield ranged between 0.368-0.372 L-CH4/g-CODremoved for BD, on the
other hand, for spiked digesters this value lied between 0.383-0.385 L-CH4/gCODremoved. In light of these findings, it could be said that NP2EO addition did not
result in any negative effect on digester performance of spiked digesters.
Following steady state achievement, the spiked digesters were dosed with NP2EO
in acetone on the 63rd day of operation. In order to demonstrate whether addition of
acetone has some effect, same volume of acetone without NP2EO was dosed into
biotic control digesters at the same day. After acetone spike, NP, NP1EO and
NP2EO concentrations remained at the same levels (nearly as 60 µg/L for NP2EO,
160 µg/L for NP1EO, 360 µg/L for NP) up to the termination day in BD. The daily
feeding/sampling did not contribute to any remarkable change in concentrations of
NP compounds in the following days of spike in these digesters. In spiked
digesters, after the spike day, NP2EO concentrations in solid phase of the spiked
digesters started to decrease gradually and reached 600.07±57.94 µg/L at the 83rd
day and 269.80±20.53 µg/L at the 115th day for SD-1 and it came down to
264
618.89±60.41 µg/L at the 83rd day and 246.45±28.62 µg/L at the 115th day for SD2. About 90% of NP2EO spike was degraded in 60 days in both spiked digesters.
The great majority of NP2EO and its products were found in the solid fraction of
sludge samples for both BD and SD digesters, whereas NP compounds were
measured in liquid fraction at much lower concentrations. The findings of this study
showed that, biodegradation of NP2EO in spiked semi-continuous anaerobic
digesters was possible. Unfortunately, NP, which is a chemical of critical concern
due to its high toxicity and endocrine-disrupting ability, seemed to accumulate in
sludge as an end product of anaerobic degradation.
The average relative abundance of Bacteria domain within control digesters was
determined as 72.6%±3.5%. On the other hand, Archaea domain was accounted for
23.1%±0.8% of total microbial community according to FISH results. Similarly, for
spiked digesters average relative abundances for Bacteria and Archaea were
determined as 71.2±4.2% and 22.8%±1.8%, respectively. For qPCR studies, the
average relative abundances for Bacteria and Archaea in microbial community of
biotic control digesters were determined as 74.2%±3.4% and 21.8%±0.6%,
respectively. For spiked digesters, 73.5%±4.6% and 20.7%±1.2% abundances were
observed for Bacteria and Archaea domains. In BD and SD, amount of αProteobacteria varied between 2.4%±0.2% and 3.1%±0.5% during the 60 days of
operation period after the spike of NP2EO or acetone into the digesters according to
FISH results. Depending on these results it could be concluded that the addition of
NP2EO into spiked digesters did not lead to any noteworthy change in the
abundance of α-Proteobacteria when compared to control digesters. Similar trend
for α-Proteobacteria was observed with qPCR assays. Relative abundance for δProteobacteria ranged between 2.8%±0.6% and 3.2±0.4% for BD, on the other
hand, 3.1±0.5% and 3.4±0.7% for SD digesters at specific sampling times during
the reactor operation based on qPCR assays. It could be said that the distribution of
bacterial strains belonging to α-Proteobacteria and δ-Proteobacteria in the spiked
digesters was similar to that in the control digesters and addition of NP2EO did not
265
lead to any remarkable change in abundance of both sub-groups of Proteobacteria
in spiked digesters according FISH and qPCR results.
The composition of the bacterial community structure in both digesters exhibited a
high abundance of β-Proteobacteria and γ-Proteobacteria compared to other subgroups of Proteobacteria. After NP2EO spike, in 20 days, relative abundance of βProteobacteria reached much higher levels compared to control digesters. Its
abundance approached to 20% at the 71st day and then decreased to 18.6±1.4% and
17.4±1.6% at 76th and 83rd days, respectively. Both FISH and qPCR results
supported this finding. Relative abundance of γ-Proteobacteria in digesters was
close to 8% in both control and spiked digesters prior to spike. Then the abundance
of γ-Proteobacteria in sludge samples of spiked digesters rose to 12.4%±0.9% at
66th day according to FISH results. At 71st day γ-Proteobacteria abundance reached
to 16.6%±1.3%, on the other hand, the level of γ-Proteobacteria remained around
7.4% in control digesters according to FISH results. Based on qPCR results, pretty
much similar trend was observed for γ-Proteobacteria. The abundance reached to
18.2%±1.3% at 71th day according to qPCR results and then remained around 15%
in the following days with minor fluctuations. It can be said that the members of βProteobacteria and γ-Proteobacteria are better at adapting sudden changes in the
environment and may play a significant role in degradation of NP compounds.
Biotic control and spiked digesters were dominated by acetoclastic Methanosaeta
species all throughout the study. As average, Methanosaeta constituted 77% of
overall Archaea domain in control digesters, on the other hand, this ratio was
determined as 65.1% for spiked digesters according to FISH results. In a similar
way, qPCR results indicated that 75.7% and 62.5% of Archaea domain were
represented by Methanosaeta in control and spiked digesters, respectively.
Although Methanosaeta dominated in all digesters during operation time, the
increase in the abundance of Methanosarcina in spiked digesters was observed with
the addition of NP2EO. The abundance of Methanosarcina increased to 17.5%
compared to 11.5% in control digesters. It can be said that NP2EO addition and
accumulation of NP with time as a last degradation product led to a noticeable shift
266
in abundance of Methanosarcina without leading to any disturbance in digester
performance. This study demonstrated that molecular techniques are useful tools in
monitoring the effect of NP2EO on bacterial and archaeal composition of anaerobic
digesters. Fluorescence in situ hybridization and quantitative results both showed
similar results and that amendment of digesters with NP2EO promoted a
considerable shift in β- and γ-Proteobacteria populations of the anaerobic digesters.
Also a remarkable shift was observed in the abundance of Methanosarcina in
spiked digesters without deterioration of digester performance.
Bacterial strains able to grow in the presence of tNP and able to use it as the sole
carbon and energy source were isolated following enrichment studies from
activated sludge taken from WWTP of Ankara. Five NP-degrading isolates were
acquired,
which
were
distributed
among
5
genera:
Stenotrophomonas,
Burkholderia, Shinella, Ochrobactrum and Citrobacter. These strains belonging to
α-, β- and γ- Proteobacteria phyla which are widely reported as NPEO-degraders.
With isolation of these strains, phylogenetically diverse bacterial strains have been
isolated as aimed. These strains exhibited growth in the presence of two isomers of
NP, namely, p353NP and p363NP, based on OD values obtained during growth
studies.
267
268
CHAPTER 9
RECOMMENDATIONS
In the light of obtained results and conclusions of this study, the following ideas
could be suggested for future studies:

The developed and optimized methods for extraction and quantifiction of
NPEs can be tested and applied to monitor NPE concentrations in different
environmental systems like lake and river water and sediment samples. In
addition methods can be tested for sludge samples of different origin such
as purely domestic or industrial as well as varying regional properties.

In this study, molecular tools, FISH and qPCR were successfully applied for
monitoring of Proteobacteria sub-groups and acetoclastic methanogens
including Methanosarcina and Methanosaeta in anaerobic digesters. More
research could be performed for investigation and monitoring of other
bacterial and archaeal species forming microbial community structure of
anaerobic digesters. Also, the response of total microbial community can be
monitored in the presence of different NP compounds. Moreover, the effects
of NP compounds on methanogenesis and digester performance at high
concentrations should be investigated.

In the present study, it was demonstrated that bacterial strains belonging to
β- and γ-Proteobacteria may play important roles in the degradation of
NP2EO in anaerobic digesters. More research is needed to find out if these
bacterial species are able to degrade NP compounds in an anaerobic
environment.
269

Five different bacterial strains belonging to α- β- and γ-Proteobacteria
phyla were isolated during enrichment studies using tNP as a substrate.
Detailed research is needed to determine their metabolic degradation
pathways and behavior in the presence of other compounds structurally
similar to NP such as octylphenol (OP) and bisphenol A (BPA).
270
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APPENDIX A
Figure A.1. Calibration curve for 4-n-NP
Figure A.2. Calibration curve for NP
307
Figure A.3. Calibration curve for NP1EO
Figure A.4. Calibration curve for NP2EO
308
Figure A.5. Calibration curve for NP1EC
Figure A.6. Calibration curve for tNP
309
Figure A.7. Calibration curve for p353NP
Figure A.8. Calibration curve for p363NP
310
APPENDIX B
Figure B.1. A standard curve constructed for quantification of abundance of αProteobacteria in sludge samples
Figure B.2. A standard curve constructed for quantification of abundance of βProteobacteria in sludge samples
311
Figure B.3. A standard curve constructed for quantification of abundance of
Bacteria domain in sludge samples
Figure B.4. A standard curve constructed for quantification of abundance of γProteobacteria in sludge samples
312
Figure B.5. A standard curve constructed for quantification of abundance of
Archaea domain in sludge samples
Figure B.6. A standard curve constructed for quantification of abundance of δProteobacteria in sludge samples
313
Figure B.7. A standard curve constructed for quantification of abundance of
Methanosaeta in sludge samples
Figure B.8. A standard curve constructed for quantification of abundance of
Methanosarcina in sludge samples
314
APPENDIX C
Figure C.1. 1H and 13C NMR spectrum of p353NP
315
Figure C.2. 1H and 13C NMR spectrum of p363NP
316
CURRICULUM VITAE
PERSONAL INFORMATION
Name
Address
Telephone
Fax
E-mail
KARA MURDOCH, Fadime
Universiteler Mah., ODTU Lojmanları, 1610/4, 06800, Çankaya,
Ankara, TURKEY
Cell Phone: (+90) 535 954 41 07 Work: (+90) (312) 210 58 62
(+90) (312) 210 26 46
fadimekara@gmail.com/fkara@metu.edu.tr
Nationality
Date of birth/place
Marital Status
Turkish
07.08.1980/GERMANY
Married
WORK EXPERIENCE
Dates (from
– to)
Name and
address of
employer
Type of
business or
sector
Occupation
or position
held
TA
Experience
April 2004-current
Middle East Technical University, Department of Biotechnology
Middle East Technical University, Department of Environmental
Engineering
Academia
Teaching Assistant and Researcher
-
Environmental Microbiology Laboratory, 2005-2013 (ENVE 202)
Environmental Chemistry Laboratory, 2005-2014 (ENVE 208)
Introduction to Environmental Chemistry, 2012-2015 (ENVE 102)
Environmental Impact Assessment, 2011-2012, 2014-2015 (ENVE
420)
Environmental Management, 2011-2012, 2014-2015 (ENVE 406)
Summer Practice I-II, 2012-2014 (ENVE 300&400)
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Other
Professional
Activities
- Organization Committee Member
25th Year Anniversary Celebration Event of METU Biotechnology
Graduate Department, 14 November 2014, Ankara, Turkey
- Organization Committee Member
International Symposium on Biotechnology: Developments and Trends, 2730 September 2009, Middle East Technical University, Ankara, Turkey
- Organization Committee Member
International Water Association (IWA) sponsored Specialist Conference of
"Facing Sludge Diversities: Challenges, Risks and Opportunities" 28-30
March 2007, Antalya, Turkey.
EDUCATION AND TRAINING
Dates (from –to)
Name and type of organisation
providing education and
training
Thesis Title
Title of qualification to
beawarded
2007-2015
Middle East Technical University, Department of
Biotechnology
Supervisor: Prof.Dr.F. Dilek Sanin
(Department of Environmental Engineering, METU),
Co-Supervisor: Prof.Dr. G. Candan Gürrakan
(Department of Food Engineering, METU)
Biodegradation of Nonylphenols, Determination of
Degradation Products and Detection of Responsible
Microorganisms Using Molecular Techniques
PhD in Biotechnology
Dates (from – to)
Name and type of organization
providing education and
training
Occupation or position held
August 2010-August 2011
Cornell University, Department of Microbiology
Supervisor: Assoc.Prof.Dr. Anthony Hay
Dates (from – to)
Name and type of organisation
providing education and
training
September 2004-June 2007
Middle East Technical University, Department of
Biotechnology
Supervisor: Prof.Dr.F. Dilek Sanin
(Department of Environmental Engineering, METU),
Co-Supervisor: Prof.Dr. G. Candan Gürrakan
(Department of Food Engineering, METU)
Investigation of Sodium and Potassium Ions In Relation
to Bioflocculation of Mixed Culture Microorganisms
MSc in Biotechnology
Thesis Title
Title of qualification awarded
Dates (from – to)
Name and type of organisation
providing education and
training
Researcher
September 1999-June 2003
Ankara University, Department of Biology
318
Principal subjects/occupational
skills covered
Title of qualification awarded
Biology and its applications
BSc in Biology
I. PUBLICATIONS LIST
INTERNATIONAL
1. Ömeroğlu S., Kara Murdoch F. and Sanin F.D. (2015). “Investigation of nonylphenol
compounds in sewage sludge samples from a metropolitan wastewater treatment plant in
Turkey”.
Talanta, 131, 650-655. doi:10.1016/j.talanta.2014.08.014
2. Kara F., Gürakan G.C. and Sanin F.D. (2008). “Monovalent cations and their influence
on activated sludge floc chemistry, structure and physical characteristics”.
Biotechnology&Bioengineering, 100:2, 231-239. doi: 10.1002/bit.21755
3. Sanin F.D., Vatansever A., Turtin İ. Kara, F., Durmaz B. and Sesay M.L. (2006).
“Operational conditions of activated sludge: influence on flocculation and dewaterability”.
Drying Technology, 24:10, 1297-1306. doi: 10.1080/07373930600838009
4. Kara Murdoch F., Ömeroğlu S. and Sanin F.D. “Development of extraction and
analysis methods to determine nonylphenol compounds in wastewater effluents” In
preparation
5. Kara Murdoch F., and Sanin F.D. “Nonylphenol compounds: Anaerobic treatment and
impact on digester performance”. Submitted
6. Kara Murdoch F., Gürakan G.C. and Sanin F.D. “Changes in microbial community
structure during the biodegradation of nonylphenol diethoxylate in anaerobic semicontinuous digesters”. In preparation
7. Kara Murdoch F., Anthony Hay and Sanin F.D. “Isolation and characterization of
novel bacterial strains capable of aerobic growth on branched nonylphenol”. In preparation
NATIONAL
1. Kara F. and Sanin F.D. “Çamur örneklerinde patojen ve indikatörler, yeni bulgular ve
çamurun tarımsal kullanımında önemi” (Pathogens and indicators in sludge samples, new
findings and their importance for sludge use in agriculture). Katı Atık ve Çevre, Sayı: 78,
2010.
2. Ömeroğlu S., Kara F., Ahmad M., Bozkurt H. and Sanin F.D. “Nonil fenollerin çevre
sistemlerinde ve arıtma çamurunda varlığı ve etkileri” (Presence and impact of
319
nonylphenols in environmental and activated sludge systems) Katı Atık ve Çevre, Sayı: 83,
2011.
II. CONFERENCE PROCEEDING PUBLICATIONS
INTERNATIONAL
1. Kara Murdoch F. and Sanin F.D. “Assessment of anaerobic microbial community
structure in the presence of nonylphenol diethoxylate using fluorescence in situ
hybridization”. ICOCEE, Cappadocia -2015, May 20-23, 2015, Nevsehir, Turkey (oral
presentation)
2. Kara Murdoch F. and Sanin F.D. “Seasonal monitoring of nonylphenol compounds in
sewage sludge”. 14th International Conference on Environmental Science and Technology,
CEST2015, 3-5 September 2015, Rhodes, Greece (oral presentation)
3. Kara Murdoch F., Murdoch R. and Sanin F.D. “Determination of relative abundance of
Sphingomonads in aerobic batch and semi-continuous digesters”. IWA World Water
Congress&Exhibition, 21-26 September 2014, Lisbon, Portugal (poster presentation)
4. Ömeroğlu S., Kara Murdoch F. and Sanin F.D. “Development of practical extraction
and analysis methods for nonylphenol compounds in water and wastewater”. IWA World
Water Congress&Exhibition, 21-26 September 2014, Lisbon, Portugal (poster
presentation)
5. Kara Murdoch F., Ahmad M. and Sanin F.D. “Degradation of nonylphenol compounds
in aerobic and anaerobic digesters”. ECSM 2014 – 4th European Conference on Sludge
Management, 26&27 May 2014, Izmir, Turkey (oral presentation)
6. Kara Murdoch F., Ömeroğlu S., Bozkurt H., Ahmad M., and Sanin F.D. "The fate of
nonylphenolic compounds in water and wastewater systems". 17th International
Symposium on Environmental Pollution and its Impact on Life in the Mediterranean
Region. 28 September-1 October, 2013, Istanbul, Turkey (poster presentation)
7. Bozkurt H., Ömeroğlu S., Ahmad M. Kara F. and Sanin, F. D. “Nonylphenols in sludge:
measurement and removal methods” V. Turkish German Solid Waste Days, TAKAG 2011,
27-30 September 2011, Stuttgart, Germany (oral presentation)
8. Kara F., Gurakan C. and Sanin F.D. “Monovalent cations and their influence on
activated sludge floc structure”. IWA, Moving Forward, Wastewater Biosolids
Sustainability: Technical, Managerial, and Public Synergy, 24 - 27 June 2007, Moncton,
Canada. Proceeding Book, 165-172 (oral presentation)
9. Vatansever A., Kara F., Turtin I. and Sanin F.D. “Effect of cation type and
concentration on bioflocculation and settling properties of activated sludge”. IWA, Facing
Sludge Diversities: Challenges, Risks and Opportunities, 28-30 March 2007, Antalya,
Turkey. Proceeding Book, 85-92 (oral presentation)
10. Doğan I., Köksoy G.T., Kara F., Kıvılcımdan Ç. and Sanin F.D. “Lab scale evaluation
of thermal, thermo-chemical and acidic pretreatment of sludge to improve anaerobic
320
digestion”. IWA, Facing Sludge Diversities: Challenges, Risks and Opportunities, 28-30
March 2007, Antalya, Turkey. Proceeding Book, 149-156 (oral presentation)
NATIONAL
1. Kara Murdoch F. and Sanin F.D. “Nonilfenol bileşiklerinin anaerobik yarı sürekli
sistemlerde akıbetinin incelenmesi” (Fate of nonylphenol compounds in semi-continuous
anaerobic systems). UKAY 2013, 5. Ulusal Katı Atık Yönetimi Kongresi, 29 May-1 June
2013, Kocaeli, Turkey (in Turkish, oral presentation)
2. Ömeroğlu S., Kara F. and Sanin F.D. “Nonilfenol bileşiklerinin arıtma çamurlarında
ölçümüne yönelik metotların geliştirilmesi ve uygulanması” (Development and application
of methods for quantification of nonylphenol compounds in activated sludge). İTU 13.
Endüstriyel Kirlenme Kontrolü Sempozyumu, 17-19 October 2012, İstanbul, Turkey (in
Turkish, oral presentation)
3. Ömeroğlu S., Kara F., Ahmad M., Bozkurt H. and Sanin F.D. “Nonil fenollerin çevre
sistemlerinde ve arıtma çamurunda varlığı ve etkileri” (Presence and effect of nonylphenol
compounds on environmental systems and activated sludge). 2. Ulusal Katı Atık Yönetimi
Kongresi, UKAY 2010, 18-20 November, 2010, Mersin, Turkey (in Turkish, oral
presentation)
4. Kara, F. and Sanin F.D. “Çamur örneklerinde patojen ve indikatörler, yeni bulgular ve
çamurun tarımsal kullanımında önemi” (Pathogens and indicators in sludge, new findings
and their importance for sludge use in agriculture” 2nd National Syposium of Treatment
Plant Sludges, 4-6 November 2009, Izmir, Turkey, Proceeding Book, 51-61 (in Turkish,
oral presentation),
5. Kara F., Vatansever A., Turtin I. and Sanin F.D. ““Katyon tipi ve konsantrasyonun aktif
çamur yumaklasma ve susuzlastırılabilme ozelliklerine olan etkilerinin incelenmesi”
(Examination of the effect of cation type and concentration on bioflocculation and
dewatering properties of activated sludge). Türkiye’de Çevre Kirlenmesi Öncelikleri
Sempozyumu V, 14-18 May 2008, Kocaeli, Turkey (in Turkish, oral presentation)
III. AWARDS AND FELLOWSHIPS
2007 -2013
Doctoral Fellowship from Turkish Scientific and Technical Research
Council (TUBITAK)
2010-2011
Doctoral Fellowship for Research Abroad from Turkish Scientific and
Technical Research Council (TUBITAK)
2009
Course Performance Award in PhD, Middle East Technical University,
Biotechnology Department
2003
Biology
1st Highest Degree in BSc Graduation, Ankara University, Department of
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