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Studies on the behaviour of endocrine disrupting compounds in a membrane bioreactor
Von der Fakultät für Mathematik, Informatik und Naturwissenschaften der Rheinisch-Westfälischen Technischen Hochschule Aachen zur Erlangung des akademischen Grades
einer Doktorin der Naturwissenschaften genehmigte Dissertation
vorgelegt von
Diplom-Ingenieurin
Magdalena Cirja
aus Borca-Neamt, Rumänien
Berichter: Univ.-Prof. Dr. rer. nat. Andreas Schaeffer Prof. Dr. rer. nat. Philippe Corvini
Tag der mündlichen Prüfung: 3. Dezember 2007
Diese Dissertation ist auf den Internetseiten der Hochschulbibliothek online verfügbar
Parts of this thesis have been published in scientific journals or are submitted for publication: Cirja M, Hommes G, Ivashechkin P, Prell J, Schäffer A, Corvini PFX (2008)
Bioaugmentation of Membrane Bioreactor with Sphingomonas sp. strain TTNP3 for the
Degradation of Nonylphenol. (Submitted)
Cirja M, Ivashechkin P, Schäffer A, Corvini PFX (2008) Factors affecting the removal of
organic micropollutants from wastewater in conventional treatment plants (CTP) and
membrane bioreactors (MBR) (Review); Reviews in Environmental Science and
Biotechnology 7 (1) : 61-78
Cirja M, Zühlke S, Ivashechkin P, Hollender J, Schäffer A, Corvini PFX (2007) Behaviour
of two differently radiolabelled 17α-ethinylestradiols continuously applied to a lab-scale
membrane bioreactor with adapted industrial activated sludge. Water Research 41: 4403-4412
Cirja M, Zühlke S, Ivashechkin P, Schäffer A, Corvini PFX (2006) Fate of a 14C-Labeled
Nonylphenol Isomer in a Laboratory Scale Membrane Bioreactor; Environmental Science and
Technology 40 (19): 6131-6136
Summary Environmental pollution with persistent chemicals becomes an increasingly important issue.
Nowadays a variety of chemicals as pesticides, dyes, detergents are introduced in a very large
scale on the surface water network. The main pathway of micropollutants into the
environment was identified as municipal wastewater. The extended use of chemicals in many
product formulations and insufficient wastewater treatment lead to an increase of the detected
micropollutant quantities in wastewater effluents. The majority of these compounds are
characterized by a rather poor biodegradability. A large spectrum of pollutants present in
waste as traces has been reported to exert adverse effects for human and wildlife. Even though
compounds are found in wastewater in a very small amount, they may have the undesirable
capability of having estrogenic activity on various high forms of life.
The present research focuses on the fate of hydrophobic micropollutants in a
membrane bioreactor (MBR) and their removal in it. MBR represents one of the most
promising innovations in the field of wastewater treatment because of the high efficiency in
removal of organics and nutrients. The high quality of the effluent is obtained by the complete
retention of suspended solids, the almost complete removal of pathogens, and the possibility
to increase biodegradation of micropollutants because of the higher sludge retention time in
MBR, in comparison to the conventional activated sludge treatment.
For the micropollutants which have hydrophobic properties the challenge is to
distinguish the real biodegradation from abiotic physico-chemical phenomena like adsorption
and volatilisation during the membrane bioreactor treatment. In order to achieve it, the
application of radiolabelled hydrophobic model markers is considered to be a powerful
technique.
In this study the possible fate of NP during wastewater treatment by using a lab-scale
MBR was investigated. After a single pulse of the 14C-labelled-NP isomer (4-[1-ethyl-1,3-
dimethylpentyl]phenol) as radiotracer, the applied radioactivity was monitored in the MBR
system over 34 days. The balance of radioactivity at the end of the study showed that 42% of
the applied radioactivity was recovered in the effluent as degradation products of NP, 21%
was removed with the daily excess sludge from the MBR, and 34% was recovered as
adsorbed in the component parts of the MBR. A high amount of NP was associated to the
sludge during the test period, while degradation products were mainly found in the effluent.
Partial identification of these metabolites by means of HPLC-tandem mass spectrometry
coupled to radio-detection showed that they were alkyl-chain oxidation products of NP. The
use of a radiolabelled test compound in a lab-scale MBR was found suitable to demonstrate
that the elimination of NP through mineralization and volatilization processes under the
applied conditions was negligible (both less than 1%). However, the removal of NP via
sorption and the continuous release of oxidation products of NP in the permeate were highly
relevant.
The fate of two differently labelled radioactive forms of 17α-ethinylestradiol (EE2),
one with the ring and the other one with the ethinyl group labelled was studied during the
membrane bioreactor (MBR) process. The system was operated using a synthetic wastewater
representative for the pharmaceutical industry and the activated sludge was obtained from a
large-scale MBR treating pharmaceutical wastewater. Before the five days radioactive
experiment, the activated sludge was adapted for 29 days continuously to non-labelled EE2.
Balancing of radioactivity could demonstrate that mineralization amounted to less than 1% of
the applied radioactivity. The EE2 residues remained mainly sorbed in the reactor, resulting in
a removal of approximately 80% relative to the concentration in the influent. The same
metabolite pattern in the radiochromatograms of the two differently labelled 14C-EE2
molecules led to the assumption that the elimination pathway does not involve the removal of
the ethinyl group from the EE2 molecule.
The bioaugmentation of membrane bioreactor in order to improve the degradation of
recalcitrant nonylphenol during the wastewater treatment was studied. The 14C-labelled NP
isomer 4-[1-ethyl-1,3-dimethylpentyl]phenol was applied as single pulse to the membrane
bioreactor bioaugmented with the bacterium Sphingomonas sp. strain TTNP3. The effects of
five successive inoculations of the membrane bioreactor with this strain able to degrade NP
were investigated in comparison to a non-bioaugmented reactor. Results showed that the
radioactivity spiked in the bioaugmented system was retrieved mostly in the effluent (56%),
followed by fractions sorbed to the system (25%), associated with the excess sludge (9%) and
collected from the gas phase as CO2 resulting from mineralization (2.3%). The degradation
products identified in the treated effluent and in the MLSS were specific metabolites of
catabolism of the NP by Sphingomonas, e.g. hydroquinone resulting from ipso-substitution.
The capacity of this bacterium to excrete biosurfactants and to increase nonylphenol
bioavailability was investigated. The presence and persistence of the strain in the membrane
bioreactor was examined by performing polymerase chain reaction–denaturing gradient gel
electrophoresis (PCR-DGGE).
Zusammenfassung Die Belastung der Umwelt mit persistenten Chemikalien ist eine Thematik von zunehmender
Bedeutung. Heutzutage gelangen verschiedenste Chemikalien, wie sie z.B. in
Pflanzenschutzmittel, Farben und Detergenzien enthalten sind, in bedeutenden Mengen in die
Oberflächengewässer. Der Haupteintagspfad für Mikroschadstoffe in die aquatische Umwelt
ist das kommunale Abwasser. Die immer umfangreichere Verwendung von Chemikalien in
vielen Produktformulierungen und deren unvollständige Eliminierung während der
Abwasserbehandlung führen zu einem Anstieg der Mikroschadstoffkonzentrationen in
Kläranlagenabflüssen. Die Mehrheit dieser Verbindungen zeichnet sich durch eine schlechte
biologische Abbaubarkeit aus. Für eine Vielzahl von Schadstoffen, die in Spuren in Abfällen
enthalten sind, wird eine mögliche nachteilige Beeinflussung des hormonalen Systems von
Mensch und Tier diskutiert. Auch wenn diese Verbindungen im Abwasser nur in sehr
niedrigen Konzentrationen vorliegen, können sie die unerwünschte Fähigkeit haben, viele
höhere Lebensformen durch ihre endokrine Aktivität zu beeinflussen.
Die vorliegende Arbeit beschäftigt sich mit dem Schicksal und der gezielten
Eliminierung hydrophober Mikroschadstoffe in einem Membranbioreaktor (MBR). Die
MBR-Technologie ist aufgrund ihres hohen Eliminierungspotentials für die organische Fracht
sowie Nährstoffe eine der vielversprechendsten Innovationen auf dem Gebiet der
Abwasserreinigung. Dabei wird die hohe Effluentqualität erreicht durch einen vollständigen
Rückhalt von suspendierten Feststoffen, die nahezu vollständige Entfernung von Pathogenen
und die Möglichkeit einer gesteigerten biologischen Eliminierung aufgrund des höheren
Schlammalters im MBR.
Bei hydrophoben Mikroschadstoffen ist es für die Untersuchung der Vorgänge
während der Behandlung im MBR wichtig, zwischen biologischem Abbau und abiotischen
Eliminationsprozessen wie Adsorption und Verflüchtigung zu unterscheiden. Um dies zu
erreichen, ist der Einsatz von radioaktiv-markierten Testchemikalien ein hilfreiches Mittel.
In der vorliegenden Studie wurde das mögliche Schicksal von Nonylphenol während
der Abwasserreinigung in einem MBR im Labormaßstab untersucht. Nach einer
Pulsdosierung des 14C-markierten Nonylphenol-Isomers 4-[1-ethyl-1,3-dimethylpentyl]phenol
wurde die Verteilung der Radioaktivität im System über 34 Tage verfolgt. Die 14C-Bilanz am
Ende der Studie zeigte, dass 42% der applizierten Radioaktivität in Form von
Degradationsprodukten des Nonylphenols mit dem Effluenten und weitere 21% mit der
täglichen Entnahme des Überschussschlamms aus dem MBR ausgetragen wurden. Am
Versuchsende waren 34% der bei Versuchsstart applizierten Radioaktivität an den Bauteilen
des MBR adsorbiert. Ein hoher Anteil von Nonylphenol lag während der Studie an den
Schlamm assoziiert vor, wohingegen Radioaktivität in Form von Metaboliten überwiegend
frei im Effluenten zu finden war. Zum Teil konnten die gefundenen Metaboliten mit Hilfe von
HPLC-Tandem-Massenspektrometrie in Kombination mit Radiodetektion als Alkylketten-
Oxidationsprodukte von Nonylphenol identifiziert werden. Durch den Einsatz von radioaktiv-
markierter Testsubstanz konnte weiterhin gezeigt werden, dass Mineralisations- und
Verflüchtigungsprozesse unter den gewählten Versuchsbedingungen für das Schicksal von
Nonylphenol in einem MBR nur eine unbedeutende Rolle spielen (jeweils unter 1%). Im
Gegensatz dazu waren der Rückhalt im MBR über Sorption sowie die oxidative
Metabolisierung die entscheidenden Prozesse für die Elimination von Nonylphenol aus dem
Abwasserstrom.
Weiterhin wurde im MBR-System das Schicksal von 2 unterschiedlich 14C-markierten
17α-Ethinylestradiol-Molekülen untersucht, einem mit Ringmarkierung und einem mit
radioaktiv-markierter α-Ethinylgruppe. Der MBR wurde hierzu mit einem speziellen
synthetischen Abwasser betrieben, das in seiner Zusammensetzung Abwasser der
Pharmaindustrie repräsentierte, sowie mit Belebtschlamm aus einer MBR-Anlage eines
pharmazeutischen Produzenten. Vor der 5-tägigen Studie mit radioaktiver Testsubstanz wurde
der Belebtschlamm im MBR für 29 Tage kontinuierlich mit nichtmarkiertem 17α-
Ethinylestradiol vorkonditioniert. Die 14C-Bilanz zeigte, dass weniger als 1% des applizierten
Ethinylestradiols bis zum Ende der Studie mineralisiert worden war. Etwa 80% der mit dem
Influenten zugeführten Radioaktivität wurde über Sorptionsprozesse innerhalb des Reaktors
aus dem Abwasserstrom eliminiert. Es konnte gezeigt werden, dass die Ethinylgruppe bei den
beobachteten Metabolisierungen der Substanz nicht betroffen war, da sich die
Metabolitenspektren in den Radiochromatogrammen der beiden unterschiedlich markierten
17α-Ethinylestradiol-Moleküle nicht unterschieden.
Weiterhin wurde mit dem MBR eine Bioaugmentationsstudie durchgeführt, um die
Eliminationsleistung für Nonylphenol während der Abwasserbehandlung zu steigern. Das 14C-markierte Nonylphenol-Isomer 4-[1-ethyl-1,3-dimethylpentyl]phenol wurde dem MBR in
einer einfachen Pulsdosierung zugegeben, dessen Belebtschlamm wiederholt mit dem
Bakterium Sphingomonas sp. Stamm TTNP3 inokuliert wurde. Die Effekte von 5
aufeinanderfolgenden Inokulationen des MBR mit diesem Nonylphenol abbauenden
Bakterienstamm wurden durch den Vergleich mit einem MBR-Ansatz ohne Bioaugmentation
untersucht. Die Ergebnisse zeigten, dass der Hauptanteil der Radioaktivität in der
Bioaugmentationsstudie den MBR mit dem Effluenten verließ (56%), gefolgt von der
innerhalb des MBR-Systems sorbierten Fraktion (25%), der mit dem Überschussschlamm
entfernten Radioaktivität (9%) und der aus Mineralisationsprozessen resultierenden
Radioaktivität in der CO2-Falle (2,3%). Die im Effluenten und im Belebtschlamm
identifizierten Degradationsprodukte bestanden aus typischen, in der Literatur beschriebenen
Metaboliten des Nonylphenolabbaus von Sphingomonas, z.B. das aus der ipso-Substitution
resultierende Hydrochinon. Die Fähigkeit dieses Bakteriums, biogene Detergenzien in das
Medium abzugeben und damit die Bioverfügbarkeit von Nonylphenol zu erhöhen wurde
untersucht. Die Präsenz von Sphingomonas im Medium und damit dessen
Überlebensfähigkeit im MBR in Koexistenz mit der konventionellen
Belebtschlammbiozoenose konnte mit der PCR-DGGE-Technik (Polymerasekettenreaktion-
Denaturierende Gradientengelelektrophorese) bewiesen werden.
I
List of content
1. General introduction ……………………….………………………………………… 1
2. State-of-the-art ………………………….……………………………………………. 3
2.1. Micropollutants in wastewater ……………………………………………… 3
2.2. Micropollutants as endocrine disrupting compounds ……….……………… 5
2.3. Conventional treatment plant and membrane bioreactor …………………… 5
2.4. Factors affecting the removal of micropollutants .……………….…………. 8
2.4.1. Hydrophobicity and hydrophilicity ……………………………... 8
2.4.2. Bioavailability …………………………………………………... 9
2.4.3. Sorption ……………………...………………………………… 10
2.4.4. Biodegradation ………………………………………………… 11
2.4.5. Abiotic degradation and volatilization ………………………… 12
2.4.6. Compound structure …………………………………………… 12
2.4.7. Sludge retention time and biomass concentration …………….. 13
2.4.8. Influence of biomass characteristics on sorption and
biodegradation ……………………………………………….. 15
2.4.9. pH value ……………………………………………………….. 16
2.4.10. Temperature ………………………………………………….. 17
2.5. Fate of model micropollutants in WWTP …………………………………. 19
2.5.1. Fate of NP in WWTP ………………………………...………. 19
2.5.2. Fate of EE2 in WWTP ………………………………..……… 26
3. Objectives and outline of this thesis ………………………………………………… 31
4. Materials and Methods ………………………………………………………………. 33
4.1. List of used chemicals and instruments …………………………………… 33
4.1.1. Chemicals and materials ………………………………………. 33
4.1.2. Instruments and devices ………………...……………………... 34
4.1.3. Other materials ………………………………………………… 35
4.2. Lab-scale MBR design …………………………………………………….. 37
4.2.1. Set-up of the pumps ...………………………………………….. 37
4.2.2. Operational parameters ………………………………………... 39
4.3. Analysis of water quality parameters ……………………………………… 41
II List of content
4.4. Liquid scintillation counting ..…….……………………………………….. 41
4.5. Determination of non-extractable residues ………………………………... 41
4.6. Radiolabelled compounds .………………………………………………… 42
4.6.1. 14C-nonylphenol preparation ……………….…………………. 42
4.6.2. 14C-17α-ethinylestradiol .………….……...…………………… 42
4.7. HPLC with radio-detection ……………………………………...………… 43
4.7.1. Analyses of NP and its degradation products …………………. 43
4.7.2. Analyses of EE2 and its degradation products ………………... 43
4.8. HPLC-MS/MS analyses …………………………………………………… 43
4.9. GC-MS analyses …………………………………………………………… 44
4.10. Preparation of Sphingomonas sp. strain TTNP3 cultures ……………..…. 44
4.10.1. Sphingomonas sp. strain TTNP3 ………………………...…… 44
4.10.2. Glycerol stock ………………………………………………... 45
4.10.3. Resting cells cultures ……………………………………….... 45
4.11. PCR-DGGE ………………………………………………………………. 46
4.11.1. 16S rDNA amplification and sequencing ……………………. 46
4.11.2. PCR-DGGE analyses ………………………………………… 46
4.12. Surface tension measurements …………………………………………… 47
4.13. Sampling and preparation of the samples ………………………………... 47
4.14. Preliminary tests .…………………………………………………………. 48
4.14.1. NP and EE2 studies …………………………………………... 48
4.14.2. Bioaugmentation study ………………………………………. 48
4.15. Experimental procedure ………………………………………………….. 49
4.15.1. NP fate study …………………………………………………. 49
4.15.2. EE2 fate study ………………………………………………... 49
4.15.3. Bioaugmentation procedure ………………………………….. 50
5. Results .……………………………………………………………………….……….. 53
5.1. Performances of MBR: optimization and operational parameters ……..….. 53
5.1.1. General prerequisites …………………………………..……… 53
5.1.2. Synthetic medium formulation ………………………………... 53
5.2. Fate of 14C-NP in lab-scale MBR ………………………………………….. 56
5.2.1. Radioactivity monitoring ……………………………………… 56
5.2.2. Sludge analyses ………………………………………………... 58
List of content III
5.2.3. Effluent analyses ………………………………………………. 60
5.2.4. Radioactivity balance ………………………………………..… 64
5.3. Behaviour of two differently 14C-EE2 continuously applied to MBR …...... 65
5.3.1. Radioactivity monitoring ……………………….……………... 65
5.3.2. Sludge analyses ………………………………………………... 68
5.3.3. Effluent analyses ………………………………………………. 69
5.3.4. Radioactivity balance ………………………………………….. 72
5.4. Bioaugmentation of MBR with Sphingomonas …………………................ 73
5.4.1. Pre-studies on the degradation of nonylphenol by Sphingomonas
sp. Strain TTNP3 …………………………………………….. 73
5.4.2. Radioactivity monitoring ……………………………………… 74
5.4.3. Sludge analyses ……………………………………………...… 76
5.4.4. Detectability of Sphingomonas in the bioaugmented MBR .….. 77
5.4.5. NP bioavailability in presence of Sphingomonas …………..…. 79
5.4.6. Radioactivity balance …………………..……………………… 81
6. Discussion ……………………………………………………………….……....…….. 83
6.1. Performances of MBR: optimization and operational parameters ………… 83
6.2. Fate of 14C- NP in lab-scale MBR ……………………………………….… 84
6.2.1. General overview of the results ……………………………….. 84
6.2.2. Radioactivity monitoring ……………………………………… 85
6.2.3. Sludge analyses ………………………………………………... 86
6.2.4. Effluent analyses ………………………………………………. 87
6.2.5. Radioactivity balance and benefits of the study …………….… 88
6.3. Behavior of two differently labelled 14C-EE2 applied to MBR …………… 89
6.3.1. General overview of the results ……………………………….. 89
6.3.2. Radioactivity monitoring ……………………………….……... 89
6.3.3. Sludge analyses ………………………………………………... 90
6.3.4. Effluent analyses ………………………………………………. 90
6.3.5. Radioactivity balance and benefits of the study ………….…… 91
6.4. Bioaugmentation of MBR with Sphingomonas ……………………..…….. 93
IV List of content
6.4.1. General overview of the results …………………….………….. 93
6.4.2. Pre-studies on the degradation of NP in presence of
Sphingomonas ……………………………….……………….. 94
6.4.3. Radioactivity monitoring ……………………………………… 95
6.4.4. Sludge analyses ………………………………………………... 95
6.4.5. Detectability of Sphingomonas in bioaugmented MBR …….… 95
6.4.6. Bioavailability of NP in presence of Sphingomonas ………….. 96
6.4.7. Radioactivity balance and benefits of the study ………….…… 97
7. Conclusions and outlook ……………………………………………….……………. 99
8. Literature …………………………………………………………………………..... 103
9. Appendix …………………………………………………………………………….. 125
Acknowledgments ……………………………………………………………..………. 127
Curriculum vitae ………………………..……………………….…..………………… 129
V
Glossary
BOD Biological oxygen demand
BPA Bisphenol A
Bq Bequerel
COD Chemical oxygen demand
Cdissolved Concentration of the dissolved substance (g/L)
Cmicropollutant Concentration of micropollutant (mg/L)
Csorbed Concentration of sorbed substance (g/L)
CTP Conventional treatment plant
DAD Diode array detector
DDT Dichloro-diphenyl-trichloroethane
DGGE Denaturing gradient gel electrophoresis
DOM Dissolved organic matter
DNA Deoxyribonucleic acid
EDC Endocrine disrupting chemical
E1 Estrone
E2 17β-estradiol
EE2 17α-ethinylestradiol
EU European Union
GC Gas chromatography
H Henry constant
HRT Hydraulic retention time
HPLC High performance liquid chromatography
Kd Sorption constant (L/g)
Kdegrad Degradation constant (L/g)
Kow Partition coefficient octanol-water
Kp Partition coeffient
LAS Linear alkylbenzene sulphonates
LSC Liquid scintillation counter
M Molar concentration (mol/L)
MBR Membrane bioreactor
MLSS Mixed liquor solid suspension
MS Mass spectrometry
VI Glossary
m/z Mass per charge
NH4+ Ammonia
NO3- Nitrate
NP Nonylphenol
NPnEO Nonylphenol polyethoxylates
OP Octylphenol
PAH Polycyclic aromatic hydrocarbons
PBDE Polybrominated diphenyl ethers
PCB Polychlorinated biphenyl
PCR Polymerase chain reaction
pNP para nonylphenol
pKa Acid dissociation constant
Rdegradation Degradation rate
RT Retention time
SRT Sludge retention time
SS Concentration of solid suspension (g/L)
TOC Total organic carbon
TN Total nitrogen
tNP Technical Nonylphenol
TP Total phosphorus
TSS Total solid suspension
VOC Volatile organic compounds
WWTP Wastewater treatment plant
1
1. General introduction
Environmental pollution with organic micropollutants is nowadays of great concern,
especially for the aquatic environment. For many years, the quantification of water pollution
was restricted to the monitoring of biochemical oxygen demand, chemical oxygen demand,
nitrates, phosphates and total suspended solids (EEC waters guideline RL-76/464/EWG,
1976). Due to the requirements of the EU water framework guideline to the condition of
European surface waters (RL-2000/60/EG, 2000) the attention is increasingly shifted to the
micropollutants and the necessity of developing wastewater treatment systems able to
enhance their removal. The European Commission proposal setting environmental quality
standards for surface waters in terms of micropollutants (COM/2006/397) includes a list of 41
dangerous chemical substances (33 priority substances and 8 other pollutants). The substances
or groups of substances in the list of priority substances include metals and diverse organic
chemicals such as plant protection products, biocides, and other groups of organics like
polyaromatic hydrocarbons (PAH) that are mainly incineration by-products and
polybrominated diphenyl ethers (PBDE) that are used as flame retardants. In the last decades,
classes of compounds like pharmaceuticals, hormones, and personal care products have been
taken in consideration. Many of these human-made compounds can exert adverse effects on
human and wildlife organisms and are referred to as xenobiotics. Although these biologically
active compounds are usually detected in trace amounts, they are unfortunately often
characterized by rather poor biodegradability and persistence in the environment.
Membrane Bioreactor (MBR) represents one of the most promising innovations in the
field of wastewater treatment because of its high efficiency concerning the removal of
organics and nutrients. The high quality of the effluent is obtained through the complete
retention of suspended solids, the almost complete removal of pathogens, and the possibility
to increase the biodegradation of micropollutants because of the higher sludge retention time
in MBR, in comparison to the conventional activated sludge treatment. In general, it is
difficult to assess the fate of micropollutants which have hydrophobic properties in the
environment. In the present study, radiolabelled forms of hydrophobic model xenobiotics
were used in order to circumvent analytical problems related to the complex sludge matrix.
The challenge of this thesis consisted in distinguishing biodegradation processes from abiotic
physico-chemical phenomena like adsorption and volatilisation during the wastewater
treatment in MBR.The comprehension of the behaviour of xenobiotics in MBR is one
prerequisite for the possible improvement of the quality of effluent treated by such system.
3
2. State-of-the-art
2.1. Micropollutants in wastewater
Increasing worldwide contamination of freshwater systems with thousands of industrial and
natural chemical compounds is one of the key environmental problems facing humanity.
Nowadays, the focus in wastewater treatment switched from macropollutants such as acids,
salts, nutrients, and natural organic matter occurring at concentrations of µg/L to mg/L to
micropollutants. Although these compounds are present at low concentrations (ng/L to µg/L),
many of them raise considerable (eco)toxicological concerns, particularly when they are
present as components of complex mixtures. These compounds have been reported to be
ubiquitous in natural waters in the past 25 years, not only in industrial areas, but also in more
remote environments.
Micropollutants comprise organic and inorganic substances belonging to industrial
chemicals, pesticides, pharmaceuticals, natural and synthetic hormones, personal care
products, and also natural chemicals. Some chemicals are not degraded (e.g. heavy metals) or
only very slowly (e.g. persistent organic micropollutants such as dichloro-diphenyl-
trichloroethane (DDT) or polychlorinated biphenyls (PCBs)). Additionally, other compounds
that are less persistent and not prone to long-range transport are also of concern if they are
emitted continuously or give raise to harmful transformation products (e.g. hormones,
surfactants as nonylphenol polyethoxylates).
The most current environmental problems related to the pollution by such compounds
are the contamination of drinking water, the bioaccumulation of xenobiotic residues in food
(vegetables, milk, fishes, etc.), the threatening of biodiversity via adverse effects on endocrine
systems of living organisms, and the spreading of resistances towards these compounds
within microbial population (Table 2.1).
4 State-of-the-art
Table 2.1. Examples of ubiquitous water pollutants (adapted from Schwarzenbach et al., 2006).
Origin Class Examples Related problems Industrial chemicals Solvents Tetrachloromethane Drinking-water
contamination Intermediates Methyl-t-butylether Industrial products Additives Phthalates Lubricants PCBs (polychlorinated
biphenyls) Biomagnification, long-range transport
Flame retardants Polybrominated diphenylethers
Consumer products Detergents Nonylphenol ethoxylates
Endocrine active transformation product (nonylphenol)
Pharmaceuticals Antibiotics Bacterial resistance, nontarget effects
Hormones Ethinylestradiol Feminization of fish Personal–care
products Ultraviolet filters Multitude of effects
Biocides Pesticides DDT Toxic effects and persistent metabolites
Atrazine Effects on primary producers
Nonagricultural biocides
Tributyl tin Endocrine effects
Triclosan Persistent degradation products
Geogenic/natural chemicals
Heavy metals Lead, cadmium, mercury
Inorganics Arsenic, selenium, fluoride, uranium
Risk for human health
Taste and odour chemicals
Geosmin, methylisoborneol
Drinking water quality problems
Cyanotoxines microcystins Human
hormones Estradiol Feminization of fish
Disinfection/oxidation Disinfection by-products
Trihalomethans, haloacetic acids, bromate
Drinking water quality, human health problems
Transformation products
Metabolites from all above
Metabolites of perfluorinated compounds
Bioaccumulation despite low hydrophobicity
Chloroacetanilide herbicide metabolites
Drinking water quality problems
State-of-the-art 5
2.2. Micropollutants as endocrine-disrupting compounds
Part of the above discussed micropollutants is classified as endocrine disrupting compounds
(EDC). In 1991, the European Commission defined this class of compounds as “exogeneous
substances that cause adverse health effects in an intact organism, or its progeny, consequent
to changes in endocrine function” (Fawell et al., 2001). The importance of detecting EDCs
was emphasized through the development of specific biotests, (e.g. specialized to identify
compounds with endocrine disrupting character), pointing out to the high biological activity
of some classes of micropollutants (e.g. synthetic hormones, alkylphenol polyethoxylates).
The ubiquitous distribution of EDCs in the environment and their harmful potential
appeared obvious consecutively to the development of very sensitive biological tests based on
immunological techniques (e.g. ELISA) or hormonal activity of the targeted compounds (e.g.
yeast estrogen screen YES) (Huang & Sedlak, 2001; Aerni et al., 2004; Bringolf &
Summerfelt, 2003; Matsunaga et al., 2003), E-SCREEN assay (Soto et al., 1995), EROD
activity assay (Ma et al., 2005) or combinations of different bioassays (Oh et al., 2006). Part
of these studies concluded that some of the compounds, e.g., alkylphenol ethoxylates,
bisphenol A (BPA), estrone (E1), 17β-estradiol (E2), and 17α-ethinylestradiol (EE2) can
possess high estrogenic activity even at extremely low concentrations (Purdom et al., 1994;
Jobling et al., 1998). Depending on the exposure dose and their mechanisms of action, the
EDCs are responsible for a wide range of adverse effects on aquatic organisms, e.g.
feminization of male fish, masculinisation of snails (Desbrow et al., 1998, Koerner et al.,
2000; Korner et al., 2001; Rajapakse et al., 2002), growth inhibition (Halling-Sørensen, 2000;
Cleuvers, 2005), immobility (Cleuvers, 2004), mutagenicity, mortality (Robinson et al.,
2005), and changes in population density (Kidd et al., 2007).
2.3. Conventional treatment plant and membrane bioreactor for
wastewater treatment
The aim of wastewater treatment is the removal of bulk organic matter (proteins,
carbohydrates, etc.) and nutrients (Tchobanoglous et al., 2004). In activated sludge from
conventional treatment plant (CTP) and MBR, organics are subject to sorption, volatilisation,
and chemical or biological degradation, while inorganics undergo sorption, chemical
transformation, precipitation and assimilation by sludge-microflora. In both systems, activated
sludge consists mainly of flocculating microorganisms held in suspension and contact with
the wastewater in mixed aerated tanks. Before the biological step in CTP, the wastewater
6 State-of-the-art
influent first passes a mechanical step where large particles are removed from water and then
is directed to a primary sedimentation stage where it is allowed to pass slowly through large
tanks. Afterwards, the wastewater is sent to the aeration tank and finally remaining suspended
flocs are removed through a separation step, which is achieved by gravity sedimentation in an
external clarifier.
The main difference between CTP and MBR is the sludge-liquid separation step
(Figure 2.1). In the MBR, the activated sludge tank includes a filtration step through
submerged micro- or ultrafiltration membrane modules which retains the solid particles in the
aeration tank and saves space by replacing the secondary separations tanks (Figure 2.2). There
are four main types of membranes that are used in industry: plate and frame, spiral, tubular,
and hollow fiber membranes. Spiral membranes offer lower energy costs due to reduced
required pumping and higher packing densities and they can also be operated at higher
pressures and temperatures if necessary. Tubular membranes can be used at high solid
concentrations because plugging is minimized. Hollow fiber membranes can be back-flushed
to maximize the cross flow filtration process. The frame membranes possess good resistance
to fouling, are easy to clean but are quite expensive and show a low pressure resistance. The biomass separation technique considerably influences the quality of wastewater
effluent (Clara et al., 2005a). Generally, CTPs are operated at 1 to 5 g mixed liquor suspended
solids (MLSS)/L of sludge, while MLSS concentration in MBR is considerably higher,
ranging from 8 to 25 g/L or even more (Stephenson et al., 2000; Galil et al., 2003;
Ivashechkin et al., 2004). MBR technology allows sewage water treatment at high MLSS
concentration due to the membrane separation step and it is not limited by the sedimentation
capacity of the secondary clarifier. As biomass continuously grows, excess sludge has to be
removed from the system in order to maintain a constant concentration of microorganisms in
the tank. One of the main parameters of activated sludge systems is the sludge retention time
(SRT), which is controlled via the removal of excess sludge. High SRTs generally correlate
with high performance of the wastewater treatment concerning COD removal. Usually, SRTs
up to 25 (in some cases 80) days are applied to MBR processes, while typical SRT values for
a CTP vary from 8 to 25 days (Winnen et al., 1996; Cote et al., 1997; Cicek et al., 1999;
Stephenson et al., 2000; Clara et al., 2004a; Johnson & Williams, 2004; Joss et al., 2005).
Due to the high SRT values and complete retention of solids inside of MBR, biodiversity of
the microorganisms is favoured and even slowly growing and free living bacteria remain in
the system (Clara et al., 2005b; Pollice & Laera, 2005; Howell et al., 2003). Furthermore, the
adaptation of some microorganisms for the degradation of persistent compounds contained in
State-of-the-art 7
sewage, e.g. nonylphenol (NP) and further estrogens, is assumed to be more likely in MBR
than in CTP (De Wever et al., 2004; Ivasheckin et al., 2004; Siegrist et al., 2004).
Figure 2.1. Conventional wastewater treatment system and membrane bioreactor. Comparative schemes.
8 State-of-the-art
Figure 2.2. Types of membranes used in the treatment of wastewater by MBR. Left: membrane plates type Kubota. Right: Hollow fibre type Zenon.
Recognized advantage of MBR is the high effluent quality in terms of COD, nitrogen,
phosphorous, ammonia, retention of suspended solids and microorganisms, reliable biomass
concentration, efficient treatment of complex waste streams and compactness of the
installation and possibility to be enlarged (Cicek et al., 1999; Abegglen & Siegrist, 2006;
Cornel & Krause, 2006). At the opposite, the high MLSS concentration used in MBR leads to
problems concerning oxygen supply of the microorganisms and the membranes require
frequent cleaning and high maintenance costs (Cicek et al., 2001; Cornel & Krause, 2006).
The factors involved in the performances of CTP and MBR are listed and exemplified in the
upcoming paragraph.
2.4. Factors affecting the removal of micropollutants from wastewater
2.4.1. Hydrophobicity and hydrophilicity
Hydrophobicity refers to the physical property of a molecule that is repelled from a mass of
water. Many organic micropollutants found in wastewater are hydrophobic compounds.
Hydrophobicity is the main property, which determines the bioavailability of a compound in
aquatic environment; and it is involved in the removal of pollutants during the wastewater
treatment through sorption and biodegradation processes (Garcia et al., 2002; Ilani et al.,
2005; Yu & Huang, 2005). The octanol-water partition coefficient Kow, reflects the
equilibrium of partitioning of the organic solute between the organic phase, i.e. octanol, and
the water phase (Lion et al., 1990; Stangroom et al., 2000; Yoon et al., 2004). High Kow is
State-of-the-art 9
characteristic for hydrophobic compounds, and implies poor hydrosolubility and high
tendency to sorb on organic material of the sludge matrix (Lion et al., 1990; Stangroom et al.,
2000; Yoon et al., 2004). For compounds with log Kow < 2.5 the organic compound is
characterized by high bioavailability and its sorption to activated sludge is not expected to
contribute significantly to the removal of the pollutants via excess sludge withdrawal. For
chemicals having log Kow between 2.5 and 4, moderate sorption is expected. Organic
compounds with log Kow values higher than 4.0 display high sorption potential (Rogers 1996;
Ter Laak et al., 2005). A wide-range of practical examples depicts the interdependency
between the log Kow of a compound and its properties of sorption to organic matter in
wastewater. Yamamoto et al. (2003) showed that the fate of compounds with log Kow ranging
from 2.5 to 4.5 (e.g. E2, E3, EE2, octylphenol (OP, BPA, and NP) is highly correlated to their
respective log Kow. Log Kow values of the compounds govern their sorption/desorption
behaviour and diffusion to remote sites. The latter are sometimes inaccessible to
microorganisms and are often also nonextractable using classical chemical extraction
procedure. A lot of information can be retrieved from fate studies concerning the formation of
pesticides residues in soil residues (e.g. lindan), where the degradation by microorganisms
and the run-off with precipitation is limited by such immobilization processes (Scholtz &
Bidleman, 2007; Kurt-Karakus et al., 2006).
2.4.2. Bioavailability
As biodegradation is the primary elimination pathway for organics in the activated sludge
treatment, bioavailability of xenobiotics to degrading microorganisms is one of the most
important prerequisite for a trace pollutant occurring in wastewater (Vinken et al., 2004;
Burgess et al., 2005). In general, the accessibility of micropollutants to the population of the
activated sludge can be defined in terms of external and internal bioavailability. External
bioavailability rather defines the accessibility of the substance to microorganisms while
internal bioavailability is limited to the uptake of the compounds into the internal cell
compartment.
In general, bioavailability consists of the combination of physico-chemical aspects
related to phase distribution and mass transfer, and of physiological aspects related to
microorganisms such as the permeability of their membranes, the presence of active uptake
systems, their enzymatic equipment and ability to excrete enzymes and biosurfactants
(Wallberg et al., 2001; Cavret & Feidt, 2005; Del Vento & Dachs, 2002; Ehlers & Loibner,
2006). High bioavailability and thus potential for biological degradation of pollutants depends
10 State-of-the-art
mainly on the solubility of these compounds in aqueous medium. The bioavailability of
organic pollutants in aqueous environment is influenced by the presence of different forms of
organic carbon like cellulose or humic acids (Burgess et al., 2005). For instance, the extensive
formation of associates consisting of one branched isomer of nonylphenol and humic acids
was observed, whereas no interactions occurred when the compound was incubated with
fulvic acids. It was assumed that the association between nonylphenol and humic acids occurs
through rapid and reversible hydrophobic interactions (Vinken et al., 2004). Another study
indicated that the apparent solubility of NP was enhanced through its association with humic
acids and the compound was less prone to volatility (Li et al., 2007). Consequently, the extent
of mineralization of this xenobiotic was increased by 15% in presence of organic matter.
Furthermore, bioavailability can be stimulated by the use of artificial surfactants. One
practical application is the use of surfactants to enhance oil recovery by increasing the
apparent solubility of petroleum components and efficient reduction of the interfacial tensions
of oil and water in situ (Singh et al., 2007). The desorption of polycyclic aromatic
hydrocarbons (PAHs) was increased by addition of non-ionic surfactants in soil-water
systems and the formation of dissolved organic matter (DOM)-surfactant complexes in the
soil-water system is a possible reason to explain the enhanced desorption of PAHs (Cheng &
Wong, 2006).
2.4.3. Sorption
Sorption mainly occurs via absorption and adsorption mechanisms. Absorption involves
hydrophobic interactions of the aliphatic and aromatic groups of compounds with the
lipophilic cell membrane of some microorganisms and the fat fractions of the sludge.
Adsorption takes place due to electrostatic interactions of positively charged groups (e.g.
amino groups) with the negative charges at the surface of the microorganisms’ membrane.
The quantity of a substance sorbed Csorbed [g/L] is usually expressed as a simplified linear
equation (1) (Siegrist et al., 2004).
dissolveddsorbed CSSKC ⋅⋅= (1)
Kd is the sorption constant [L/g], which is defined as the partitioning of a compound between
the sludge and the water phase. SS [g/L] represents the concentration of suspended solids in
the raw wastewater and Cdissolved [g/L] is the dissolved concentration of the substance.
An example for a substance which strongly sorbs to suspended solids is the antibiotic
norfloxacin. Its sorption relies on electrostatic interactions between positive charged amino
State-of-the-art 11
group of norfloxacin and the negatively charged surface of microorganisms (Golet et al.,
2003).
Different types of sludge play a role concerning the sorption. For instance,
microorganisms in the secondary sludge represent the greater proportion of the suspended
solids resulting in a relatively high sorption constant of Kd = 25 L/g. In comparison, for the
primary sludge the sorption constant of norfloxacin is only Kd = 2. Based on the same amount
of suspended solids, the primary sludge has an essentially fewer content of microorganisms
but contains a large fat fraction. Thus, only 20% of norfloxacin is sorbed to the primary
sludge (Poseidon Final Report, 2004).
2.4.4. Biodegradation
Biodegradation defines the reaction processes mediated by microbial activity (biotic
reactions). In aerobic processes, microorganisms can transform organic molecules via
oxidation reactions to simpler products, for instance other organic molecules or even CO2
through mineralization (Siegrist et al., 2004; van der Meer, 2006). At low concentrations, the
kinetics of decomposition of trace pollutants occurs mostly according to a first order reaction
(see equation 2, Siegrist et al., 2004).
Rdegradation = Kdegradation • SS • Cmicropollutant (2)
Rdegradation is the degradation rate, Kdegradation is the degradation constant, SS [g/L] is the
concentration of suspended solids and Cmicropollutant [mg/L] is the concentration of
micropollutants supposed to be degraded.
The degradation rates are strongly dependent upon environmental conditions, such as the
redox potential of the system and the microbial population present. The process takes time for
acclimatization of microorganisms to the substrate. The affinity of the bacterial enzymes to
the trace pollutants in the activated sludge influences the pollutant transformation or
decomposition (Spain et al., 1980; Matsumura, 1989).
Biological degradation of pollutants occurs via two mechanisms (Tchobanoglous et al., 2004).
The first is the mixed substrate growth for which the bacteria use the trace substance as a
carbon source and energy source and hence, the compound is mineralized. The second
possibility is the co-metabolism for which the bacteria only break down or convert the
substance and do not use it as carbon source, while using other present substrates.
12 State-of-the-art
For organic compounds the probability of biodegradation generally increases with the
age of the sludge. From a variety of compounds studied, bezafibrate, sulfomethoxazol and
ibuprofen were the only compounds degraded at SRT of 2-5 days (Poseidon Final Report,
2004). By increasing the SRT to 5-15 days also diclofenac and iopromide were degraded,
while carbamazepine and diazepam remained non degradable even at SRT < 20 days. One
interpretation of the authors was a more diversified bacterial biocoenosis with enlarged SRTs
because also slowly growing bacteria could colonize such biosystems. Furthermore, the
bacteria become able to assimilate more complex, less easily degradable compounds with
increasing sludge age (Siegrist et al., 2004). However, biodegradation can be impaired in
spite of a high sludge age if easily degradable substrates are present. The same problem can
occur during periods of increased substrates loading (Andersen et al., 2003).
2.4.5. Abiotic degradation and volatilization
Abiotic degradation comprises the degradation of organic chemicals via chemical (e.g.
hydrolysis) or physical (e.g. photolysis) reactions (Acher, 1985; Doll & Frimmel, 2003; Iesce
et al., 2006) and can be man induced and natural processes. Abiotic processes are not
mediated by bacteria and have been found to be of fairly limited importance in wastewater
relatively to the biodegradation of micropollutants (Stangroom et al., 2000; Lalah et al., 2003;
Ivashechkin et al., 2004; Soares et al., 2006; Katsoyiannis & Samara, 2007). The elimination
of trace pollutants by volatilization during the activated sludge process depends on the
dimensionless Henry’s constant (H) and Kow of the analyzed trace pollutant, and becomes
significant when the Henry’s law constant ranges from 10-2 to 10-3 (Stenstrom et al., 1989). At
very low H/Kow ratio, the compound tends to be retained by particles (Roger, 1996; Galassi et
al., 1997). The rate of volatilization is also affected by the gas flow rate and therefore,
submerged aeration systems such as fine bubble diffusers should be used to minimize
volatilization rates in wastewater treatment plants (Stenstrom et al., 1989).
2.4.6. Compound structure
The structure of a compound and its elementary composition can influence the elimination
rates in wastewater. A compound with simple chemical structure will be prone to removal via
degradation during the wastewater treatment, while a compound with complex structure is
likely to persist as parent compound or partially degraded compound in sewage water. For
instance, pharmaceuticals are complex molecules and are notably characterized by their ionic
nature. Compounds such as ketoprofen and naproxen were not eliminated during CTP
State-of-the-art 13
process, while MBR treatment led to their elimination (Kimura et al., 2007). It was assumed
that the poor removal in CTP is due to their complex molecules including two aromatic rings
making the compound more resistant to degradation processes. Compounds like clofibric acid
and diclofenac are small molecules harbouring chlorine groups and were not efficiently
removed by both CTP and MBR. Therefore, the recalcitrance of these pharmaceuticals and
personal care products (PPCP) was attributed to the presence of halogen groups, based on
many observations with other chemicals substances. On the basis of the removal extent and
the chemical structure the same authors proposed a classification of PPCP into compounds,
which are easily removed by both CTP and MBR (i.e. ibuprofen), not efficiently removed in
both systems (i.e. clofibric acid, dichlofenac), and not satisfactorily removed by CTP but well
removed by MBR (i.e. ketoprofen, mefenamic acid, and naproxen).
The extent of removal of polar compounds such as naphthalene sulphonates (anionic
surfactants) during MBR treatment depends strongly on their respective molecular structure
(Reemtsma et al., 2002). The removal of the naphthalene monosulphonates was almost
complete, while it was limited to about 40% in the case of naphthalene disulfonates.
Degradation and partitioning behaviour have also been reported to be a function of the polar
to non-polar group ratio of the molecule, the presence of aromatic moieties (Chiou et al.,
1998), and the organic carbon content (Yamamoto et al., 2003) which characterize the
molecule. Linear alkylbenzene sulphonates (LAS) with long alkyl chains were preferentially
adsorbed to the sludge matrix, while the short homologues of this anionic surfactant were
found in the effluent in a comparative study in CTP and MBR (Terzic et al., 2005). On the
whole, the chemical structure of an organic pollutant does not only provide information
concerning the class to which the compound belongs, but also indications on the
degradability, chemical stability, volatility and sorption of an xenobiotic reaching the aquatic
environment.
2.4.7. Sludge retention time (SRT) and biomass concentration
The sludge retention time (SRT) is the mean residence time of microorganisms in activated
sludge wastewater treatment systems. A study states that a sufficiently high SRT is essential
for the removal and degradation of micropollutants in wastewater (Joss et al., 2005). High
SRTs allow the enrichment of slowly growing bacteria and also the establishment of a more
diverse biocoenosis able to degrade a large number of micropollutants. It was demonstrated
that at short SRTs (< 8 days) slowly-growing bacteria are removed from the system and in
this case, the biodegradation is less significant and adsorption to sludge will be more
14 State-of-the-art
important (Jacobsen et al., 1993). For instance, nitrifying bacteria possess generation times of
minimum 8 days. Nitrification leads to the conversion of ammonia to nitrate under strong
aeration conditions and this process is mediated by endogeneous microorganisms in the
aerated tanks. Complete nitrification was demonstrated in MBRs at sludge ages of 5-72 days
and high organic loading rates of 0.05-0.66 kg BOD m-3d-1 (Fan et al., 2000). The Byrns
model (Byrns, 2001) concerning xenobiotics degradation shows that at low SRTs, most of the
compounds are removed through sludge discharge. As the SRT increases, the proportion of
sludge wasted from the system decreases and higher amounts of trace pollutants remain in the
system and are amenable for further degradation. Nevertheless, physico-chemical properties
of the compounds also play an important role.
In order to remove pharmaceuticals from wastewater, SRT is one of the parameters
easy to handle to improve the process efficiency. Two MBRs operated at high SRTs of 26
days showed removal efficacy of 43% for benzothiazoles (Kloepfer et al., 2004). By varying
the SRT in MBRs, Lesjean et al. (2005) noticed that the removal of pharmaceutical residues
increased with a high sludge age of 26 days and inversely decreased at lower SRT of 8 days.
SRT values between 5 and 15 days are required for biological transformation of some
pharmaceuticals, i.e., benzafibrate, sulfamethoxazole, ibuprofen, and acetylsalicylic acid
(Ternes et al., 2004). Nevertheless, the application of high SRT is not automatically leading to
the removal of all pollutants. For SRTs of 2 days, Clara et al. (2005a) found out that none of
the investigated pharmaceuticals, i.e. ibuprofen, benzothiazole, dichlorofenac and
carbamazepine was eliminated, while applying SRT of 82 days in MBR and 550 days in CTP,
removal rates above 80% were obtained. Nevertheless, removal rates of carbamazepine
remained below 20% for all applied SRTs. Similar results were reported by Ternes et al.
(2004) where carbamazepine and diazepam were not degraded even at SRT longer than 20
days. In MBR containing adapted sludge, the removal of diclofenac ranged from 44 to 85% at
SRTs of 190–212 days (Gonzales et al., 2006). The biodegradability of trimethoprim was
studied during different sewage treatment steps using batch systems (Perez et al., 2005). The
main outcome of this study was that the activated sludge treatment comprising the
nitrification process was the only treatment capable to eliminate trimethoprim. The capacity
of nitrifying bacteria to break down trimethoprim was quite unexpected, because a subsequent
study reported that this xenobiotic cannot be degraded by microorganisms (Junker et al.,
2006). The degradation of selected pharmaceuticals was tested in a lab scale MBR with high
sludge concentration ranging between 20 and 30 g/L and a SRT of 37 days (Quintana et al.,
2005). Bezafibrate was transformed (60%) but not mineralized and the metabolites were
State-of-the-art 15
tentatively identified. Naproxen was degraded over a period of 28 days. Ibuprofen
degradation started after 5 days and was complete after 22 days. Tetracyclines were highly
sorbed to sludge and the sorption correlated well with the SRTs during adsorption tests in a
batch system (Sithole & Guy, 1987). The adsorption kinetics for tetracyclines was determined
at various biomass concentrations in sequencing batch reactors at different SRTs and HRTs
(Kim et al., 2005). Between 75 and 95% of the applied tetracyclines was adsorbed onto the
sludge after 1 h. At long SRTs (10 days) the removal of tetracyclines was 85–86%, while the
decrease of the SRT to 3 days resulted in a lower removal rate (78%). The lower degradation
rates were assigned to the reduced biomass concentrations in system with shortened SRT.
2.4.8. Influence of biomass characteristics on sorption and biodegradation
The biomass characteristics are important factors for biodegradation and differ between CTP
and MBR treatment (Brindle & Stephenson, 1996). The possibility for genetic mutation and
adaptation of microorganisms to assimilate persistent organic compounds increases at higher
SRTs (Cicek et al., 2001). Furthermore, some enzymatic activities increase proportionally to
the higher specific surface of MLSS, which is directly related to the floc structure. Comparing
the MBR and CTP systems, Cicek and collaborators (1999) showed that the biomass in the
MBR has higher viable fraction than in the conventional treatment plant. This phenomenon
can be attributed to improved mass-transfer conditions in the MBR favoured by smaller flocs
and the presence of many free-living bacteria. The size of bacterial flocs contained in the
activated sludge can be another factor causing the difference between CTP and MBR
wastewater treatment processes. In MBR it varies between 10 and 100 µm, and in the CTP
between 100 and 500 µm (Zhang et al., 1997). In this study the specific flocs surface per unit
of reactor volume was ten times higher in MBR than in CTP systems. The small size of
microorganisms and the floc surface implies short distances to be overcome by the substrate
during the diffusion into the flocs.
The activated sludge composition varies with the influent composition and the
operating conditions adapted to the wastewater treatment system (Chang & Judd, 2003).
Changing the inflow from artificial to municipal wastewater provoked a prominent increase of
metazoan and ciliate abundances (Fried et al., 2000). Communities of ciliates and metazoan
can rapidly change as a function of plant type and respective operating conditions. Comparing
lab scale and pilot scale reactors, it could be pointed out that ciliate species numbers turned
out to be lower in small scale reactors compared to large scale.
16 State-of-the-art
2.4.9. pH value
The acidity or alkalinity of an aqueous environment can influence the removal of organic
micropollutants from wastewater by influencing the physiology of microorganisms, the
activity of extracellular enzymes, and the solubility of micropollutants present in wastewater.
Depending on their pKa values, pharmaceuticals can exist in various protonation states as a
consequence of pH variation in the aquatic environment. For instance, at pH 6–7 tetracyclines
are not charged and therefore, adsorption to sludge becomes an important removal mechanism
(Kim et al., 2005). It was also demonstrated that the hydrophobicity of norfloxacin varies
with the pH values, being very low at pH < 4 and pH > 10 and the maximal hydrophobic
value was reached at a pH of 7.5 (Advanced Chemistry Development, ACD Labs). Another
study identified the pH value as critical parameter affecting the removal of micropollutants
during the MBR treatment, pH values varied from neutral to acidic as nitrification became
significant in the MBR (Urase et al., 2005). At pH values below 6, high removal rates of up to
90% were observed for ibuprofen. Ketoprofen was removed from MBR up to 70% when the
pH dropped down below 5, but obviously these are exceptional conditions, not applicable to
the municipal wastewater treatment.
The sorption of E1 and E2 to the organic matrix was reported to be strongly dependent
on the pH value (Jensen & Schaefer, 2001). In these studies, 23% of the steroid estrogens
were sorbed to the activated sludge at a pH value of 8, while this proportion increased up to
55% when the pH value was maintained at 2 and it was shown that increasing pH value up to
the compounds’ pKa (pH > 9) lead to an increased desorption of steroids. The same behaviour
was observed in the investigations of Clara and collaborators (2004 a, b), where the
compound’s solubility increased at pH of 7–12. During the sludge treatment like sludge
dewatering and conditioning with lime, the pH is increasing above 9 and the micropollutants
can be desorbed from sludge solids. For instance, the recovery of BPA in the aqueous phase
took place at pH > 12 and desorption was attributed to the increased hydrosolubility of the
deprotonated form of BPA (Clara et al., 2004b; Ivashechkin et al., 2004). The consequence of
such high release was a high back-loading of the treatment plant via the recycling of the
process water. In another study, the sludge-water partition coefficients (Kp) of estrogens in
activated sludge from a CTP were increased with decreasing pH for almost all the
investigated compounds (BPA, E2, EE2) (Kikuta, 2004). In the case of compounds
harbouring one carboxyl group, the Kp values at pH 5.6 were 2.5-3 times higher than those at
pH 6.7, while for compounds with phenol groups such as E1, EE2, and BPA the increase of
the partition coefficient varied to a lower extent within this range of pH values.
State-of-the-art 17
Although few studies focused on this parameter, the control of the pH value might be a
solution for the removal of micropollutants in WWTPs. The pH of industrial wastewater is
often subject to variations and may also negatively influence the elimination of the
micropollutants. One solution would be the adjustment of the pH of the influent before
biological treatment (Tchobanoglous et al., 2004). Enhanced removal rates of deprotonable
micropollutants from wastewater can be achieved at low pH values, as the protonation state
influences both sorption and degradation processes. On the one hand, acidic conditions are
not usual in CTP or MBR, but could be adapted for systems treating wastewater from highly
contaminated sites or industrial wastewater in order to increase degradation rates. On the
other hand, since the dissolution of a compound can be controlled by varying the pH value,
one can use this advantage in order to avoid further contamination. A possible alternative can
be the pH adjustment of the sludge to alkaline values prior usage for soil amendment in
agricultural applications.
2.4.10. Temperature
In general, biodegradation and sorption of organic micropollutants are temperature dependent
processes. The temperature influences the microbial activity in wastewater treatment systems
as microbial growth rates strongly vary according to the applied temperature conditions (Price
& Sowers, 2004). The rate of change of a biological or chemical system as a consequence of
increasing the temperature by 10°C is known as temperature coefficient (Q10). For most
compounds, sorption equilibrium decreases with increase of temperature, whereas the
biodegradation is less efficient at lower temperatures. For an exhaustive review, the readers
are referred to the publication of Hulscher & Corelissen (1996). The temperature influences
the solubility and further physicochemical properties of micropollutants as well as the
structure of the bacterial community. Concerning the elimination of micropollutants, it seems
that CTP shows better stability than MBR during seasonal temperature variations. The larger
surface of CTP in comparison to MBR would attenuate the variations of temperature and thus,
protect bacterial activity against temperature shock produced in the system. The temperature
increase in microbial activity also favors higher biodegradation rates of micropollutants.
Wastewater treatment systems situated in areas with prolonged high temperature may
efficiently eliminate micropollutants.
In a recent study, the removal of pharmaceuticals, i.e., ibuprofen, benzafibrate,
diclofenac, naproxen, and ketoprofen was reported to increase during the summer time when
the temperature reached 17oC in comparison to the winter season when the water temperature
18 State-of-the-art
was around 7oC (Vieno et al., 2005). A temperature of 20oC was beneficial for the removal of
pharmaceuticals in CTP and MBR and for instance more than 90% of bezafibrate was
eliminated under the conditions applied (Clara et al., 2004a). At low temperature during
winter season, the degradation rates decreased. In the case of diclofenac, naproxen, and
ibuprofen better performances of removal are reached when the systems are operated at 25oC
than at 12oC (Carballa et al., 2005). In another study, comparing the removal of
pharmaceuticals (phenazone, carbamazepime and metabolites) during CTP and MBR process,
the performance of the CTP process remained relatively constant over time despite the
winter/summer changes of temperature (10–25oC), while in MBR removal rates were strongly
affected by the seasonal changes (Lesjean et al., 2005). The higher temperature registered in
summer in MBR and the long sludge age (26 days) led to the removal rates to 80–100%. The
same study states that the extent of removal in MBR units was up to 99% for pharmaceuticals
and up to 80% for the steroids initially present in the incoming wastewater during the summer
period.
The adsorption of the antibiotic fluoroquinolone to the particles in the raw water is
influenced by the temperature. Lindberg et al. (2006) observed 80% adsorption at 12oC, while
Golet et al. (2003) showed an adsorption of 33% when the temperature was higher. Studies on
the influence of high temperature on the removal of COD from wastewater of pharmaceutical
industry led to the conclusion that temperature serves as pressure of selection for the bacterial
community development during aerobic biological wastewater treatment (LaPara et al.,
2001). In the same time, it stimulates higher degradation rates of pharmaceuticals.
Temperature was reported to influence also the mineralization of E2 (Layton et al.,
2000). An increase of 10oC leads to a duplication of microbial activity (based on Arrhenius
equation) and mineralization rates. Changes of approximately 15oC had statistically
significant effect on the mineralization rate of E2 present in the aqueous phase.
The temperature of wastewater treatment seems to play an important role concerning
the removal of xenobiotics; WWTP in countries with average temperature of 15-20oC may
better eliminate micropollutants than those in cold countries with a mean temperature mostly
under 10oC (the seasonal temperature changes between summer and winter are critical for
biodegradation processes and generally for the removal of pollutants).
State-of-the-art 19
2.5. Fate of model micropollutants in wastewater treatment plants
2.5.1. Fate of nonylphenol in wastewater treatment plants
Nonylphenol: occurrence, environmental concentration and effects
Nonylphenol (NP) is a starting material to produce the non-ionic surfactants nonylphenol
polyethoxylates (NPnEO) used in household and commercial products. NPnEO have been
used over decades in the preparation of a wide range of consumer products (e.g. personal care,
laundry products and cleaners), commercial products (e.g. floor and surface cleaners), and in
many industrial cleaning processes (e.g. textile scouring and preparation). The NPnEO are
produced by coupling ethoxy groups of various chain lengths to NP. In 1997, 78,500 tons of
NP was produced in Europe and more than half of this amount was used for the production of
NPnEO (EUR20387/EN, 2002).
The history of the NP problematic started with a report describing this compound as a
persistent pollutant in anaerobically treated sewage sludge (Giger et al., 1984). It was stated
that the biodegradation of NPnEO can lead to toxic metabolites (e.g. NP) during the
mesophilic anaerobe sludge stabilization in WWTP and can afterwards be released in the
environment. Reports on aerobic formation of NP from NPnEO are rather scarce. Only
indirect proof is available for the formation of NP under these conditions. There are some
publications known which report an actual slight increase in NP concentration during NPnEO
degradation in well-aerated environments (Jones et al., 1998; Corsi et al., 2003). In lab
studies NP has not been detected as a product of the aerobic degradation of NPnEO (Mann &
Boddy, 2000; Jonkers et al., 2001). In conclusion, NP is believed to be formed only under
anaerobic conditions during the sludge treatment and returned back to the treatment plant with
the leachate (Jonsson et al., 2003).
NP is recognized as hazardous chemical for its adverse effects on the endocrine
system of aquatic organisms above concentration of 8.3 µg/L in fish (Soto et al., 1995).
Beside the disruption of hormonal system of living organisms the need for (eco)toxicological
studies was emphasized since other investigations concluded on the toxicity and
bioaccumulation of NP in aquatic species (Ekelund et al., 1990; Hu et al., 2005). On the
whole, NP is assumed to pose environmental risks because of its toxicity to aquatic
organisms, its xenoestrogenic effects in aquatic fauna, and its lipophilicity which is
responsible for the bioaccumulation potential and its persistence (half-life of 20-104 days in
sediment) (Hecht et al., 2004). In 2000, concentrations of NP in wastewater were estimated to
range between 0.1 and 3.6 µg/L in Germany (Koener et al., 2000), 1-14 µg/L in Switzerland
(Johnson et al., 2005) and 0.7 to 2.6 µg/L in Italy (Bruno et al., 2002). Since 2005, the use of
20 State-of-the-art
NP in Europe is strongly restricted in order to decrease the residual concentration of NP
reaching WWTP (Directive 2003/53/EC).
NP used for the production of NPnEO is in fact a technical mixture (tNP) consisting of
a wide range of isomers, whereby the alkyl chain structure varies. According to Guenther et
al., (2006), 211 possible constitutional isomers of NP are numbered based on a hierarchical
and logical system. Many of these isomers possess up to three chiral C-atoms, in total 550
compounds are possible (Guenther et al., 2006). Many of the structures could be identified
using cutting edge analytical method such as GCxGC-MS, LC/LC-MS (Moeder et al., 2006,
Zhang et al., 2007). The structure of 23 isomers were identified and characterised (Wheeler et
al., 1997, Thiele et al., 2004)
NP is a hydrophobic compound due to the presence of the nonyl chain, (log Kow =
4.48) (Table 2.2) and its tendency to sorb to different materials inclusive glassware is an
established fact (Ahel et al., 1993, Vinken et al., 2004). The strong adsorption can impair the
determination of NP concentration in environmental samples as well as in mineral media.
This behaviour led several authors to overestimate the degradation rates of NP during several
fate studies (Clara et al., 2005a, Terzic et al., 2005). Careful examination of the experimental
procedures applied for the extraction and the measurements should not be neglected when
judging on the relevance of the results. Further physico-chemical properties can also hamper
the accurate determination of NP concentration.
NP is a semi-volatile organic compound capable of relatively reduced water/air
exchange. Porter and Hayden (2001) examined volatilization as a mechanism responsible for
the loss of NP during wastewater treatment and concluded that up to 3% of NP in wastewater
influent could be released into the atmosphere.
Table 2.2. Physico-chemical properties of nonylphenol.
Molecular formula: C15H24O Molecular weight: 220.34 g/mol Density at 25oC: 0.952 g/cm3 (a)
Solubility in water: 4.9 mg/L (b)
pKa: 10.28 (c)
Partition coefficient (log Kow): 4.48 (d)
Vapor pressure at 25oC: 2.07·10-2 Pa (e)
Boiling point: 310oC (f)
Viscosity at 20oC: 3160 mPa·s (a)
Henry’s law constant:8.39·10-1 Pa·m3/mol (e)
a(Fiege et al., 2000); b(Brix et al., 2001); c(Muller & Schlatter, 1998); d(Ahel et al., 1993); e(Shiu et al., 1994); f(Lorenc, 2003).
State-of-the-art 21
Removal of nonylphenol in wastewater treatment systems
The removal of nonylphenol is often given in the literature with respect to sorption and
biodegradation. Many research studies focused on the removal of NP from wastewater in
conventional treatment plants and membrane bioreactors (Wintgens et al., 2004; Clara et al.,
2005a; Nakada et al., 2006). Unfortunately, many of the studies were performed at different
scale, used different parameters and more important, the real comparison is often not possible
since important gaps in data sets and interpretations can be observed (Terzic et al., 2005;
Gonzalez et al., 2007; Hernandez-Raquet et al., 2007). Many examples from the literature are
based on the results of analytical techniques used for the measurements of grab samples from
influent and effluent of WWTP. Often, the removal rates in these studies are based on the
difference between inflow and outflow concentration. In such a black box the respective
extent of sorption or biodegradation occurring during the WWTP steps remain unknown
(Esperanza et al., 2004; Nakada et al., 2006).
The technical mixture of NP is recalcitrant to degradation because more than 85% of
the isomers possess a quaternary α-carbon on the branched alkyl chain (Wheeler et al., 1997).
Such stable structure is resistant to ω- and β-oxidation of the nonyl chain (Van Ginkel, 1996).
This is considered the main reason for the different biodegradability of tNP and the linear
alkyl chain NP (4-n-NP) which is not constituent of tNP. It was reported that the degradation
of branched isomers of NP is resulting in metabolites with incomplete degraded alkyl chain
while the ultimate degradation of the linear NP isomer is faster (Cirja et al., 2006; Corvini et
al., 2006a).
Starting from the problemacy concerning the real fate of NP in WWTP, in the next
paragraph, the removal of NP in terms of sorption and biodegradation is presented in details.
Particular attention is paid to some operational parameters of the wastewater treatment
process, which are assumed to play a role in the removal performance, e.g. sludge retention
time, biomass concentration or temperature of WWTP system.
Sludge retention time
The estimation of NP removal is often correlated to the SRT at which the wastewater
treatment system is operated. Johnson and his collaborators (2005) operated many CTPs in
order to evaluate the removal of NP and further endocrine disrupters from wastewater. Good
degradation of the investigated compounds was registered at a high SRT of 30 days. In the
same study it was shown that no significant difference is observed between the MBR and
CTP in term of degradation performances. Similar observation on the removal of NP was
22 State-of-the-art
described by Ivashechkin and collaborators (2004) operating MBR and CTP processes at
SRTs of 12 and 25 days with degradation rates around 80%. The high removal efficiency
(95%) associated to the operation of a MBR at high SRTS was confirmed as well by Terzic et
al. (2005) during the investigation of NPnEO elimination from wastewater.
From all pilot-scale studies it remains unclear whether NP is really biodegraded or
simply removed by withdrawing excess sludge. A reliable determination of NP adsorbed and
entrapped into sludge matrix is often difficult to achieve by means of classic analytical
methods and in this way the most results are resumed at giving removal rates of NP from
water phase without further insight.
In general, the SRT is directly depending on the biomass concentration at which the
WWTP is operated, as described further.
Biomass concentration
The biomass concentration in wastewater treatment system is playing and important role for
the removal rate of NP from wastewater through biodegradation and sorption. It is controlled
via the operational SRT applied to the process. Ahel et al. (1994) measured the concentration
of NP in sewage treatment plants and they determined that 44 to 48% of NP was adsorbed to
the sludge in different running systems. Similar removal rates were observed during a study
investigating the performances of CTP and MBR in parallel (Clara et al., 2005b). CTP was
operated with SRTs of up to 237 days and a biomass concentration of 4.0 g/L, while SRT of
up to 55 days and a biomass concentration of 11.8 g/L were applied to the MBR treatment.
Other two CTPs with SRTs lower than 50 days and biomass concentrations between 3 and 4
g/L did not show any removal performance (Clara et al., 2005b). A similar situation was
reported by Gonzalez et al. (2006) where the removal of NP and NPnEOs was higher in MBR
(96%) with mixed liquor solid suspension concentration inside of the reactor ranging from 11
to 20 g/L, while in CTP a removal rate of around 54% was reported with a biomass
concentration maintained at 4 g/L. A lab-scale experiment showed sorption to the sludge
matrix of 60 and 90% (Tanghe et al., 1999), when the sludge concentration was 3-4 g/L and
operated at 10-15oC while 90% removal was registered also in investigation of two CTP pilot
plants (one with aerobic and the other anaerobic sludge digestion). The performance of this
study was correlated with the use of high solid suspension concentrations of 16 g/L in primary
sludge and 2.1 g/L in secondary sludge (Esperanza et al., 2004). In a pilot scale MBR with
nanofiltration membranes and 22 g/L TSS treating the dumpsite leachate, 80 to 85% NP was
eliminated (Wintgens et al., 2002). In this case, the authors assumed that the adsorption of the
State-of-the-art 23
compound on suspended matter in the bioreactor was the main removal pathway. It is obvious
from these studies that the removal via the adsorption to the suspended solids in the
wastewater treatment system cannot be avoided. The consequences are positive for the
effluent but the source of pollution remain active for field application of the sludge still
practiced nowadays in some countries.
Temperature
The removal of NPnEO from municipal wastewater was investigated in dependence on
temperature variation by Terzic et al. (2005). The results of this study showed positive
influence on process efficiency up to 95% and the efficacy improved as the temperature in the
treatment system varied from 3 to 30oC. In the case of NP it was stated that at 10-15oC, which
are typical temperature values for Europe, the compound is preferentially distributed into the
sludge fraction (Brunner et al., 1988). The degradation of NP in a packed bed bioreactor
containing cold adapted bacteria (Soares et al., 2006) showed an optimal biodegradation rate
at a temperature of 10oC, conditions feasible for northern countries. By decreasing the
temperature from 10 to 5.5oC a negative effect on the bioreactor efficiency was observed. The
explanation was based on the lower diffusion of organic pollutants (limited solubility), which
is limited by the low temperature.
The increase of temperature stimulates the microorganisms on growth, activity and
obviously on the substrate consumption. This was investigated by Tanghe et al., (1999) where
the NP added at concentrations of 8.33 mg/L in a lab-scale system was almost totally removed
and biodegraded when the reactors were operated at 28oC. By lowering the temperature from
28 to 10-15oC, elimination capacities decreased to 13-86%, depending on the process
conditions. Upon re-establishing the temperature at 28oC, the accumulation of NP in the
sludge could be counteracted, proving the dependence of microbial growth and high
temperature.
Anaerobic degradation of nonylphenol in wastewater
As described above, NP source in the wastewater is the degradation of NPEOs. Complete
degradation of NPnEO occur under aerobic conditions (Jones et al., 1998) while, they have
been reported to be persistent in anaerobic environments (Giger et al., 1984) (Figure 2.3).
Ejlersson et al. (1999) characterized the products of NPEO degradation by mixed microbial
communities in anaerobic batch tests. Within seven days of anoxic incubation, nonylphenol
diethoxylate (NP2EO) appeared as the major degradation product. After 21 days, NP was the
24 State-of-the-art
main species detected, and was not degraded further even after 35 days. The anaerobic
degradation of NPnEO seems to be dependent on a nitrate source. The observed generation of
NP coupled to nitrate reduction suggests that the microbial consortium possessed an
alternative pathway for the degradation of NPEO (Luppi et al., 2007).
In another study NP1EO was spiked to an anaerobic digestion testing apparatus that
was operated at a retention time of approximately 28 days and a temperature of 35oC with
thickened sludge. Approximately 40% of NP1EO was converted to NP (Minamiyama et al.,
2006).
The further degradation of NP in activated sludge systems is rather scarce investigated
in the literature. Chang et al., (2005) investigated the effects of various parameters on the
anaerobic degradation of NP in sludge. The optimal pH for NP degradation in sludge was 7
and the degradation rate was enhanced when the temperature was increased. The addition of
aluminum sulfate (used for chemical treatment of wastewater) inhibited the NP degradation
rate within 84 days of incubation. The same study stated the theory that sulfate-reducing
bacteria, methanogen, and eubacteria are involved in the degradation of NP as the sulfate-
reducing bacteria are a major component of anaerobic treated sludge.
Figure 2.3. Degradation of NPnEO in the wastewater treatment plant (adapted from Corvini et al., 2006b; Giger et al., 1984).
State-of-the-art 25
Nonylphenol degradation by isolated microorganisms
NP can be degraded aerobically by specialized microorganisms like bacteria, yeast and
fungi under aerobic conditions (Yuan et al., 2004; Chang et al., 2005). Most of the NP-
degraders belong to sphingomonads. They were isolated from municipal wastewater treatment
plants and identified as Sphingomonas sp. Strain TTNP3, Sphingomonas cloacae,
Sphingomonas xenophaga strain Bayram (Tanghe et al., 1999; Fujii et al., 2001; Gabriel et
al., 2005). These bacterial strains are known for their capability of utilizing NP as a carbon
source. The rates of NP degradation are ranging from 29 mg NP/L·day (Tanghe et al., 1999)
up to 140 mg NP/L·day (Gabriel et al., 2005). The metabolites of NP degradation varied
depending on the bacterial strain and NP isomer used in the different studies (e.g. Candida
aquaetextoris degrades linear NP to 4-acetylphenol, Sphingomonas xenophaga Bayram
degrades branched NP to the corresponding branched alcohol and benzoquinone) (Vallini et
al., 2001; Gabriel et al., 2005).
The degradation pathways of NP by different Sphingomonads share similarities as the
corresponding alcohol of the alkyl side-chain of NP are degradation intermediates produced
by the isolated strains (Tanghe et al., 2000; Fujii et al., 2001; Gabriel et al., 2005).
Some scientists assumed that the degradation of para-NP (pNP) starts with the fission of the
aromatic ring (Tanghe et al., 1999; Fujii et al., 2001). Corvini et al., (2004) provided evidence
that hydroquinone is the central metabolite in the degradation pathway of NP isomers
(Corvini et al., 2006b, Figure 2.4). The degradation mechanism is a type II ipso substitution
requiring a free hydroxyl group at the para position. Hydroquinone is formed via
hydroxylation at C4 position of the alkyl chain, where alkylated quinol leaves the molecule as
a carbocation or radical species (Corvini et al., 2006b). Hydroquinone is further degraded to
organic acids, such as succinate and 3,4-dihydroxy butanedioic acid. The degradation of NP is
influenced by some factors like position, structure and length of the alkyl chain, concentration
of NP in the media, oxygen limitation or the temperature limitation (Tanghe et al., 1999;
Gabriel et al., 2005; Corvini et al., 2006).
26 State-of-the-art
Figure 2.4. Pathway for the degradation of NP isomers by Sphingomonas sp. strain TTNP3 (adapted from Corvini et al., 2006b).
2.5.2. Fate of 17α-ethinylestradiol in wastewater treatment plants
17α-ethinylestradiol: occurrence, environmental concentrations and effects
17α-ethinylestradiol (EE2) is a synthetic estrogen used as active agent of contraceptive pills
and is the most prescribed world-wide (Williams et al., 1996). A contraceptive pill contains
35 µg EE2 and up to 80% of this is eliminated as unmetabolised conjugates (Katzung, 1995).
EE2 is recognized as a compound with strong estrogenic activitiy. The contribution of EE2 to
the total amount of excreted estrogens in the water environment is only 1%, but this
compound is considerably more persistent in wastewater treatment systems compared to the
natural hormones. On the base of in vivo tests carried out in fish, the hormonal activity of EE2
was reported to be 11- to 130- folds higher than that of the natural estrogen estradiol (E2)
(Ternes et al., 1999).
Estimation of the maximum EE2 concentration to be recovered in wastewater was
13.4 ng/L, but the measured concentrations in municipal influents are generally lower than the
estimated values (de Mes et al., 2005). For example, measured values from 0.3 to 5.9 ng/L in
the Netherlands (Vethaak et al., 2002), from 1 to 11 ng/L in Sweden (Larsson et al., 1999)
and Germany (Ternes et al., 1999), and 42 ng/L in Canada (Ternes et al., 1999) were
reported.
The fate of EE2 is related to its physico-chemical characteristics (Table 2.3). Since the
log Kow is approximately 4, it is expected that the compound sorbs to sludge in high
proportion. Due to the presence of an additional ethinyl-group, the ring becomes extremely
State-of-the-art 27
stable against oxidation in comparison to the natural E2. Volatilisation is not expected to play
a significant role in the removal of EE2 from wastewater since its Henry’s law constant is
lower than 10-4 and H/Kow ratio is lower than 10-9 (Roger 1996).
Table 2.3. Physico-chemical properties of 17α-ethinylestradiol Molecular formula: C20H24O2
Molecular weight: 296.40 g/mol
Solubility in water: 4.8 mg/L (a)
pKa: 10.7 (b)
Partition coefficient (log Kow): 3.9 (c)
Vapour pressure at 25oC: 4.5E-011(a)
Henry’s law constant:7.94•10-12 Pa·m3/mol (d)
a(Lai et al., 2000); b(Clara et al., 2004); c(Jürgens et al., 2002); d(de Mes et al., 2005)
Removal of EE2 in wastewater treatments systems
Recent studies on the fate of estrogens during wastewater treatment investigate mainly their
degradation in activated sludge from municipal, conventional and membrane bioreactor
WWTP and rarely from industrial sewage treatment plants. These approaches lead to
disparate data sets concerning the degradation of EE2. For instance, elimination of 10% of
incoming EE2 was reported in a German WWTP (Ternes et al., 1999), while other studies
stated that the activated sludge process removes approximately 85% (Johnson & Sumpter,
2001). Some steps of the wastewater treatment process can be crucial for the removal of EE2.
Sludge retention time
Like for other compounds refractory to biodegradation, sludge retention time (SRT) is one
operational parameter which can influence EE2 degradation. Therefore, it is expected that a
long SRT supports high degradation rates of EE2. Based on this concept, SRT higher than
five days were applied in CTP and degradation rates of EE2 above 90% were obtained
(Kreuzinger et al., 2004). At SRT of twelve days and HRT of one day, a similar performance
of EE2 removal was reported from the investigations of Andersen et al. (2003). Reports on
the high sorption of EE2 to sludge particles of wastewater treatment systems are known
(Holbrook et al., 2004). Beside this, in the MBRs a relevant amount of EE2 from
contaminated influents is also sorbed to membranes. In fact, EE2 absorbed to microbial flocs
or colloids is retained by the ultra- and microfiltration membranes (Nghiem & Schaefer, 2002;
Clara et al., 2005b).
28 State-of-the-art
Satisfying removal rates ranging from 90% to 95% were reported in CTP and MBR
and no significant difference was observed between the two different wastewater treatment
systems (Clara et al., 2005b). The high performance of both treatments was attributed to the
high sludge age applied. With SRT of 12 to 15 days, both of the treatment systems were
adapted to the nitrification-denitrification process. In a full scale municipal plant including a
nitrification step, degradation rates of estrogens ranged between 79 and 95% and the extended
biodegradation was mainly related to the nitrifying activity (Vader et al., 2000). When SRT
are shorter than eight days, slow growing specific degraders are washed out and adsorption
remains the main process for the removal of EE2 (Jacobsen et al., 1993).
Biomass
The sorption of EE2 to different materials plays a significant role in the removal of EE2
(Yamamoto et al., 2003). Between 15 and 50% of the compound was bound to natural organic
material in natural waters, which contain typically 5 mg /L total organic carbon (TOC). Tests
carried out with activated sludge at MLSS concentrations of 2-5 g/L demonstrated that only
20% of EE2 remained in the aqueous phase after one hour of incubation (Layton et al., 2000)
and by use of 14C-labelling it could be shown that the compound was in fact mainly associated
to the sludge matrix. On the base of tests carried out with activated sludge from MBR and
CTP, the type of sludge was reported to affect significantly the extent of biodegradation of
EE2 (Joss et al., 2004). MBR sludge led to 2-3 fold faster transformation of estrogens than
that of CTP. This observation was interpreted as the consequence of the small size of flocs in
MBR sludge, which possess a high specific surface area favouring the transfer of pollutant
molecules to the degrading microorganisms. Kikuta et al. (2004) carried out CTP and MBR
treatment in order to remove EE2 from wastewater. The CTP was operated at 1 g/L MLSS
and HRT of 8 hours, while MBR was operated with higher MLSS of 10 g/L MLSS and
shorter HRT of 3 hours. The results indicated that 13% of EE2 was removed in the case of
CTP and 33% in MBR, while the rest of EE2 remained in the treated effluent. The authors of
this study concluded that the increase of MLSS concentration in MBR and the use of a longer
contact time (HRT) will help the migration of EE2 from the liquid phase to the sludge phase,
resulting in higher removal via sorption to sludge matrix.
Temperature
The influence of temperature was reported to influence the mineralization of estrogens during
activated sludge treatment (Layton et al., 2000). An increase of 10oC generally leads to a
State-of-the-art 29
doubled microbial activity (from Arrhenius equation) and should correspondingly improve the
mineralization rates of EE2. Variations of approximately 15oC had statistically significant
effect on the mineralization of natural hormones, while degradation rates of synthetic
estrogens like EE2 were not improved. The period from May to July indicated removal of
EE2 from 60 to 70% in CTP and MBR (Clara et al., 2004a). In the winter period, EE2
removal reached 60% in CTP, while no EE2 removal was observed in MBR. The inactivity of
microorganisms was assumed to be the consequence of temperature decrease. Similar results
were reported in a study of Desbrow et al. (1998); the biomass was less active during the
winter and high concentrations of EE2 were observed in the effluent.
Anaerobic degradation of ethinylestradiol
Research on the anaerobic degradation of EE2 until now is very rare. In river water samples
under anaerobic conditions EE2 was not degraded over a period of 46 days of incubation
(Jürgens et al., 2002). Another attempt to carry out investigations on anaerobic degradation of
estrogens, including EE2 was the batch test described in Poseidon project report (2004) with
sludge from an anaerobic sludge digester. The results did not show any degradation of the
investigated compounds.
Ethinylestradiol biodegradability by isolated specialized microorganisms
Recent studies have been dedicated to the isolation of microoganisms that can specifically
convert EE2. Many indices point out to relations between the nitrification processes and the
degradation of EE2. In a German WWTP, an EE2 removal efficiency of 90% was reported
and satisfying biodegradation occurred in the nitrifying tank (Andersen et al., 2003). A
significant degradation of EE2 was observed in nitrifying activated sludge, where the
ammonia-oxidizing bacterium Nitrosomonas europaea was present (Shi et al., 2004). These
results were in agreement with those of previous studies demonstrating the fast degradation of
EE2 with the concomitant formation of hydrophilic compounds in nitrifying activated sludge
(Vader et al., 2000). Additionally to reports on the positive influence of nitrifying bacteria
good degradation rates were obtained in axenic cultures of EE2-degrading fungi and bacteria
isolated from activated sludge. In some cases, the microorganisms could degrade up to 87%
of the added substrate (30 mg/L EE2) (Yoshimoto et al., 2004; Haiyan et al., 2007). The
fungus Fusarium proliferatum has been isolated from a cowshed sample and is capable of
converting 97% of EE2 at an initial concentration of 25 mg/L in 15 days at 30oC at an
optimum pH of 7.2 (Shi et al., 2002).
30 State-of-the-art
Recently, few studies have been conducted on the metabolic pathway of EE2
degradation. Shi et al., (2002) and Yoshimoto et al., (2004) reported the existence of EE2-
degradation products but did not identify them. Horinouchi et al., (2004) proposed a pathway
for the degradation of testosterone by Comamonas Testosteroni TA441, consisting of
aromatic-compound degrading genes for seco-steroids and 3-ketosteroid dehydrogenase.
Studies on the metabolisms of EE2 by the isolated bacterium Sphingobacterium sp. JCR5 led
to the identification of three metabolites. From these metabolites a degradation pathway for
EE2 could be concluded. Further metabolites, i.e. two acids (e.g. 2-hydroxy-2,4-dienevaleric
acid and 2-hydroxy-2,4-diene-1,6-dionic acids) were the main catabolic intermediates for the
formation of E1 (Shi et al., 2004; Haiyan et al., 2007). Furthermore, Shi el al. (2004) reported
that E2 was most easily degraded via E1 by nitrifying bacteria and it is assumed as being also
the degradation product of EE2. The catabolic pathway of EE2 by Sphingobacterium Sp.
JCR5 is proposed to start with the oxidation of C-17 to a ketone group and the hydroxylation
and ketonization of C-9α followed by the cleavage of B ring. A ring is hydroxylated by the
addition of molecular oxygen. The complete metabolic scheme is presented in figure 2.5
(Haiyan et al., 2007).
Figure 2.5. Proposed catabolic pathway for EE2 degradation by strain JCR5 (adapted from Haiyan et al., 2007).
31
3. Objectives and outline of this thesis
Studies on the fate of hydrophobic micropollutants presented in the literature are often
contradictory with respect to terms like removal or elimination from wastewater. Often we are
faced with contradictory analytical results or a lack of important information on the
performance of quantification. For instance, some studies are pointing out the relevance of the
SRT on the removal and biodegradation of hydrophobic micropollutants, but not all necessary
information concerning the system evaluation are provided. Furthermore, “missing” amounts
of the micropollutant are often associated to removal via volatilization. Data on the removal
of micropollutants from industrial wastewater is lacking, even if it is well known that the
composition of the wastewater and the microorganism communities responsible for the
biological treatment are different to that of municipal wastewater.
According to the literature, MBR processing is often indicated as superior to the
conventional wastewater treatment concerning the removal of hydrophobic micropollutants.
Starting from this statement, the challenge in the present work was on the one hand the
development and optimisation of a MBR in a small-scale format (1 L) able to reproduce the
functionality and performances of a real system. On the other hand this work should provide a
deeper and reliable insight into the real fate of micropollutants during the wastewater
treatment in MBR, with the help of radiomarkers.
In the frame of the present dissertation entitled “Studies on the behaviour of endocrine
disrupting compounds in a membrane bioreactor” the following points are intended to be
focussed:
• Design, development and optimization of a small-scale MBR with similar functionality
to that of a real scale system and variation of operational parameters in order to provide
satisfying results on wastewater treatment. In order to have reliable results, the wastewater
treatment system even at lab-scale size has to fulfil the characteristics of a real system. Would
it be possible to optimize and adapt the operational parameters and to maintain them constant
for long-term periods in a lab-scale MBR?
• Fate and balancing of nonylphenol in MBR simulating the treatment of municipal
wastewater; one pulse spiking strategy investigations. What happens with the radiolabelled
NP isomer spiked once in the MBR adapted to synthetic wastewater? Will it be possible to
32 Objectives and outline of this thesis
differentiate between real degradation and sorption just after one pulse of NP was applied
directly inside the MBR? How big is the risk that NP will reach the surface waters via the
effluent of MBR systems?
• Fate and balancing of 17α-ethinylestradiol in industrial adapted activated sludge of a
pharmaceutical factory; continuous adaptation of the system to EE2. An industrial
factory producing contraceptive pills, is supposed to release traces of EE2 into the
wastewater. How efficient is a MBR inoculated with industrial activated sludge to eliminate
EE2 from wastewater? Is this endocrine disrupter degraded by specialized degraders or will it
cross the treatment system.
• Attempts on the isolation and quantification of first metabolites formed by real
degradation of targeted compounds in a MBR system. Until now there are no known
metabolites of NP or EE2 formed in the real environment. Could it be feasible to quantify and
identify the chemical structure of such metabolites by the help of radio-markers? What kind
of compounds are they?
• Implementation of a bioaugmentation strategy for a MBR with specialized degrading
bacteria in order to enhance the biodegradation of NP. In pure culture, Sphingomonas sp.
strain TTNP3 showed high performances on fast NP degradation. Could this advantage be
used for practical application in real NP contaminated sites? Would it be possible that low
amounts of Sphingomonas will survive and degrade NP also in the complex matrix of a
MBR?
33
4. Materials and methods
4.1. List of used chemicals and instruments
4.1.1. Chemicals and materials
14C radiolabelled compounds
[U-Ring-14C] 4-[1-ethyl-1,3-dimethylpentyl]phenol
Synthesized for use in the present research work
(chapters 5.2 and 5.4)
Specific radioactivity 1.492 MBq/mg
Radiochemical purity: 98%
[U-14C] Phenol
Hartmann Analytic, Germany
Specific radioactivity: 2.70 GBq/mmol
Radiochemical purity: 99.7%
[U-Ring-14C] 17-α-ethinylestradiol
Bayer Schering Pharma, Germany
Specific radioactivity: 6.377 MBq/mg
Radiochemical purity: > 98%
[20-14C] (ethinyl) 17-α-ethinylestradiol
Hartmann Analytic, Germany
Specific radioactivity: 3.748 MBq/mg
Radiochemical purity: > 98%
Non-radiolabelled chemicals
3,5-Dimethyl-3-heptanol 99%, Avocado Research Chemicals Ltd., Great Britain
MSTFA N-Methyl-N-trimethylsilyl-trifluoroacetamid,
derivatization agent for GS, Fluka, Switzerland
Phenol 99.5%, Fluka, Germany
17α-ethinylestradiol 98%, Fluka, Germany
Sodium hydroxide Carl Roth, Germany
Monoethylene glycol Carl Roth, Germany
34 Materials and methods
Solvents
Acetonitrile HPLC grade > 99.9%, Rotisolv, Carl Roth, Germany
Ethyl acetate HPLC grade, Carl-Roth, Germany
4.1.2. Instruments and devices
Analytical devices
GS-MS Agilent 6890 gas chromatograph equipped with a 30
m*0.32 mm*0.25 µm phenyl methyl siloxane column,
0.46 µm film thickness, CS-Chromatography Service,
Germany; HP 5971A mass selective detector, Agilent
Technology, Germany
HPLC Agilent HPLC–System, equipped with a 250/4
Nucleosil 100-5 C18 HD column, Macherey Nagel;
250 mm x 4 mm with pre-column 20 mm x 4 mm, CS
Chromatographie Service, 5 µm Nucleosil C18-
material, Macherey-Nagel
HPLC radiodetector Radioisotope detector Ramona 171, Raytest, Germany,
with B11 29 370 TSX, 370 µL, cell volume 1300 µL
HPLC-MS/MS Thermo Finnigan surveyor LC-Pump with Synergi 4i
Fusion RP 80A column (150, 4.6 mm, 3 µm particle
size), Phenomenex, Torrance, CA
LC/MS with radiodetector Thermo Finnigan surveyor LC-Pump; Synergi 4 µ
Fusion RP 80A column (150 x 4.6 mm, 3 µm particle
size), and Luna C18 column (50 x 3 mm, 3 µm particle
size) Phenomenex, Torrance, CA, USA. 3139 quartz
cell and TSQ quantum ultra AM tandem MS, Thermo
Finnigan, USA
Liquid scintillation counter (LSC) LC-5000 TD, Beckman, Germany
Other used devices
Analytical balance P1200N, d = 0.01 g, Mettler–Toledo, Germany
Biological oxidizer OX500, RJ Harvey Instrument Corporation, USA
Materials and methods 35
Centrifuge J21C Centrifuge, rotor JA-14, Beckman, Germany,
Model ‘5414’, Eppendorf, Germany
Electrophoresis device (DGGE) BIORAD D-CODE™ Universal Mutation Detection
System, BIO-RAD Laboratories, USA
Drying stove KVTS11, Salvis, Swizerland
Horizontal shaker GFL 3017, Burgwedel, Germany
Lyophilizer Alpha 1-2, Braun Biotech International, Germany
MBR system Realized for the present work, including mechanical
and electronic components
Milli-Q Water System Millipore, Germany, with 0.22 µm membrane filter
Peristaltic pump MULTIFIX-Constant MC 1000 PEC, Alfred
Schwinherr KG, Germany
PCR Eppendorf® Mastercycler® personal, Eppendorf,
Germany
Rotary evaporator Rotavapor R-114 and RE, Buechi, Switzerland,
combined with membrane vacuum pump MZ 2C and
vacuum controller CVC2
Photometer DU-600 Photometer; Beckman, Germany
Spectrophotometer Labor photometer C 214, Hanna Instruments, Italy
Vacuum pump MZ 2C and vacuum controller CVC2
Tensiometer Krüss GmbH, Germany
4.1.3. Other materials
Activated sludge Inoculum from Soers Wastewater Treatment Plant,
Aachen
Agarose gel Eurogentec, Belgium
Bacterial strain Sphingomonas sp. strain TTNP3
Bacterial peptone Carl Roth, Karlsruhe
Beef extract Carl Roth, Karlsruhe
Glucose Carl Roth, Karlsruhe
Complex medium Standard I medium, Merck, Germany
DNeasy® Plant Mini Kit Qiagen, Germany
36 Materials and methods
Membrane cellulose acetate plates A3-Abfall-Abwasser-Anlagentechnik GmbH,
Germany
Microtubes Eppendorf, Germany
MBR metal connections Carl Roth, Germany
Flow-through-system flask Carl Roth, Germany
Nitrogen gas 3.0 Westfalen AG, Germany
DNA primers MWG Biotech, Germany
PCR amplification mixture Eurogentec, Belgium
pH-Meter Model pH 27, with glass electrode SE 103, Knick,
Germany
2x RedTaq ready mix Sigma, USA
Scintillation vials Pocto vials (6 mL) and Super Polyethylene vials
(20 mL), Packard Bioscience B.V., The Netherlands
LSC scintillation cocktail Luma SafeTM Plus, Lumac LSC, The Netherlands
Carbomax Plus LSC cocktail, Canberra-Packard
HPLC scintillation cocktail Quicksafe Flow 2; Zinsser Analytic GmbH, Germany
Silica hoses Carl-Roth, Germany
Test kits for analyses Hanna Instruments, Italy
Glass centrifugation tubes VWR International, Germany
Materials and methods 37
4.2. Laboratory-scale membrane bioreactor (MBR) design
The submerged lab-scale MBR is schematically presented in Figure 4.1 and Picture 4.1. The
aeration tank consists of a 1.5 L glass cylinder with a length of 32 cm and a diameter of
11 cm. The cylinder was fixed at both sides into steel lids with 4 screws.
4.2.1. Set-up of the pumps
Peristaltic pumps coordinated by an electronic board controller are used. The influent pump is
automatically controlled by magnetic level controller and started once per hour to maintain
the level at 1 L. The effluent pump was set up for filtration with a flow rate of approximately
4.0 mL/min and with breaks every 10 min for 1 min. The backwashing pump was set-up
automatically for 1 minute pumping every 10 min to clean the membranes (during the break
of the filtration pump).
The aeration tank contained a metal frame for four membrane plate modules, each
sized 9 cm x 3.5 cm. The total surface area of the four membrane modules is 0.024 m2. The
used membranes are made of microfiltration cellulose acetate with a pore size of 0.45 µm.
The membranes were back-washed with effluent flow provided from the permeate collector.
Mixing and aeration of the mixed liquor solid suspension (MLSS) was achieved by using a
porous bubble air diffuser supplying a gas flow of 0.09 m3/h and a magnetic lab stirrer. The
pressure (0.1 – 0.5 bar) was measured by means of a manometer inserted on-line. The MBR
was inoculated with 1 L of activated sludge and the stabilization phase of the process was
initiated. As parts of the radioactive test compound from MBR can be stripped out, an
additional trapping device for possibly volatilized radioactive compounds was connected to
the system for safety reasons and in order to perform an exhaustive balancing of the total
radioactivity. The equipment consisted of a 1 L flow-through-system flask filled with 500 mL
of monoethylene glycol and connected to the gas outlet of the bioreactor (air supply and
biomass respiration) via a pipe. 14C volatile organic compounds (VOC) are absorbed into the
monoethylene glycol phase by trapping the gas flow through this solution. After this step the
gas flow is sparged further through three similar flasks, each filled with 0.5 L of 1M NaOH in
order to trap the 14CO2 released from mineralization processes taking place inside the MBR.
The content of the flasks was replaced before the NaOH solution was saturated with CO2. For
detection of the saturation limit of CO2, phenolphthalein was added to the NaOH solution and
the radioactivity was measured in all three flasks. The outlet of these four recipients was
38 Materials and methods
connected to a vacuum pump working under a slightly reduced pressure (-0.1 to -0.2 bar).
This pump allowed counterbalancing the positive pressure created inside the MBR by the air
supply.
2
1
34
5 6
7
8 9
10
11
12 13 13 13
14
Figure 4.1. Laboratory-scale membrane bioreactor (MBR) system.
Legend: 1 - Bioreactor vessel; 2 - Membrane plate modules; 3 – Influent tank; 4 – Effluent tank; 5 – Influent peristaltic pump; 6 – Air pump; 7 – Level controller; 8 – Manometer; 9 – Controlling valve; 10 – Effluent peristaltic pump; 11 – Backwashing peristaltic pump; 12 – Monoethylene glycole flask for the trapping of volatile organic compounds; 13 – NaOH flasks for the trapping of CO2; 14 – Vacuum pump.
Picture 4.1. Laboratory-scale MBR system.
Materials and methods 39
4.2.2. Operational parameters
The MBR was operated at a sludge retention time (SRT) of 25 days, a hydraulic retention
time (HRT) of 15 hours, and total suspended solids (TSS) have been established at a
concentration of 8 g/L and 10 g/L, respectively.
The chosen values are in agreement with those used at pilot and real scale MBRs.
During the test period the temperature in the system was within 20 to 25°C and the
concentration of dissolved oxygen was maintained at approximately 6 mg/L. Mixed liquor
solid suspension (MLSS) present in the MBR had a clear, dark brown colour due to the
composition of the synthetic feed and revealed a homogenized biomass characteristic of MBR
sludge (Picture 4.2).
)
)
Picture 4.2. Photographs of biomass from the optechniques: a) bright field 100 X; b) Niesser 100Gram staining 1000 X. The MBR was fed with synthetic wastewater
(DEVL41). This synthetic wastewater contains
and urea and the mineral salts K2HPO4, NaCl,
missing nitrification and anoxic conditions, the
MBR proved to be much higher than water q
remove the urea and part of the peptone (rich so
b)
ac
d)tical microscope observed with different 0 X; c) acridin orange staining 400 X; d)
prepared as described in DIN ISO 11733
the organic components peptone, beef extract
CaCl2•2H2O and MgSO4•7H2O. Due to the
amount of nitrogen and phosphorus fed to the
uality standards required. It was decided to
urce in organic nitrogen and phosphorus) and
40 Materials and methods
to replace it with glucose, as an easy degradable carbon source for microorganisms. To
maintain the biomass concentration of 10 g/L under the established operational parameters the
final composition of wastewater contained 700 mg/L chemical oxygen demand (COD), 40
mg/L total nitrogen (TN), and 5 mg/L total phosphorus (TP). The influent was prepared
freshly every five days by autoclavation of the listed ingredients in Table 4.1, and use of
sterile connection tubes to avoid contamination. During the experiment the influent tank was
continuously stirred to prevent sedimentation and to maintain the concentration of nutrients
constant in the whole volume.
Table 4.1. Composition of the used synthetic wastewater.
Component Concentration (mg/L)
Peptone 192
Beef extract 132
Glucose 324
K2HPO4 28
NaCl 7
CaCl2•2 H2O 4
MgSO4•7 H2O 2
During the radioactive test period as well as in the adaptation period, TSS in the activated
sludge, flow rates, pH value and temperature were monitored daily in the MBR and excess
sludge was removed corresponding to sludge retention time. COD, TP, and TN were
measured for each influent charge and effluent collected.
4.3. Analysis of water quality parameters
Total solid suspensions (TSS) in the mixed liquor, flow rate, pH and temperature were
monitored daily in the MBR system. Chemical oxygen demand (COD) and the concentration
of total nitrogen (TN), total phosphorus (TP), nitrate (NO3-) and ammonia (NH4
+) in the
effluent were measured every second day. COD, TP and TN were measured additionally for
each fresh influent charge, which was renewed every 5th day. All the parameters (COD, TP,
TN, NO3-, ammonia) are analysed photometrically by using reactive test kits.
Chemical oxygen demand (COD) was measured photometrically by means of a
tungsten lamp as illumination source equipped with a narrow band interface filter of 420 nm.
Materials and methods 41
The analytical method was adapted from the US EPA 410.4 approved method for the COD
determination in surface waters and wastewaters. Oxidable organic compounds reduce the
dichromate ion to the chromic ion, and the amount of remaining dichromate is determined.
Total phosphorus (TP) analyses is an adaptation of the Standard Methods for the
Examination of Water and Wastewater, 20th edition, 4500-PC, described in
vanadomolybdophoshoric acid method. A persulfate digestion converts organic and
condensed inorganic forms of phosphates to orthophosphate. The reaction between
orthophosphate and the reagents leads to a yellow coloration of the sample, which is
quantified photometrically.
Total nitrogen (TN) and nitrate analyses involve a chromotropic acid method. A
persulfate digestion converts all forms of nitrogen to nitrate, which turns into a yellow tint by
reacting with the reagents of the kit. The reaction product is quantified photometrically.
Ammonia analyses which consist of an adaptation of the ASTM Manual of Water and
Environmental Technology, D1 426-92, described in Nessler method where the reaction
between ammonia and which reagents causes a yellow tint in the sample, measured
photometrically.
4.4. Liquid scintillation counting (LSC)
The radioactive samples were analysed by means of a Beckman LS 5000 TD liquid
scintillation counter. Aliquots of different fractions coming from MBR were added to 5 mL of
Lumasafe scintillation cocktail and liquid scintillation counting was performed. The 14C
atoms of the labelled substance emit ß-particles. These particles activate the molecules of the
aromatic organic solvent contained in the liquid scintillation cocktail. The resulting
fluorescent light allows the quantification of the radioactivity.
4.5. Determination of non-extractable residues
In order to analyze the non-extractable fraction, 0.1 g sample of pre-dried sludge pellets was
packed in cellulose and combusted in a biological oxidizer OX500. Resulting 14CO2 was
trapped in a scintillation vial containing Carbomax Plus LSC cocktail, and was subjected to
LSC. The radioactivity value measured was helpful to recalculate the whole amount of non-
extractable radioactivity present in the sludge sample.
42 Materials and methods
4.6. Radiolabelled compounds
4.6.1. 14C- 4-[1-ethyl-1,3 dimethylpentyl]phenol (NP) preparation
A radiolabelled single isomer of NP, i.e. 4-[1-ethyl-1,3 dimethylpentyl]phenol (isomer
number 111) (5, 27), was used as radioactive model compound. The compound was
synthesized by means of a Friedel–Crafts alkylation using phenol, 3,5-dimethyl-3-heptanol in
the presence of boron trifluoride as catalyst (Vinken et al., 2002). The synthesized product
was characterized by means of high performance liquid chromatography coupled to a diode
array detector and a liquid scintillation counting detector (HPLC-DAD-LSC) and by gas
chromatography-mass spectrometry (GC-MS). Compound characteristics are given in Table
4.2.
Table 4.2. Characterisation of 14C-4-[1-ethyl-1,3 dimethylpentyl]phenol.
Test compound characteristics Chemical structure
Name: 4-[1-ethyl-1,3-dimethylpentyl]phenol
Formula: C15H24O
Molecular weight: 220.36 g/mol
Radioactivity of the stock solution: 14 MBq/mL
Specific radioactivity: 1,492 MBq/mg
Concentration: 9.45 mg NP /mL CH2Cl2
Radiochemical purity: 94.4%
Chemical purity: 94.7%
Applied radioactivity: 4 MBq/L
4.6.2. 14C-17α-ethinylestradiol
The two 14C-17α-ethinylestradiols were differently labelled, one [4-14C] A-ring labelled and
the other [20-14C] ethinyl-labelled in order to gain more information on the possible
degradation pathway (Table 4.3).
Materials and methods 43
Table 4.3. Characteristics of both 14C-EE2 used in the studies.
Test compound
characteristics
[4-14C] (A-ring)
labelled EE2
[20-14C] (ethinyl)
labelled EE2
Chemical structure
Formula
Molecular weight
Specific activity (MBq/mg)
Radioactive purity
Applied radioactivity (MBq)
C20H24O2
296.4
6.377
98%
2.6
C20H24O2
296.4
3.748
98%
2.6
C-20
C-4
4.7. HPLC with radio-detection
Samples were analyzed using HPLC Series 1100 with flow rate of 1 mL/min, equipped with
autosampler, degasser, diode array detector, and online liquid scintillation flow radiodetector
with a cell volume of 1300 µL. The flow rate of the scintillation cocktail was 2.0 mL/min.
50 µL of the sample was injected onto a Nucleosil column.
4.7.1. Analyses of nonylphenol and its degradation products
The compounds were eluted with a gradient of water (A) and acetonitrile (B) (both HPLC-
grade) as follows: 25% B with linear gradient to 90% B during 22 minutes. Finally, after 90%
B isocratic for 5 minutes, the system returned to its initial conditions (25% B) within 10
minutes, and was kept in this composition for 3 minutes before the next run started. The UV
signals were measured at 280 and 294 nm.
4.7.2. Analyses of 17α-ethinylestradiol and its degradation products
The linear gradient program used was: initially 80% A for 15 min, followed by 50% A for 7
min, 5% A for the next 6 min, and finally 80% A until the end of the measurement (39 min)
and kept in this composition for 2 min before the next run was started.
4.8. HPLC-MS/MS analyses
HPLC-MS/MS analyses were carried out with the help of Dr. Sebastian Zühlke at INFU,
Dortmund University.
44 Materials and methods
LC method for NP. Samples were separated on a RP 80A column using a gradient
program as follows: 20% B with linear gradient to 30% B during 3 min and to 100% B during
another 18 min. After 100% B isocratic for 6 min, the system returned to its initial conditions
(20% B) within 0.2 min, and was kept in this composition for 4 min before the next run was
started.
LC method for EE2: Compounds were separated on a Luna C18 column. The mobile
phase consisted of water, 10 mM Ammonia acetate, 0.1% acetic acid (A) and
acetonitrile/methanol (v:v, 9:1) (B). Samples were separated using a gradient program as
follows: 20% B held for 1 min, linear gradient to 50% B during 5 min and to 100% B during
another 6 min. After 100% B isocratic for 7 min, the system returned to its initial conditions
(20% B) within 0.2 min, and was kept in this composition for 4 min before the next run was
started.
After splitting of the solvent flow (flow rate of 800 µL/min) compounds were detected
simultaneously via radioisotope detector with a 3139 quartz cell and a TSQ quantum ultra
AM tandem mass spectrometer. Full scan mass spectra were obtained using the TSQ quantum
equipped with an APCI ion source (Ion Max) operating in negative mode in the m/z range of
92 to 500. Nitrogen was employed as both the drying and nebulizer gas. Product ion spectra
were obtained after fragmentation at collision energies of 15, 30, and 45 V.
4.9. GC–MS analyses
The GC–MS analyses of NP were carried out using an Agilent 6890 gas chromatograph
equipped with a 30 m x 0.32 mm x 0.25 µm phenyl methyl siloxane column. The system was
operated in scan mode (m/z = 50-600) with an ionisation energy of 70 eV. The temperature
programme was 50oC for 5 min, then rising 10oC per min until 280oC and holding 280oC for 5
min. The temperature of the interface was 280oC. The injection volume was 1 µl and the
carrier gas helium was applied at a flow rate of 1 mL/min.
4.10. Preparation of Sphingomonas sp. strain TTNP3 cultures
4.10.1. Sphingomonas sp. strain TTNP3
Sphingomonas is a Gram-negative, non-spore-forming, chemoheterotrophic, strictly aerobic
bacterium that typically produces yellow- or off-white pigmented colonies (Picture 4.3 a and
b). It is unique in its possession of ubiquinone 10 as its major respiratory quinone, and of
glycosphingolipids (GDLs) instead of lipopolysaccharide in the cell envelope. The bacterium
Materials and methods 45
is metabolically versatile, i.e. as carbon source it can utilize a wide range of naturally
occurring compounds as well as some types of environmental contaminants, e.g. NP (Tanghe
et al., 1999).
Picture 4.3. Sphingomonas sp. strain TTNP3. a) Petri dish colonies. b) Microscopic view of single cells. 4.10.2. Glycerol stock
Glycerol stock cell suspension of Sphingomonas sp. strain TTNP3 was prepared with the help
of Boris Kolvenbach in the laboratory of Biology V Institute.
100 mL of preliminarily autoclaved (121°C, 20 min) Standard I medium were
inoculated with 100 µL of an existing glycerol stock and the flasks were incubated at 28°C for
48 hours on a rotary shaker until the turbidity reached a value of 6.4 (measured
photometrically at 660 nm). 25 mL of sterile glycerol were added, and after vigorous shaking
the mixture was aliquoted into 2 mL reaction tubes and stored at -80°C after shock freezing in
liquid nitrogen.
4.10.3. Resting cells cultures
Resting cell cultures of Sphingomonas sp. TTNP3 were done according to Corvini et al.
(2004). 100 mL of sterile Standard I medium were inoculated with 100 µL of a glycerol stock
and incubated at 28°C for 48 hours on a rotary shaker until the turbidity (660 nm) reached a
value of 6.4. The resulting culture was washed twice in 50 mM phosphate buffer with a pH of
7.0 (referred to as phosphate buffer; KH2PO4, K2HPO4) by means of successive centrifugation
at 3000 g during 15 min and resuspension of the pellets in 50 mL of phosphate buffer. Finally,
cells were resuspended in phosphate buffer at a turbidity value of 25 to 30 at 660 nm. The cell
46 Materials and methods
suspension was divided into aliquots of 5 mL and stored at -20°C. Unless stated otherwise,
resting cells suspensions used in this work consisted of frozen cells.
4.11. Polymerase chain reaction–denaturing gradient gel electrophoresis
(PCR-DGGE)
4.11.1. 16S rDNA amplification and sequencing
The amplification and sequencing of Sphingomonas sp. Strain TTNP3 has been carried out by
Dr. Jürgel Prell, University of Reading, UK.
A large fraction of the 16S rDNA sequence was amplified from strain TTNP3
genomic DNA using primer BAC27f (5´-AGAGTTTGATCCTGGCTCAG-3’) and primer
1488r (5´-CGGTTACCTTGTTACGACTTCACC-3´ (Herrera-Cervera et al., 1999). The
reaction was performed using 2x RedTaq Ready Mix, primers at concentrations of 0.5 µM
and approximately 40 ng of template DNA. Cycler conditions were: 1 cycle of 3 min at 94°C.
30 cycles of 45 sec at 94°C, 45 sec at 48°C, 1min at 72°C and 1 final cycle of 5 min at 72°C.
The amplified fragment was sequenced on both strands using the primers BAC27f, 927r (5´-
CCGCTTGTGCGGGCCC-3´ (Amann et al., 1995) and 1488r. A clean sequence of 1398 bp
was obtained and deposited on NCBI under accession number EF514925. A bootstrap tree
was produced using Clustal X1.81 (Thompson et al., 1997) and TreeView 1.6.1 (Page 1996).
4.11.2. PCR-DGGE analyses
DNA was extracted from representative samples of MBR activated sludge taken before,
during and after bioaugmentation as well as from pure cultures of Sphingomonas sp. strain
TTNP3. For DNA extraction, samples were poured into 1.5 mL microtubes together with
equivalent amounts of quartz particles and mixed vigorously with the help of a metal stick
before being frozen in liquid nitrogen. The procedure was repeated five times. The
purification steps were performed using the DNeasy® Plant Mini Kit. After purification, 50
µL of pure DNA for each sample were obtained and tested as 1:10, 1:100 and 1:1000
dilutions on a 1.5% agarose gel. PCR was performed using a Mastercycler personal. For
amplification a Sphingomonas specific set of 16S rRNA gene primers was used (Leys et al.,
2004). The primers were forward primer Sphingo108f (5’-GCGTAACGCGTGGGAATCTG-
3’) and the reverse primer Sphingo420r (5’-TTACAACCCTAAGGCCTTC-3’). A 40-bp GC
clamp (CGCGGGCGGCGCGCGGCGGG CGGGGCGGGGGCGCGGGGGG) was attached
to the 5’-end of the reverse primer to allow DGGE analysis of the amplicons. The temperature
Materials and methods 47
program used with the Sphingo108f/GC40-Sphingo420r primer pair consisted of a short
denaturation of 15 s at 95oC, followed by 50 cycles: denaturation for three seconds at 95oC,
annealing for ten seconds at 62oC and elongation for 30 s at 75oC. The last step included an
extension for two minutes at 74oC (Leys et al., 2004). The PCR mixture contained 100 ng of
extracted DNA, 25 µl of the 2X PCR (containing HotGoldStar DNA polymerase, 20 µM
dNTPs, MgCl2, Tris-HCl, KCl and red dye loading buffer), 25 pmol forward primer, 25 pmol
of reverse primer, and the volume was adjusted to 50 µL using sterile water. The PCR
products were checked on a 1.5% agarose gel and used further for DGGE on a
polyacrylamide gel. A 6% polyacrylamide gel with a denaturation gradient of 60% to 47%
(100% denaturant gel contained 7 M urea and 40% formamide) was used for DGGE.
Electrophoresis was performed at a constant voltage of 50 V for 16 h in 1x TAE (Tris-acetate-
EDTA) running buffer at 60oC in a universal mutation detection system. After electrophoresis
the gel was stained in a silver bath and scanned for further analysis.
4.12. Surface tension measurements
The surface tension coefficient was measured by means of ring method (Du Nouy method)
using a tensiometer. The platinum ring hanging from the balance hook is first immersed into
the liquid and carefully pulled up away from the surface of the liquid, while the microbalance
is recording the force applied. Samples containing different proportion of cells of
Sphingomonas sp. strain TTNP3 and activated sludge were shaken and centrifuged before
determining the surface tension coefficient in the resulting supernatant. The device was
calibrated with distilled water and measurements were performed in triplicates at room
temperature.
4.13. Sampling and preparation of the samples
COD, TN, TP, ammonia (NH4+) and nitrate (NO3
-) in the effluent were measured
photometrically every two days using reactive test kits. Immediately as the radiotracer was
spiked to MBR and homogenized, samples were taken from MLSS, VOC trap and CO2 trap
for daily LSC measurements. Three samples of 0.5 mL (for sludge and effluent) and 1 mL
volume (from NaOH and monoethylene glycol) respectively, were mixed with scintillation
cocktail for LSC measurement. Samples from the collected effluent and from the activated
sludge (divided in water phase and pellets) were extracted daily, in order to avoid further
degradation during the storage. 40 mL of MLSS freshly collected from MBR were transferred
48 Materials and methods
into glass centrifugation tubes and centrifuged (10 min, 3,000 g). Then, the phases were
separated in water phase and pellet and the radioactivity measured by means of LSC.
Afterwards, the two phases were extracted three times with ethyl acetate and the resulted
organic phase was dried over Na2SO4 and filtrated. The filtrate was concentrated by means of
a rotary evaporator (water bath at 45oC) to 1 mL, transferred into a HPLC vial and dried under
a N2 stream at room temperature. The residue was dissolved in HPLC grade methanol to a
total volume of 0.5 mL and the sample was ready for further analyses. Measurements of the
radioactivity in the extracts before drying and after resuspension did not evidence losses of
radioactivity. After the extraction step, part of the pellet was prepared for non-extractable
fraction analyses. At the end of the study, the MBR was stopped, sludge was removed and
vessel, hoses, connection parts, porous stone air diffuser and the membrane modules were
washed several times with water and then submerged in ethyl acetate for 5 days until no
radioactivity was released. The quantity of radioactivity in the different fractions of washing
water and solvent were determined by means of LSC.
4.14. Preliminary tests
4.14.1. Nonylphenol and ethinylestradiol studies
A preliminary radioactive test was performed in order to test an extraction method for NP and
EE2 and their respective degradation products. 40 mL excess of sludge was spiked with 14C-
NP or 14C-EE2 respectively each at 100 µg/L, shaken for 1 h and then centrifuged. The
radioactivity was measured by means of LSC after the separation of the phases in pellet and
water phase followed by the extraction and concentration steps as described above (chapter
4.13).
4.14.2. Bioaugmentation study
Suspensions of resting cells of strain TTNP3 were prepared as described in Chapter 4.10. For
an easy-to-apply solution allowing a longer period of storage, the cells were lyophilized for
24 h at -40oC. Then, the activity of fresh cells and lyophilized cells was investigated in order
to gain information on the degradation capacity of pure bacteria and to quantify losses of
activity caused by the freeze-drying step. Furthermore, preliminary tests on the degradation
capacitiy of Sphingomonas sp. strain TTNP3 in mixtures with activated sludge were carried
out by adding freeze-dried cells (resuspended in distilled water) to MLSS at ratios varying
between 0.5% and pure bacterial suspension. 14C-NP was applied to vials with final
Materials and methods 49
concentrations of 100 µM and placed under a gentle nitrogen stream to evaporate the solvent.
1 mL of the freeze-dry resting cells and of the Sphingomonas-MLSS mixtures, respectively
was added to each vial and incubated for 1 h in a water bath at 35oC under continuous stirring.
After incubation, the reaction was stopped with 1 mL of ice-cold methanol. Samples were
centrifuged for 10 min at 15,000 g and the supernatant analyzed by means of HPLC.
4.15. Experimental procedure
4.15.1. Nonylphenol (NP) fate study
The efficiency of the MBR in treating synthetic wastewater was monitored over a period of
120 days. The synthetic composition was adapted from DIN ISO 11733 (DEVL41)
autoclaved at 120°C during 20 min, and applied into the MBR by means of a peristaltic pump.
After this period, radiolabelled NP was applied as single pulse to the MBR at a concentration
of 4 MBq/L corresponding to 2.3 mg 14C-4-[1-ethyl-1,3- dimethylpentyl]phenol/L. The
monitoring was performed daily from all MBR compartments. Aliquots from MLSS, effluent,
volatilization and mineralization flask were measured in LSC. Effluent and MLSS were
extracted daily in order to prevent possible further degradation during the sample storage. The
monitoring and analyses lasted over 34 days, until the radioactivity in the system was
exhausted.
4.15.2. 17α-ethinylestradiol (EE2) fate study
During the adaptation period of 29 days, the inoculated MBR was continuously supplied with
EE2 via influent containing 100 µg of non-radiolabelled EE2/L in order to adapt the bacteria
(to this compound) and to saturate possible adsorption sites. The MBR influent was
simulating the industrial wastewater and was prepared with the components presented in
Table 4.4. The organic components were prepared separately by adding the necessary
amounts to 500 mL of sterile distilled water to prevent evaporation. The inorganic
components were prepared in the glass influent tank and sterilized. After cooling down the
separately prepared organic component were added to the media together with 100 µg/L EE2.
Afterwards the solution was shaken and ready for analyses of the parameters.
In order to characterize and identify degradation products released in the MBR
effluent, an EE2 concentration higher than environmental concentrations was used. The EE2
solution was freshly prepared in pure ethanol and added to each new influent charge supplied
to the MBR. After a stabilisation period of 29 days, the radioactive test period was started by
50 Materials and methods
replacing the non-radiolabelled EE2 in the influent with a solution containing 14C-labelled
EE2, to which non-radiolabelled EE2 was added to maintain the total concentration of EE2 at
100 µg/L influent. Two different 14C-labelled forms of EE2, i.e. [20-14C] (ethinyl) and [4-14C]
(A-ring) labelled EE2 were used. Experiments for each radio-marker were carried out
separately and a stabilization period of the MBR process was operated for each run. For each
form of the radiolabelled EE2, the radioactivity of the influent was adjusted to 0.32 MBq/L
synthetic wastewater. The monitoring of the radioactivity in the system was carried out over a
period of five consecutive days.
Table 4.4. Composition of the synthetic wastewater.
Component Concentration (mg/L)
1. Organic solvents Methanol 40 Ethanol 60 Acetone 50 Propanol 35
2. Inorganic growth medium Na2HPO4•7 H2O 16.7 KH2PO4 4.25 NH4Cl 0.85 MgSO4•7 H2O 4.5 CaCl2•2 H2O 7.28 FeCl3•6 H2O 0.0625
3. Trace elements CoCl2•2 H2O 0.02 MnCl2•2 H2O 0.003 ZnSO4•7 H2O 0.01 H3PO 0.03 Na2MoO4•2 H2O 0.003 NiCl2•6 H2O 0.002 CuCl2•2 H2O 0.001 NaCl 6600
4.15.3. Bioaugmentation procedure
The lab-scale MBR was inoculated with activated sludge from the pilot MBR of wastewater
treatment plant Soers, Aachen and fed with synthetic wastewater (Table 4.1). It was
acclimatized for 63 days. During this period the effluent parameters COD, TN, TP and TSS
were regularly measured. At the end of the stabilization period (day 64), 3.8 MBq of 14C-NP
were spiked at a concentration of 2.5 mg/L. After 10 min, the MBR was inoculated with
100mg/L lyophilized resting cells of Sphingomonas sp. strain TTNP3 (resuspended in
Materials and methods 51
distilled water). Similar inoculations of the system were repeated after 24, 48, 72, and 96
hours. At defined time intervals (e.g. before addition of Sphingomonas and three hours after
inoculation, which was a sufficient time to observe the effects of freshly added bacteria to the
system), samples of MLSS (0.5 mL), effluent (1 mL) and solutions for the trapping of
volatiles (0.5 mL) and CO2 (0.5 mL) were sampled in triplicate and analyzed by means of
liquid scintillation counter (LSC). In order to gain information on the impact of
bioaugmentation on the microbial community present in the MBR, the system was kept
running for further two weeks after the radioactivity reached the limit of detection in the
effluent. The control MBR was operated under the same conditions without addition of the
bacterium
53
5. Results
5.1. Performances of MBR: optimization and operational parameters
5.1.1. General prerequisites
The lab-scale MBR developed and used for this study had to fulfil two main criteria. First of
all, the quality standards for surface waters set in Germany were taken into account in order to
evaluate the reliability of the MBR treatment developed at laboratory scale. Even at small
scale, it was intended to achieve the performances on wastewater treatment at the level of a
real scale activated sludge treatment that later on the results obtained with this system can be
in the further studies used for up scaling. A second criterion taken into account for the
development of the MBR was the work under radioactive conditions. That imposes the
restrictions on the amount of waste production, exhausted gas from the system, tightness of
the system and handling of the samples. The MBR system was dimensioned for the
production of less than 1.6 L effluent/day to avoid additional problems with radioactive waste
discharge and laboratory safety. Reported data on the MBR efficiency were collected over a
period of 120 days to prove the stability and robustness of the investigated system.
5.1.2. Synthetic medium formulation
At the beginning of the study, the synthetic wastewater influent used was an adaptation from
DIN ISO 11733 (DEVL41) and the C:N:P ratio of the nutrients was 100:20:2, equivalent to
700 mg/L COD, 80 mg/L TN and 10 g/L TP. It is usually stated that the ratio of C:N:P in
wastewaters to be treated should be approximately 100:5:1 for aerobic treatment and 250:5:1
for anaerobic treatment (Tchobanoglous et al., 2004).
The designed MBR was provided just with an aeration tank, the usual anaerobic-
anoxic tanks of a municipal wastewater treatment were missing. Therefore, the nitrogen and
phosphorus recovered in the effluent of the MBR’s highly exceeded the allowed
concentrations in surface waters. The only satisfactory parameter was COD with high
biodegradation rates of about 90% (Figure 5.1.1.).
54 Results
0
10
20
30
40
50
60
70
80
0 20 40 60 80 100 120Time (days)
CO
D, m
g/L
COD monitoring
Influent composition changed
Figure 5.1.1. COD concentration in the permeate during the adaptation period.
The unconsumed nitrogen and phosphorus source prescribed in the initial receipt have been
removed by trial. The concentrations of the two nutrients have been measured in solutions
containing the organic ingredients of synthetic wastewater (e.g., peptone, beef extract, urea).
Analysed contents of N and P of beef extract and peptone taken together were already two
times higher than the above mentioned concentrations required for the growth of activated
sludge. After these results it was decided to take out the additional nitrogen source (i.e. urea)
from the wastewater composition. Furthermore, half of the peptone and beef extract amounts
were replaced by glucose as carbon source, in order to maintain the biodegradable COD at
700 mg/L necessarily for growth of the 10 g/L biomass present in the system. Phosphorus was
supplemented by the addition of inorganic source (K2HPO4). The final composition applied to
the MBR was 700 mg/L COD, 40 mg/L TN and approximately 5 mg/L TP. After a short
equilibration period, the recorded effluent parameters respected the frame of surface water
quality (Water Quality, Surface Water Standard TCVN 5942/1997) with concentrations of
phosphorus below 2 mg/L (Figure 5.1.2) and of total nitrogen not exceeding 18 mg/L (Figure
5.1.3). No supplementary treatment step was necessary to improve the quality of the effluent.
A slight acidity (pH = 5-6) in the MBR vessel and in the resulting permeate was
observed. In literature, it was reported that low pH value result from nitrification processes
and occur when the concentration of buffering substances present in the system is too low
(Tchobanoglous et al., 2004). In general it is admitted, that there should be at least 50 mg/L of
Results 55
alkaline substances (e.g. Na, K, Ca) remaining in the effluent to prevent problems with low
pH. The performance of MBR fed with synthetic wastewater during the adaptation period is
summarized in Table 5.1.1.
0
10
20
30
40
50
60
70
80
0 20 40 60 80 100 120
Time (days)
TN
, mg/
L
TN monitoring
Influent composition changed
Figure 5.1.2. Total nitrogen concentration in the permeate.
0
2
4
6
8
10
12
14
16
18
0 20 40 60 80 100 12Time (days)
TP,
mg/
L
TP monitoring
Influent composition changed
0
Figure 5.1.3. Total phosphorus concentration in the permeate.
56 Results
Table 5.1.1. Summary of the MBR’s performances.
Parameter Wastewater
influent
MBR effluent Removal efficiency
in the MBR, %
COD, mg/L 700 ± 20 20 ± 5 > 97.0
Total nitrogen (TN), mg/L 40 ± 5 10 ± 2.5 > 75.0
Nitrate (NO3-), mg/L < 0.5 8 ± 2 -
Total phosphorus (TP), mg/L 5 ± 0.5 1.5 ± 0.5 > 70.0
5.2. Fate of 14C-4-[1-ethyl-1,3-dimethylpentyl]phenol in the lab-scale MBR
5.2.1. Radioactivity monitoring
After the MBR was adapted to synthetic wastewater under laboratory conditions,
radiolabelled NP was applied as single pulse to the MBR at a concentration of 4 MBq/L
corresponding to 2.3 mg/L 14C-4-[1-ethyl-1,3-dimethylpentyl]phenol. The aim of using 14C-
label was to distinguish between biodegradation, sorption, mineralization and volatilization of
NP in the MBR.
After spiking NP, the radioactivity was monitored by means of LSC in all relevant
compartments of the system, i.e. MLSS, effluent, CO2 and VOC-trap in order to get
information on the fate of the compound in MBR. During the first day, the radioactivity inside
the MBR (radioactivity recovered in the MLSS fraction) dropped markedly by around 60% of
the initial applied amount. The daily monitoring showed a continuous decrease of
radioactivity in MLSS, until only 1.8% of the initial applied amount was measured at day 34
(Figure 5.2.1). This was corroborated with the low amount of radioactivity found in the
mineralization fraction (< 1%), which increased slowly during the first half of the experiment
and remained almost constant in the last days (Figure 5.2.2).
In the VOC trap containing monoethylene glycol, a very low level of radioactivity (<
1%) was observed. The radioactivity of the VOC fraction remained nearly constant after the
first week (Figure 5.2.2).
Results 57
0
10
20
30
40
50
60
70
80
90
100
0 5 10 15 20 25 30 35
% o
f ini
tially
app
lied
radi
oact
ivity
MLSS monitoring
Time (days)
Figure 5.2.1. Radioactivity monitoring of MLSS.
Figure 5.2.2. Radioactivity monitoring in NaOH and monoethylene glycol solutions for trapping 14C-CO2 and 14C-VOC, respectively.
In the effluent the first traces of radioactivity were detected in the samples analysed two hours
after application of the 14C-NP and the concentrations measured after 2 days were up to 70
Bq/mL. During the second period of the study the concentration of radioactivity decreased
58 Results
until reaching the detection limit of the LSC measurement method (i.e. approximately 1 Bq)
(Figure 5.2.3).
Figure 5.2.3. Time course of daily radioactivity measured in the MBR effluent. Percentages refer to the initially applied radioactivity.
5.2.2. Sludge analyses
In order to gain information on the fate of NP in sludge, daily sludge samples were
fractionated by means of centrifugation and the radioactivity was measured in the supernatant
and in the MLSS pellet. Further more, the MLSS pellet fraction was extracted 3 times with
ethyl acetate. The extractable 14C fraction was quantified by means of LSC measurement. The
remaining solid fraction was combusted in order to determine the bound 14C residues of NP in
the sludge pellet. These residues which were not extractable with the applied method of
extraction, account for strong association mechanisms like e.g. covalent binding, ionic
interaction of the parent compound or its metabolites. The recovered radioactivity in the
organic extracts (pellets and water phase) and in the non-extractable fraction is presented in
Figure 5.2.4.
Results 59
0
10
20
30
40
50
60
70
80
90 R
adio
activ
ity [%
]
pellet extractionwater phase from MLSS non-extractable compounds associated to MLSS pellet
day 1 day 2 day 3 day 6 day 9 day 12 day 20 day 26 day 33
Figure 5.2.4. Distribution of radioactivity in MLSS pellets and supernatant fractions of activated sludge. Percentages of radioactivity in each fraction are expressed relatively to the radioactivity measured in the samples before fractioning (100%).
In the samples from the first couple of days, the radioactivity recovered in the water phase of
MLSS accounted for less than 5% of the total sample and increased with the time slowly to
8%. This was partially in agreement with the results of preliminary experiments concerning
the development of the extraction method. In these tests with 14C-NP more than 90% of the
radioactivity measured was associated to the sludge and only up to 10% remained in the
supernatant after 2 h of incubation. During these experiments the sorbed fraction was easily
extractable shortly after the spiking of NP and thus was considered as weakly sorbed to the
sludge matrix. The biodegradation of NP and the formation of hydrosoluble compounds was
assumed not to take place in such short period. In the case of the MBR process, more than
70% of the radioactivity contained in the MLSS pellets could be extracted after the first day,
while it decreased to 10% by the end of the experiment. In contrary, the fraction of
radioactive bound residues of NP increased proportionally. The examination of the organic
pellet extracts by means of HPLC-LSC showed that the recovered radioactivity was assigned
to NP itself.
60 Results
5.2.3. Effluent analyses
The LSC measurements of the effluent samples showed that part of the radioactivity applied
to the MBR was leaving the system. HPLC equipped with a radio detector was employed in
order to analyze the organic extracts of the effluent samples. The respective chromatograms
show a whole range of radioactive signals for the measured samples, starting from the first
days of the incubation (Figure 5.2.5). The radioactive peaks accounted for compounds which
were more polar than the parent compound and elute before NP on a reversed phase column
(retention time 21 min). The peaks present in the HPLC chromatograms of the different
incubation periods are characterized by reproducible retention times but are not constant over
the time course of the testing period (Table 5.2.1).
The NP standard solution was analyzed using the same method to ensure that it did not
contain traces of degradation products formed during the MBR treatment of wastewater. The
metabolites of NP have a smaller molecular size than the membrane (0.45 µm) and present
increased polarity. Thus, they were able to cross the membrane barrier and could be recovered
in the effluent. This abundance decreased to the end of the incubation period as at this time
almost no further NP is available in the water phase of MLSS (seeTable 5.2.1).
Fm
igure 5.2.5. HPLC chromatograms of the effluent extracts. Peak 4 contained two etabolites, which were not separated under the applied chromatographic conditions.
1 6 11 16 21 26 31 36
Retention time in HPLC (min)
Rad
ioac
tivity
day 2
day 30
NP
Peak 3
Peak 4
Peak 2
Peak 1
Results 61
Table 5.2.1. Concentration profiles of the metabolites during the time course of the incubation.
Day Radioactivity
(Bq/L)
Peak 1
(%)
Peak 3
(%)
Peak 4
(%)*
NP Peak
(%)
2 54681 36 16 30 12
6 27745 32 21 27 18
15 17342 35 13 35 3
17 12355 41 16 27 2
20 7896 33 18 29 -
23 5888 24 11 26 -
26 4370 15 11 33 3.5
30 3348 9 19 46 5
32 2933 6 12 58 7
34 2340 7 9 50 10 *Peak 4 contains two unresolved compounds
HPLC/MS/MS analyses
After the organic extracts of the effluent samples were analysed in the HPLC with radio
detection, the further interest turned to get information on the moleculare structure of the
degradation products of NP. The peaks were separated according to the fragmentation pattern,
collected as fractions, concentrated and tried to be identified by coupling of chromatographic
and mass spectrometric techniques (Zühlke et al., 2004). Three of the metabolites were
tentatively identified by means of HPLC tandem mass spectrometry and radiodetection. The
proposed structures are presented in Table 5.2.2 and the ion spectra of NP and the 3
metabolites in LC-MS/MS are given in Figure 5.2.6.
62 Results
Table 5.2.2. Proposed structures of the metabolites of 4-[1-ethyl-1,3-dimethylpentyl]phenol according to HPLC/MS/MS analyses.
Peak no. (see Figure 5.2.5), retention time (min)
Proposed structure of the metabolite MW (g/mol)
Precursor ion [M-H]- and product ions (m/z)
Peak 4, compound 1
RT: 13.33
HO
O
3-(4-hydroxyphenyl)-3,5-dimethylheptan-4-one
234
233,
149, 147, 133, 117, 93
Peak 4, compound 2
RT: 13.80
HO
O
5-(4-hydroxyphenyl)-3,5-dimethylheptan-2-one
234
233,
191 (traces), 149 (traces),
147 (traces), 133, 117, 93
Peak 3, compound 3
RT: 11.53
HO
O
4-(4-hydroxyphenyl)-4-methylhexan-3-one
206
205,
147, 133, 117, 93
Results 63
50 100 150 200 m/z
0
10
20
30
40
50
60
70
80
90
100
Rel
. Abu
ndan
ce
233
147
133
LC-MS/MS Peak at RT 13.33 min
: :
50 100 150 200 m/z
0
10
20
30
40
50
60
70
80
90
100
Rel
. Abu
ndan
ce
233
133
147
LC-MS/MS Peak at RT 13.80 min
205
50 100 150 200
m/z
0
10
20
30
40
50
60
70
80
90
100
Rel
. Ab
unda
nce
147
133
LC-MS/MS Peak at RT 11.53 min
50 100 150 200 m/z
0
10
20
30
40
50
60
70
80
90
100
Rel
. Abu
ndan
ce
219
133
147
NP- Peak at RT 18.61 min
Figure 5.2.6. Product ion spectra of NP and the 3 metabolites at collision energy of 20 V; precursor ions at m/z 233, 219 and 205 (RT = retention time).
Interpretation of the LC-MS/MS spectra
The molecular weights of the NP metabolites in the effluent were determined by measuring
[M-H]- in the full scan mode in the LC-MS. These even-electron ions are efficiently produced
in the ion source and contain less energy. Collision-induced dissociation leads to fragment
ions that can help to identify the structure of unknown metabolites. Characteristic ions of the
64 Results
product ion spectra (LC-MS/MS) were 93, 117, 133, and 147 (m/z). The fragmentation of
even-electron anions include homolytic bond cleavage and radical loss neighboring to hetero
atoms, e.g. keto groups (de Hoffmann & Stroobant 2002). Cleavage between the quaternary C
atom and the rest of the alkyl-chain would lead to relatively stable ions of m/z 147 if this
formation is assisted by a heteroatom at C4. Thus, abundant fragments at m/z 147 were
indicative of a keto substitution at C4 of the alkyl chain of 3-(4-hydroxyphenyl)-3,5-
dimethylheptan-4-one and 4-(4-hydroxyphenyl)-4-methylhexan-3-one (see Figure 4.2.4). The
fragments of m/z 133, 117 and 93 were obtained after cleavage of CHx-positions from alkyl
phenols at higher collision energies. All these fragments (133, 117, 93) were also observed for
the original NP. The proposed 5-(4-hydroxyphenyl)-3,5-dimethylheptan-2-one at RT 13.80
indicates a product ion of m/z 191. This can additionally confirm the proposed structure, as
the loss of m/z 42 may have been due to the loss of the acetyl group.
5.2.4. Radioactivity balance
After the radioactivity reached the low detection limit in the MBR and the experiment was
finished, a balance based on the radioactivity was performed (Figure 5.2.7). During the
experiment, part of the radioactivity was taken out daily with the excess sludge (40 mL). The
excess sludge contributed significantly to the elimination of 14C-labelled material from MBR,
amounting to 21.3% of the applied radioactivity over the whole process (34 days). After
stopping the bioreactor, the component parts of the system (i.e. silicon hoses, connections,
membranes, air porous stone, metal and steel components) where first washed with water and
after submerged in ethyl acetate for 3 days. The total radioactivity in the desorbed fraction
was 33.7%. The total radioactivity measured in the collected effluent (1.6 L/day) over the 34
days amounted to 42% from initially applied 14C. At the end of the study, 1.8% of the
radioactivity was still sorbed to the MLSS present in the MBR. The 14C belonging to the
volatilization and mineralization fraction together amounted to less than 1% from the initially
applied radioactivity. At the end of the study was recovered 99% of the applied radioactivity
in the system.
Results 65
MBR1.83%
excess sludge21.3%
effluent 42.2%
adsorption33.7%
CO2
VOC0.16%
0.65%
Radioactive balance in MBR after 34 days
Figure 5.2.7. Radioactivity balance at the end of the 34 days incubation period of MBR operation with one pulse spiking of radioactive NP (2.3 mg/L).
5.3. Behaviour of two differently radiolabelled 17α-ethinylestradiols
continuously applied to MBR with adapted industrial activated sludge
5.3.1. Radioactivity monitoring
The aim of this study was to investigate the capacity of an industrial activated sludge taken
from a MBR of a pharmaceutical factory to degrade EE2. In order to adapt the biomass to
EE2, over a period of 29 days the system received continuously 100 µg EE2/L influent before
the radioactive spiking started. Assuming that the industrial activated sludge was already
adapted to EE2 present in the industrial wastewater of the factory it was decided to use the
continuous application of this compound in the present study to maintain the conditions
similar to the real system. Due to the exposure conditions, it was assumed that the application
of radio-marked EE2 for the last 5 days of the experiment should be sufficient to get
information on the fate of EE2 in MBR and to identify degradation products formed. In the
same time, we were restricted to 5 days of radioactive experiment due to the costs and amount
of available radioactive material.
66 Results
First, quantification of 14C in the MBR (MLSS) and in the effluent was performed by
means of LSC during MBR treatment in order to gain a general overview of the fate of EE2.
The mean values of the radioactivity measured in the activated sludge and in the effluent for
both studies (14C-ring and ethinyl-labelled EE2) are shown in Figure 5.3.1 and Figure 5.3.2.
Until now, no degradation products of EE2 are reported in the environment. The aim
of this study was to get a deeper insight into the degradation pathway of EE2. The
investigations were carried by using two differently labelled forms of EE2, A-ring and
ethinyl- group, respectively. We expected that the idea to label different positions of the EE2
molecule will lead to the detection of radioactive peaks helpful for the identification of
metabolites and to understand the preferential degradation pathway in an industrial
wastewater treatment system.
0
100000
200000
300000
400000
500000
600000
700000
800000
0 1 2 3 4 5
Time (days)
A
MLSS monitoring
pplie
d ra
dioa
ctiv
ity, B
q/L
6
Figure 5.3.1. Monitoring of the radioactivity measured in MLSS samples of both MBR during the period of continous addition of 14C-labelled EE2. Error bars represent the minimum and maximum values of two replicates.
Results 67
0
20000
40000
60000
80000
100000
120000
140000
160000
0 1 2 3 4 5 6
Time (days)
App
lied
radi
oact
ivity
, Bq/
L effluent monitoring
Figure 5.3.2. Monitoring of the radioactivity measured in the effluent samples of both MBR during the period of continous addition of 14C-labelled EE2. Error bars represent the minimum and maximum values
Within the first hours of continous application of 14C-EE2 into the MBR no radioactivity
could be detected in the permeate, while the LSC measurements showed a rapid accumulation
of radioactivity in the activated sludge (MLSS). In the case EE2 would not sorb to the reactor
and the MLSS, the experimental conditions used would have allowed its detection in the
effluent within the first ten minutes if one takes into account the detection limit of 0.5 Bq, the
amount of radioactivity in the influent, the reactor volume and the feed flow-rate. The
concentration of 14C in the effluent remaining below the detection limit for several hours after
the application of 14C-EE2 with the influent was started, suggested that EE2 was entirely
retained in the reactor. The low amount of radioactivity in the permeate was in agreement
with the values measured in MLSS which increased continuously up to the third day. After
three days, the radioactivity in the permeate increased as the concentrations of 14C in the
MLSS tended to stabilize. At the end of the operation of the reactors, the radioactivity
contained in one litre of MLSS was twice that contained in one litre of the incoming feed-
solution (320,000 Bq/L). At the same time, approximately one third of the applied
radioactivity was released from the system with the effluent.
These results demonstrated the high tendency of EE2 residues to accumulate on
suspended solids, although the reactors were fed with non-radiolabelled EE2 during the
68 Results
stabilization phase. The radioactivity measured in the NaOH and in the monoetylene glycol
traps was at the lower limit of detection of the LSC after the five days incubation period.
5.3.2. Sludge analyses
In order to gain information on the chemical species remaining sorbed to the solid fractions of
MLSS, daily samples were worked out by separation of water phase and pellets. These
fractions were treated by a liquid-liquid extraction with ethyl acetate for 3 times. After the
extraction, aliquots were injected in a HPLC system with radio detection. The subsequent
HPLC analysis confirmed that the fraction of radioactivity sorbed to the sludge pellets, which
remained extractable, consisted exclusively of unmodified EE2.
The radioactivity measured in the water phase was constant around 15%, while the
fraction extractable from the pellets constantly amounted to approximatey 70%. The
radioactivity corresponding to the non-extractable fraction bound to the sludge solids was
quantified after combustion of the samples. Bound residues constituted 12 to 15% of the total
radioactivity of the samples. As shown in Figure 5.3.3, the partitioning between extractable
radioactivity and bound residues of EE2 remained almost constant. In fact, the binding to the
sludge was observed from the first application day of 14C-EE2, despite the fact that EE2 was
continuously applied to the sludge over 29 days and possible also before in the industrial
wastewater. The radioactivity recovered from the sludge was analysed by means of HPLC-
LSC and could be completely identified as unmodified EE2.
Results 69
0
10
20
30
40
50
60
70
80R
adio
activ
ity [%
]
water phase from MLSS pellet extraction non-extractable compounds associated to MLSS
day 1 day 2 day 3 day 4 day 5
Figure 5.3.3. Partitioning of the radioactivity between liquid phase and solid phase (extractable and non-extractable residues) of MLSS during the continuous application of radiolabelled EE2 to the MBR. Error bars represent the minimum and maximum of radioactivity values (expressed in % relatively to the radioactivity measured in the samples before fractioning) for samples from MBR separately fed with 14C-ethinyl-labelled EE2 and 14C-A ring–labelled EE2.
5.3.3. Effluent analyses
In order to gain information on the radioactive compounds released with the permeate,
chemical analyses of ethyl acetate extracts from 24 hours samples were performed by means
of HPLC coupled to a radiodetector (Figure 5.3.4).
First remarkable observation was the similarity of the patterns of the two differently
labelled 14C-EE2 molecules studied. A total of five different peaks could be distinguished; the
largest one (peaks no. 3) eluting at 17 minutes was EE2. The respective peaks in each sample
had similar retention times, but the intensity of the peaks of the various compounds increased
correspondingly to the increased load of radioactivity in the reactor. In comparison to an EE2
standard, peak no. 3 corresponding to EE2 eluted slightly earlier in all effluent extracts. A
verification carried out by means of LC-MS/MS coupled to radiodetection confirmed that this
peak effectively corresponded to EE2. Molecular ion as well as product ion spectrum was
identical. This was confirmed by performing GC-MS analyses of the collected peak eluting at
17 min and comparing its mass spectrum to that of the EE2 reference standard. It was
70 Results
assumed that the slight shift in retention time was due to the complex matrix contained in the
samples, in comparison to the standard sample dissolved in pure solvent. The analysis of
concentrated standards in deeper details showed that in fact the compounds corresponding to
the peaks no. 2, 4, and 5 consisted of impurities originally contained in both radiochemicals.
Obviously, only the compound eluting in peak no. 1 (RT 7.5 min) was a metabolite resulting
from a transformation process catalyzed by some microorganisms present in the industrial
activated sludge used to inoculate the MBR.
Peak no. 1 accounted for approximately 5% of the total radioactivity contained in the
effluent extract. Since this metabolite was detected in both studies it was assumed that the
acquisition of further structural information on this compound could support the
comprehension of the biological processes occurring in these bioreactors.
3
Figure 5.3.4. Stacked radio-chromatograms of organic extracts of filtrated effluent from various sampling dates for the reactor fed with 14C-ethinyl-labelled EE2. Insert shows the pattern obtained in the study with 14C-A ring-labelled EE2. LC-MS/MS analyses
Metabolites of organic trace compounds in water samples might be identified by the coupling
of chromatographic and mass spectrometric techniques (Zühlke et al., 2004). Information
about the structure of metabolites can be achieved via decay and analysis of the fragmentation
pattern. An almost complete guarantee for correct structure identification can only provide the
Results 71
synthesis of the proposed metabolite and the comparison of retention time and MS
fragmentation pattern (Zühlke et al., 2004).
Metabolites can also be tentatively identified by means of HPLC tandem mass
spectrometry coupled to radiodetection. Peak no. 1 (radio-chromatograms Figure 5.3.4) was
separated and fraction analyzed via LC-MS/MS-radiodetection. Unfortunately, the
concentration of this compound was too low to give unambiguous results in the radiodetector.
Therefore, the complete extract of the effluent samples was analysed via LC-MS/MS coupled
to a radiodetector. Under conditions of LC method 1, the signal of EE2 (Rt 11.45 min)
accounts for most of the radioactivity (about 95%) and only three further minor compounds
could be detected at Rt 10.07, 12.01 and 16.20 min. The identification of the molecular ion of
these three peaks was difficult due to the concentration level and the low detector response in
all tested ionization modes (APCI as well as ESI in positive and negative mode). The results
of negative APCI confirmed the main peak as EE2 also with identical product ion spectra of
the compound in the sample and the reference. One compound was assumed to have a [M-H]-
of 365.1 and another with 311.1. The compound with the proposed molecular weight of 312
might be the hydroxylated derivative form of EE2. This could be checked with product ions
typical for estrogenic steroids (143, 183). The product ions of the peak at m/z 365.1 could not
be correlated to fragments of estrogens.
Therefore, another LC separation under different conditions was performed (see
description of LC-method 2). The compound with the molecular weight 366 could now be
excluded as the signal from radiodetector and MS differed in their retention times. The signal
for the postulated hydroxylated EE2 was still confirmed. Analyzing of the radiolabelled EE2
standard resulted in small concentrations of the proposed metabolite in the substrate. Due to
the small concentrations of the compound in the extracts used for the identification of the
compound, it remained unclear whether the hydroxylated EE2 detected in the effluent was
spiked as impurity with the EE2 into the MBR or whether EE2 was transformed during
sewage treatment.
72 Results
5.3.4. Radioactivity balance
The radioactivity balances obtained from the experiments with two differently 14C-labelled
EE2 molecules are presented separately in Figure 5.3.5. At the end of the studies, between 95
and 97% of the total radioactivity applied was recovered.
Over the whole test period, the amount of volatilized EE2 and of the mineralized
fraction (14CO2) were negligible and accounted for less than 1% of the total radioactivity
applied independently of the labelling of EE2. The total radioactivity recovered in fractions
sorbed to the MLSS was 31-37% and the fraction sorbed to component parts of the MBR
amounted to approximately 30-37% of the applied radioactivity. A significant portion of the
applied radioactivity was detected in the fouling cake of the membranes (5-6%). The daily
withdrawal of excess sludge contributed to the removal of 4-5% of the applied radioactivity
over the test period, while a high amount of radioactivity was released within the permeate
(16-19%).
0
20
40
60
80
100
120
Rad
ioac
tivity
, (%
)
MBR run with 14C-ethinyl-labelled EE2
MBR run with 14C-A ring-labelled EE2
<1 <1Mineralization Volatilization Effluent MLSS Fouling
cakeAdsorption Excess
sludgeTotal recovery
1619
3731
6 4.6
3037
3.9 4.7
95 97
Figure 5.3.5. Radioactivity balance after 5 days of MBR operation under continuous application of radioactive EE2.
Results 73
5.4. Bioaugmentation of MBR with Sphingomonas sp. strain TTNP3 for the
degradation of nonylphenol
5.4.1. Pre-studies on the degradation of nonylphenol by Sphingomonas sp. strain TTNP3
Although fast degradation of NP by pure cultures of Sphingomonas strain TTNP3 was
reported in literature (Corvini et al., 2004, Soares et al., 2005), preliminary tests to the
bioaugmentation experiment were carried out.
Mixtures containing various amounts of Sphingomonas and activated sludge from an
acclimated MBR were tested in presence of 14C-NP (Figure 5.4.1). NP recovered after one
hour of incubation at 37oC was quantified by means of HPLC-LSC analysis. The values are
expressed according to the radioactivity concentration injected into the HPLC. The
degradation rate of NP decreased inversely to the amount of Sphingomonas cells added in the
test. In order to compensate the decrease of degradation activity observed by mixing strain
TTNP3 with activated sludge and to prevent a strong disturbance of MBR biocoenosis, a five-
time inoculation of MBR strategy (once per day) applying each time 100 mg of cell dry
weight of Sphingomonas sp. strain TTNP3 per liter of activated sludge was chosen for the
main study.
0
10
20
30
40
50
60
70
80
90
100
% of Sphingomonas added in the test system
NP
degr
adat
ion,
(%)
pure culture
3.0 1.5 1.0 0.5
Figure 5.4.1. Degradation of NP in assays containing various amounts of dry weight Sphingomonas cells/activated sludge (w/w) after 1 h incubation at 37oC. The error bars represent the minimum and maximum values.
74 Results
5.4.2. Radioactivity monitoring
14C-NP was spiked in the MBR and the first bioaugmentation step with Sphingomonas was
accomplished. The, radioactive monitoring was performed in all compartments of the MBR
(MLSS, effluent, NaOH and ethylene glycol solution) in order to follow the effects of
bioaugmentation compared to the non-amended experiment.
In the MLSS of the control MBR, the radioactivity decreased rapidly to approximately
75% of the initial amount within 24 hours followed by a phase with a slower effect and after a
steady slope for the rest of the experimental period (Figure 5.4.2). The concentration of 14C in
the MLSS dropped down to 60% from intial applied radioactivity after the first addition of
Sphingomonas and the effect continued during the next five days following the four
successive pulses (one per day) with pure Sphingomonas strain. After each pulse, the
radioactivity concentration in the MLSS declined markedly, and this decrease was even
observed when the MBR was no longer inoculated with fresh Sphingomonas cells.
The radioactivity quantified in the effluent of the bioaugmentation study was six fold
higher than that in the effluent of the control experiment starting from the first day. Finally,
the radioactivity in the MBR with bioaugmentation was exhausted after 8 days, in contrast to
the control experiment with continous release of radioactivity over 30 days (Figure 5.4.3).
The monitoring of volatilized fraction showed that NP is not stripped out from the system, the
measured concentration remained at the lower detection limit of the LSC for both
experiments, while the 14C-CO2 measured in the bioaugmentation study was a bit higher than
that of the control experiment. The highest amounts of radioactivity in the CO2 trap were
registered during the first days and they decreased the same way as the concentrations of NP
in the MBR.
Results 75
0
10
20
30
40
50
60
70
80
90
100
0 5 10 15 20 25 30 35
Time (days)
% o
f ini
tially
app
lied
radi
oact
ivity
control MBR bioaugmented MBR
Figure 5.4.2. Radioactivity measured in the MLSS of the MBR (control and bioaugmentation experiment).
0
2
4
6
8
10
12
14
1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30
Time (days)
% o
f ini
tially
app
lied
radi
oact
ivity
MBR control
MBR bioaugmentation
Figure 5.4.3. Radioactivity monitoring in the MBR effluent of the MBR (control and bioaugmentation study).
76 Results
5.4.3. Sludge analyses
After each bioaugmentation step, the formation of non-extractable residues in the activated
sludge was investigated. The sludge samples from the MBR were extracted by liquid-liquid
extraction. The same procedure was followed also for the control experiment. The
extractability of the radioactivity from the sludge pellet was higher in the bioaugmented
system than in the control and slowly decreased in the time course of MBR operation.
Conversely, the proportion of non-extractable residues increased during the time
course of both experiments and the amounts were quantified by means of combustion. This
phenomenon was more prominent in the non-amended system where the formation of bound
residues was probably favored by the long contact time between NP and the sludge matrix
(Figure 5.4.4.). The radioactivity in the water phase separated from the MLSS in the control
study amounted to less then 10% for the whole experimental period while it was markedly
higher in the bioaugmentation study (around 30%) (Figure 5.4.5).
0
10
20
30
40
50
60
70
80
90
Rad
ioac
tivity
[%]
water phase from MLSSpellets extraction non-extractable compounds associated with MLSS pellets
day 1 day 3 day 4 day 6 day 8 day 10 day 18 day 22 day 30
Figure 5.4.4. Distribution of the radioactivity in the MLSS pellets and supernatant fractions of the activated sludge in the control MBR. Percentages of radioactivity in each fraction are expressed relatively to the radioactivity measured in the samples before fractionation (100%).
Results 77
0
10
20
30
40
50
60
70R
adio
activ
ity [%
]
water phase from MLSS pellets extractionnon-extractable compounds associated to MLSS pellets
day 1 day 2 day 3 day 4 day 5 day 6 day 7
Figure 5.4.5. Distribution of the radioactivity in the MLSS pellets and supernatant fractions of the activated sludge in the bioaugmented MBR. Percentages of radioactivity in each fraction are expressed relatively to the radioactivity measured in the samples before fractionation (100%).
5.4.4. Detectability of Sphingomonas in the bioaugmented MBR
PCR-DGGE
In order to ensure that the differences observed between the two MBR studies were due to the
addition of amended microorganisms, PCR-DGGE was implemented. Practically, this method
allowed gathering evidences for the presence and persistence of Sphingomonas sp. strain
TTNP3 in the bioaugmented MBR system.
In order to choose adequate primer sets allowing for the detection of this strain in
activated sludge, partial sequencing of the 16S rDNA of this strain was performed. The
sequence of 1398 bp obtained was compared to other sequences in the NCBI-Blast and a 99%
sequence identity with Sphingomonas wittichii RW1 was found (Yabuuchi et al., 2001) (see
Scheme A1 and Sequence A2). Then, PCR primer sets specific for sphingomonads allowing
for the detection of S. wittichii were used (Leys et al., 2004).
16S rRNA gene fragments obtained from non-amended activated sludge, from a pure
culture of Sphingomonas strain TTNP3 and from the MLSS of the bioaugmented MBR were
loaded on a DGGE gel (Figure 5.4.6). Lane 12 corresponding to MBR inoculum prior to
bioaugmentation was clearly different to lanes corresponding to samples containing
78 Results
Sphingomonas strain TTNP3. DGGE profiles of pure cultures of Sphingomonas strain TTNP3
(lane 11) and of direct colony PCR of this strain (lane 1) consisted in a single band,
corresponding to previous observations (Leys et al., 2004). Lane 10 representing a 1:1(w/w)
mixture of Sphingomonas strain TTNP3 and inoculum from MBR (before bioaugmentation)
mixed shortly before DNA extraction showed a single band corresponding to Sphingomonas
strain and thus proved that the DNA of this strain can be extracted from a mixture that
contains a complex matrix. Lanes 5-9 corresponding to amplified DNA extracts after
inoculation of the MBR at the days 1-5 displayed the band characteristic of Sphingomonas.
Lanes 2-4 corresponding to the days 19-21 after the last addition of Sphingomonas and show
DGGE profiles which still contain the characteristic band of this strain.
1 2 3 4 5 6 7 8 9 10 11 12
Figure 5.4.6. DGGE profile of rDNA samples of Sphingomonas sp. strain TTNP3 (band 1, 11), bioaugmented MBR (bands 5, 6, 7, 8, 9) and days 19, 20, 21 after the bioaugmentation period (bands 2, 3, 4). DNA fragments were amplified with primers Sphingo 108f/GC40 and Sphingo 420r. Band 10 corresponds to a mixture of Sphingomonas sp. strain TTNP3 and inoculum from MBR (before bioaugmentation) freshly mixed prior to DNA extraction. Band 12 is corresponding to the endogeneous microflora present in MBR before introduction of Sphingomonas.
Metabolite pattern in effluent extracts
The activity of Sphingomonas in the amended MBR could be investigated by analyzing the
degradation intermediates in the effluent extracts of the investigated systems by means of
HPLC-LSC (Figure 5.4.7). In the effluent of the control MBR, degradation products of 14C-
NP were released and corresponded to compounds with shortened and oxidized alkyl chains
Results 79
as previously described for the MBR operated under the same conditions (see Chapter 5.2.)
(Cirja et al., 2006). The NP remaining in the effluent amounted to approximately 5% of the
radioactivity recovered in the effluent extracts. In effluent extracts from the bioaugmented
system, only two major peaks in the HPLC chromatograms were observed. Subsequent GC-
MS analyses of respective fractions collected from the HPLC (between three and five
minutes) led to the identification of hydroquinone. This compound accounted for 45% of the
radioactivity detected in effluent extracts, while original NP (retention time 22 min) amounted
to 55%.
Figure 5.4.7. HPLC chromatograms of effluent extracts from bioaugmented (top) and control MBR (bottom).
5.4.5. NP bioavailability in presence of Sphingomonas sp. strain TTNP3
The results of a detailed MLSS analysis (Figure 5.4.5.) indicated 26 to 32% of radioactivity in
the water phase of the bioaugmented MBR, while in the control the radioactivity in the water
phase remained below 10%.
The impact of Sphingomonas addition was also visible by measuring six times more
radioactivity in the effluent treated by the bioaugmented system (after the first amendment)
than in the control MBR. From the effluent analyses at the time of inoculation as well as three
80 Results
hours later it was obvious that the addition of fresh resting cells led to an increased release of
radioactivity with the effluent (Figure 5.4.8).
This phenomenon could be attributed to the ability of Sphingomonas to excrete
biosurfactants resulting in a decreasing surface tension in the aqueous environment
(Rosenberg et al., 1999, Deziel et al., 1996, Brown 2006). Surface tension measurements
were performed. It could be demonstrated that activated sludge had a liquid surface tension
coefficient of 62 mN•m-1, which was lower than that of pure water (71 mN•m-1) (Figure
5.4.9). Addition of Sphingomonas strain TTNP3 to sludge at a ratio of 1% dry weight clearly
decreased the surface tension to 50 mN•m-1.
0
100000
200000
300000
400000
500000
600000
0 20 40 60 80 100 120 140 160 180Time (h)
Rad
ioac
tivity
(Bq/
L)
effluent monitoring
IIIII
IV V
I
Figure 5.4.8. Influence of successive inoculations of the MBR with Sphingomonas sp. strain TTNP3 on the radioactivity measured in the effluent. I, II, III, IV and V indicate the time of inoculation of Sphingomonas strain TTNP3 to the MBR.
Results 81
0
10
20
30
40
50
60
70
80Su
rfac
e te
nsio
n (m
N/m
)
Water MLSS MLSS + Sphingomonas
Figure 5.4.9. Surface tension coefficient of water, sludge and a mixture of activated sludge and Sphingomonas.
5.4.6. Radioactivity balance
At the end of the study a balance based on radioactivity was performed. The radioactivity
recovery rates were higher then 90% for both MBRs (Table 5.4.1).
At day 7 (eight days after spiking) less than 1% of the initially applied radioactivity
was still present in the MLSS of the bioaugmented MBR, while in the non-bioaugmented
system 10% remained after 30 days. The removal of radioactivity through withdrawal of
excess sludge was 9% and 20% in the bioaugmented and control MBR, respectively. The
difference in removal through excess sludge was due to the longer operation period of the
control MBR and to the higher proportion of radioactivity associated to the MLSS of the
control MBR in comparison to the bioaugmented MBR.
In the bioaugmented MBR, 56% of the initially applied radioactivity was measured in
the effluent collected over the 8 days operation period, while 30% were recovered in the
treated effluent of the control MBR within the 30 days of operation. Ultimate degradation of
NP, i.e. mineralization to14C-CO2, increased from 1% in the control experiment to 2.3% in the
bioaugmented MBR. Sorption to the components of MBR was 25% for bioaugmented system
and 30% in the control, respectively.
82 Results
Table 5.4.1. Radioactive balance performed after bioaugmentation and control studies in MBR.
Experiment Control Bioaugmentation
Effluent 30 56
Sorption to the sludge (in MBR) 10 1
Excess sludge 20 9 14C-CO2 <1 2.3 14C-VOC <1 <1
Sorption to MBR components 30 25
Sum 92 94
83
6. Discussion
6.1. Performances of MBR: optimization and operational parameters
The present MBR system was designed and operated as close as possible to the parameters
and performances of a full-scale MBR. The efficiency of the MBR during the whole operation
was evaluated in terms of COD, TN, and TP removal as well as robustness of the membranes
and stability of the process over long operation periods.
Due to the missing P-removal and denitrification steps in this MBR system, it was
necessary to adjust the values of the nutrients in the synthetic wastewater used as influent in
order to keep the effluent quality within the required standards. After these modifications,
parameters remained almost constant after an equilibrium period of one week. Nevertheless,
the MBR was operated over a period of 120 days in order to proof the robustness of the
reactor in long-term operation. Typical variations such as higher nitrification rates, decrease
of pH, or total elimination of COD during some particularly hot days of the summer period
were evidenced during the daily monitoring of the treated effluent.
Beside the requirements for water quality, the lab-scale MBR had to fulfill the
conditions for working with radioactive material. The most important aspects were the
tightness of the system to avoid the uncontrolled release of radioactive gases in the laboratory
and the production and disposal of radioactive waste. For the gas problemacy, an efficient
solution was applied which consisted in gas trap liquids able to absorb volatile organic
compounds (e.g. monoethylene glycol) and CO2 (e.g. NaOH 1M) respectively. The
production of radioactive liquid waste was managed by adjusting the operational parameters
of the MBR (SRT, HRT) in a way that radioactive effluent and excess sludge amounts were
limited to 1.6 L/day. The disposal of radioactive wastes is in fact one of the largest practical
limitation of working with radio-markers.
Disposal of radioactive waste is a complex issue, not only because of the nature of the
waste, but also because of the complicated regulatory structure for dealing with radioactive
waste. There are a number of regulatory entities involved. Proper disposal is essential to
ensure health protection, safety of the public and quality of the environment including air,
soil, and water supplies (EPA, 2004).
This work is the first one developing a MBR and using it for fate studies of
hydrophobic organic micropollutants with the help of radio-markers. The radioisotope tracing
method is a suitable technique to investigate the removal of trace pollutants and the fate in
84 Discussion
different environmental compartments. It provides the advantageous possibility of calculating
a radioactivity balance of the target compounds. This method has already been applied in
many studies concerning the fate of micropollutants in various environments (Junker et al.,
2006, Layton et al., 2000). In conclusion, the use of radiotracers in environmental studies is
almost unavoidable due to the scientific benefit, although radioactive waste is produced.
6.2. Fate of 14C-4-[1-ethyl-1,3-dimethylpentyl]phenol in lab-scale MBR
6.2.1. General overview of the results
This study aimed to give a better insight into the possible fate of NP during wastewater
treatment by using a lab-scale MBR designed and optimized for fate studies carried out with
radiolabelled compounds. After a single pulse of the 14C-labelled-NP isomer (4-[1-ethyl-1,3-
dimethylpentyl]phenol) as radiotracer, the applied radioactivity was monitored in the MBR
system over 34 days. The balance of radioactivity at the end of the study showed that 42% of
the applied radioactivity was recovered in the effluent as degradation products of NP, 21%
was removed with the daily excess sludge from the MBR, and 34% was recovered as
adsorbed in the component parts of the MBR. A high amount of NP was associated to the
sludge during the test period, while degradation products were mainly found in the effluent.
Partial identification of these metabolites by means of HPLC-tandem mass spectrometry
coupled to radio-detection showed that they were alkyl-chain oxidation products of NP. The
use of a radiolabelled test compound in a lab-scale MBR was found suitable to demonstrate
that the elimination of NP through mineralization and volatilization processes under the
applied conditions was negligible (both less than 1%). However, the removal of NP via
sorption and the continuous release of oxidation products of NP in the permeate were highly
relevant.
The results concerning the study on the fate of 14C- 4-[1-ethyl-1,3-
dimethylpentyl]phenol in the lab-scale MBR also were partly published elsewere (Cirja et al.,
2006).
Discussion 85
6.2.2. Radioactivity monitoring
One advantage of using radiolabelled compounds for fate studies are the fast measurement of
radioactivity by means of LSC and the overview of its distribution in the investigated system.
After spiking the MBR with 14C-NP, a fast decrease of radioactivity inside of the MBR was
registered. Almost 40% of the initially applied radioactivity was not longer measured in
MLSS in just one day. It was assumed that the NP had a fast distribution into the system
immediately after spiking, due to its highly hydrophobic properties (Kow = 4.8).
A fast sorption can be favoured by the fact that the wastewater treatment system was
not previously fed with NP. In this study, the most amenable sites for a fast sorption were
component parts of the MBR like membrane modules, connection tubes, the aeration system
or the glass walls. The high sorption capacity was corroborated with the low amounts of
volatilized and mineralized NP in the MBR (< 1%). Due to the fact that the MBR was not
equipped with an open surface the volatilization was not stimulated, and it remained at the
lower detection limit over for the whole experimental period. The background for the low
volatilization rate detected can also be an experimental artifact: If NP is volatilized and then
sorbed to the hoses directing the exhaust gases, this amount was recovered in the sorption and
not in the volatilization fraction. In this way, the volatilization should have been
underestimated.
The release of radioactivity into the effluent was small in the first day, but reached the
highest peak (around 2% from initially applied radioactivity) in the next two days. The loss of
radioactivity in the first day was corroborated with the sorption of NP to the connection tubes
from the MBR, to the effluent tank, and to the membrane modules. The daily monitoring of
the MBR over 34 days, showed a continuous decrease of radioactivity in the MLSS to finally
1.8% of the initially applied radioactivity at the end of the study. During the last period of the
monitoring it was obvious that the removal of radioactive compounds from MBR occured just
via excess sludge withdrawal. This observation was in line with the radioactivity
measurments in the effluent for this time period, which showed concentrations in the range of
the lower LSC detection limit of 1 Bq. The slow release of radioactivity lasted for a long
period and was assumed to be caused by the decrease of NP bioavailability in the MBR. High
contact time of NP with the MLSS and continuous growth of the biomass inside the
bioreactor can lead to the immobilization of hydrophobic molecules (Ehlers & Loibner 2006).
86 Discussion
6.2.3. Sludge analyses
The fate of NP in activated sludge can be investigated by means of not adapted sludge
analyses. The MLSS separated in water phase and pellet, emphasised the fast distribution of
NP to the solid fraction. Around 10% of NP was attributed to the water phase during the
experiment while the rest was contained in the sludge pellet in reversibly or irreversibly
bound form. The fast repartition of NP to the sludge confirmed the results found in the
literature (Fauser et al., 2003; Langford et al., 2005). This suggests that the sorption
mechanism is determined by the affinity between the alkyl chain of NP and the organic matter
of the MLSS. Vinken (2005) investigated the association of organic compounds with humic
acids (e.g. phenol, BPA with C3 chain, NP with C9 chain and heptadecanol with C16 chain)
and the sorption increased in the order of phenol<BPA<NP<heptadecanol. This statement is
also supported in the present study by the fact that the metabolites were recovered in the
effluent, while NP remained bound to the sludge matrix. If the sorption was determined by the
phenolic ring, no degradation products should have been occurred in the effluent. In literature
the dependence of the adsorption intensity of nonylphenol ethoxylates on the length of the
ethoxylate chain is reported as the adsorption slightly increased for homologues from NP3EO
to NP13EO (John et al., 2000).
The radioactivity extractable from the sludge pellets with the applied extraction
method was defined as the reversible bound residues. This fraction decreased during the time
course of the experiment, while the formation of non-extractable residues measured by
combustion increased proportionally. The HPLC-LSC analyses of the extractable fraction
indicated that the radioactivity was exclusively due to NP itself. In contrast, in the water
phase similar compounds to that in the effluent extracts were measured (see Chapter 6.2.3).
The phenomenon of the formation of non-extractable residues is a restriction of the
access of microorganisms to NP and thus an impediment to further biodegradation. This
statement was already described in the literature for the problemacy of the entrapment of
pesticides into the soil matrix and the subsequent reduced access to microbial degradation
(Golovleva et al., 1990). In contrast, Ortega-Calvo et al. (1998) reported increased
degradation values of phenantrene when immobilized in the presence of humic fractions.
The sorption of NP to the sludge and its retention practically is equal to a removal
from the effluent and thus, is a benefit for the environment. On the other hand its persistence
in the sludge in combination with the spreading of the sludge to the agricultural fields
represents a risk in countries where the sludge from wastewater treatment plants is still used
as amendment to the fields. The main objective of sludge treatment (aerobic or anaerobic) is
Discussion 87
to reduce its volume and the level of its fermentable organic matter (Hernandez-Raquet et al.,
2007). Additionally, at the moment there is an increasing interest in enhancing the efficiency
of sludge treatment processes, with respect to the reduction of the content of micropollutants
(Nowak et al., 2007).
6.2.4. Effluent analyses
The aim of the effluent analyses was to assess whether the parent compound persists in the
effluent or the radioactivity is recovered as degradation products. From the HPLC-LSC
measurements of the effluent extracts a whole range of radioactive peaks was obtained. All
peaks possessed shorter retention time than NP and emphasized the presence of polar
degradation products.
The group of polar metabolites in the permeate extracts with retention times (RT)
shorter that of NP (RT 21 min) could not be detected in the organic extracts from the MLSS
pellets. Due to their increased hydrosolubility, the metabolites were most probably not sorbed
to the sludge flocs. This indicates that NP is able to persist in the sludge phase while the
dissolved and bioavailable NP fraction present in the water phase is easily degraded.
Partial identification of the metabolites present in the effluent extracts has been
performed by means of HPLC-tandem mass spectrometry coupled to radio-detection. The
metabolites could be characterized as alkyl-chain oxidation products of NP. Information about
the structure of metabolites can be achieved via decay and analysis of the fragmentation
pattern. Definitive conclusion on the structure identification can only be provided via the
synthesis of the proposed metabolite, followed by comparisons of the retention times and the
MS fragmentation patterns (Zühlke et al., 2004).
LC-MS/MS product ion spectra could not completely enlighten the structures of the
compounds, but general structure elements were identified and confirmed the presence of a
keto group at the sub-terminal position on the alkyl chain of the metabolites. The oxidation of
the alkyl chain has been reported for metabolites of nonylphenolethoxylates in a lab scale
bioreactor and in activated sludge (Di Corcia et al., 2000, Jonkers et al., 2001). Carboxy alkyl
polyethoxy carboxylates were formed via ω-oxidation at both alkyl and ethoxylate parts of the
NPnEO molecules and were strictly observed only for non highly-branched compounds. With
the exception of a study reporting 1,1-dimethyl-substituted carboxylated pentylphenols as a
metabolite of NP in a pure fungal culture (Moeder et al., 2006), neither an alkyl chain
shortening of branched NP isomers nor the formation of alkyl chain-keto derivatives as the
88 Discussion
result of sub-terminal oxidation have yet been documented. The accumulation and release of
metabolites during the MBR process indicates a strong persistence of these compounds and
merits ecotoxicological consideration.
6.2.5. Radioactivity balance and benefits of the study
The amount of NP sorbed to the component parts of the MBR and recovered after desorption
in the organic solvent was the fraction prevented from biodegradation and in the same time
considered as removed from the effluent (Cirja et al., 2006).
If the amount of radioactivity bound to the components of the MBR during the first
day is excluded from the balance a better overview on the real degradation of NP occurring in
the continously working flow-through-system can be achieved. Starting with this reasoning,
66.3% of the initially applied radioactivity remains in the MBR and this amount is prone to
further discussion. Assuming this as the available amount (and normalyzing it to 100%), most
of the radioactivity was eliminated as metabolic products of NP (63.8%). Furthermore, 32.2%
of the radioactivity was removed daily with the excess sludge and the volatilized and
mineralized fractions were 0.24% and 0.99%, respectively. Inside the MBR 2.76% of the
initially applied radioactivity remained at the end of the study.
The radioactivity recovered in the effluent mostly as metabolites of NP proved the
degradation capacity of a non-adapted biomass growing in the MBR. Furthermore, in this
study degradation products of NP could be identified for the first time for “real” environment.
Another important aspect drawn the identification of the NP degradation products was to
discover the preference of the MBR biocoenosis to start the degradation of NP via alkyl chain
oxidation. This finding confirmed the low mineralization rates found, and the mineralization
could have required a degradation starting with the ring (Corvini et al., 2006a). On the other
hand, Telscher et al., (2005) reported mineralization rates of 20-30% in a soil/sludge
incubation of the same 14C-NP isomer over 135 days. Another study reported an increase of
the mineralization rates of NP from 20% to above 35% in the presence of dissolved humic
acids due to an enhancement of its solubility in the medium by its association (Li et al., 2007).
The removal of radioactivity via of excess sludge withdrawal is an indication for pilot
and real scale wastewater treatment plants operated with short SRT that important amounts of
hydrophobic compounds leave the system in this way (Wintgens et al., 2002).
Discussion 89
The high recovery rate (99%) of initially applied radioactivity in the MBR is helpful
for a precise differentiation between the different elimination mechanisms for NP occuring in
the investigated system, i.e. biodegradation, sorption, volatilization and mineralization.
6.3. Behavior of two differently radiolabelled 17α-ethinylestradiols
continuously applied to a MBR with adapted industrial activated sludge
6.3.1. General overview of the results
The fate of two differently labelled radioactive forms of 17α-ethinylestradiol (EE2) was
studied during the membrane bioreactor (MBR) process. The laboratory-scale MBR specially
designed for studies with radioactive compounds was operated using a synthetic wastewater
representative for the pharmaceutical industry and the activated sludge was obtained from a
large-scale MBR treating wastewater from a pharmaceutical factory. Applying mixed liquor
solid suspensions of 8 g/L and a sludge retention time of 25 days over the whole test period of
35 days, C-, N- and P-removals of >95%, 75 and 70%, respectively could be achieved.
Balancing of radioactivity could demonstrate that mineralization amounted to less than 1% of
the applied radioactivity. The NP residues remained mainly sorbed in the reactor, resulting in
a removal of approximately 80% relative to the concentration in the influent. The same
metabolite pattern in the radiochromatograms of the two differently labelled 14C-EE2
molecules led to the assumption that the elimination pathway does not involve the removal of
the ethinyl group from the EE2 molecule.
The results concerning the study on the behavior of the two differently radiolabelled
17α-ethinylestradiols continuously applied to the MBR containing adapted industrial
activated sludge were partly published elsewhere (Cirja et al., 2007).
6.3.2. Radioactivity monitoring
EE2 is generally recognized as persistent in the environment. The presence of the ethinyl
group in the molecule hinders the degradation comparing with the other estrogens E1 and E2.
Two differently labelled EE2 were used in this study, e.g. [20-14C] ethinyl labelled EE2 and
[4-14C] A-ring labelled EE2 in order to gain information concerning the fate of this endocrine
disrupter and possibly to support the identification and characterization of degradation
products of EE2 necessary for the elucidation of the degradation pathways.
90 Discussion
The 14C-EE2 added continuously to the MBR via the influent was monitored in all
compartments of the system. At the beginning of the study it could be observed that the
concentration inside of the MBR (fraction associated to MLSS) increased almost
proportionally to the total amount fed with the inflow, while in the effluent no radioactivity
could be detected. It was assumed that still a sorption capacity existed even if the MBR was
fed with non-labelled EE2 for 29 days prior to the initiation of the radioactive phase.
Saturation equilibrium could not be determined since the concentration of non-radiolabelled
EE2 was not determined.
Sorption equilibria carried out for estrogens showed that an equilibrium nearly was
reached after two days in river sediments (Jürgens et al., 2002) while a final equilibrium after
50 days has been reported in river water sediments (Bowman et al., 2003). The time to reach
equilibrium is obviously related to the type of sorption material as well as the test conditions.
It was assumed that after a shorter period of several hours or days, more than 90% of the
equilibrium concentration is already reached (Mes et al., 2005).
6.3.3. Sludge analyses
After the monitoring, excess sludge samples were worked out in order to gain information on
the radioactivity distributed to the sludge matrix. The composition of non-extractable residues
was undefined and just can be attributed to the parent compound or to degradation products.
Nevertheless, at least 65 to 70% of the radioactivity from MLSS was extractable. The high
amount of radioactivity extractable from the pellet remained almost constant over the five
days and was an indication for a reversible sorption of EE2 to the sludge. This emphasized a
possible replacement of the non-labelled EE2 added in the adaptation period by the 14C-EE2.
Ren et al., (2007) confirmed the reversibility of EE2 adsorption and assumed that the
behavior of EE2 could be considered as an exothermic, physical and reversible process,
resulting in a higher adsorption at lower temperatures.
6.3.4. Effluent analyses
Analyses of the MBR effluent are decisive in order to estimate a possible biodegradation of
EE2 in the MBR. Despite the poor production of CO2, a polar metabolite resulting from EE2
could be detected in the effluent even in minute amounts. The rest of radioactivity (almost
95%) recovered in the effluent extracts was identified as EE2 itself. This result demonstrated
once more the highly resistance of EE2 to biological degradation (Joss et al., 2004).
Discussion 91
Until now only a few studies were performed to investigate the degradation pathway
of EE2, although oxidation metabolites were detected in sewage systems. For instance,
nitrifying bacteria seem to be involved in the oxidation of EE2 into more hydrophilic
metabolites (Vader et al., 2000). Recent studies carried out with Sphingobacterium sp.
isolated from activated sludge established that EE2 catabolism processes via the formation of
E1 (Haiyan et al., 2007), the latter known as an oxidation product of E2 at C17 produced under
both biotic and abiotic conditions (Shi et al., 2004, Layton et al., 2000, Fahrbach et al., 2006).
Additionally to E1 two further catabolic intermediates (i.e. 2-hydroxy-2,4-dienevaleric acid
and 2-hydroxy-2,4-diene-1,6-dionic acids) were described in literature and the authors
assumed that the downstream part of the reported degradation pathway was similar to that of
testosterone in Comamonas testeroni (Haiyan et al., 2007).
In both experiments of the present work, a similar polar metabolite with an RT in the
applied HPLC method of 7.5 min was detected. This led us to conclude that the initial
degradation of EE2 was not initiated by the splitting of the ethinyl group. In the case that the
ethinyl- group is substituted first, the 14C-labelled ethinyl-group would recovered in the
injection peak of the HPLC-radiochromatograms from the MBR study with ethinyl-labelled
EE2 while, radioactive E1 would be detected in the HPLC-radiochromatograms from the
MBR study with A ring-labelled EE2. Consequently, the pattern of radioactivite peaks in the
radiochromatograms of both runs would be completely different, which actually was not the
case.
6.3.5. Radioactivity balance and benefits of the study
In the present study, radioactive EE2 was primarily recovered in suspended solids and solids
remaining attached to the reactor (in total approximately 70%) to which the amount removed
through the withdrawal of excess sludge (5%) should be added. These results confirmed
former literature studies, which describe sorption as the main mechanism of EE2 removal
from wastewater (Braga et al., 2005).
The amount of EE2 adsorbed into the component parts of the MBR system is not
taken in account, the remaining radioactivity was dispersed as follows: 25-31% in the
effluent, 6-7% in the excess sludge, 7-9% in the fouling cake and the highest amount, 53-58%
retained in the MBR as sorbed to the MLSS matrix. The pore size of the used microfiltration
membranes was not able to retain the EE2 molecules. Nevertheless, it was advantageous to
92 Discussion
filtrate just the liquid phase, while the sludge flocs remain in the system and in the same time
to attenuate the release of pollutant compound to the environment.
The formation of a fouling cake at the membrane surface should have contributed to
the retention of EE2 inside the system as similar phenomena were reported previously in the
literature (Nghiem et al., 2002).
Sorption properties of EE2 to activated sludge are well documented in the literature. In
a test of Layton et al. (2000), 20% of 14C-EE2 was distributed in the water phase and 60%
was bound to the sludge after 1 h of incubation at a MLSS concentration of 2 to 5 g/L. In
another study, 68% of EE2 was calculated to be sorbed to the sludge after 3 h of incubation
(Kozak et al., 2001). These findings confirm the results of the present study (70% of
radioactivity was associated to the sludge).
The removal performance of EE2 was comparable to that observed in pilot and real
MBR reported in the literature. For instance, 60% to 70% elimination was observed in a MBR
with an initial EE2 concentration in the wastewater of 3 ng/L (Clara et al., 2004). Initial
concentrations of the pollutant in raw water in the range from 2 to 100 ng/L led to EE2
removal rates above 80% during the MBR process (Zühlke et al., 2006; Joss et al., 2004).
Although in the present study the applied concentration of EE2 was quite high (100
µg/L), the elimination rate from the effluent remained high. In a study with a ten-fold higher
EE2 concentration in the wastewater (1 mg/L) gained from a hospital the elimination rate was
only 43% (Pauwels et al., 2006). In a further study carried out in an MBR by Urase et al.,
(2005), the applied EE2 concentration was similar to the one applied in our work and removal
rates of 60 to 70% were reached. However, the MLSS concentration in this study was
relatively low in this case, i.e. 2.7 to 3.5 g/L.
The low extent of mineralization observed in our study is in agreement with a report
stating the recalcitrance of EE2 in batch reactors (Braga et al., 2005). Although some authors
reported the production of 14CO2 from radioactive EE2 in batch cultures, the extent of
mineralization is not realistic to be compared with data on the potential of industrial sludge to
degrade EE2.
A comparable study where a radiolabelled form of EE2 is applied continuously to a
bioreactor has not been yet reported in the literature. Nevertheless, from the results of a study
on the removal of EE2 and other estrogens in a wastewater treatment plant of a contraceptive
producing factory the authors hypothesized that the removal rate in a WWTP treating highly
contaminated effluent is smaller than in municipal treatment systems (Cui et al., 2006). The
explanations were the higher loading with steroid estrogens in comparison to municipal plants
Discussion 93
and the toxic composition of industrial effluent wastewaters, in general. In the present work
the feeding of the microbiocoenosis with a mixture of organic solvents and the steroid EE2
(i.e. recalcitrant xenobiotic) without the addition of vitamins constitute a rational explanation
for the lack of ultimate biodegradation.
6.4. Bioaugmentation of MBR with Sphingomonas sp. strain TTNP3 for the
degradation of nonylphenol
6.4.1. General overview of the results
The aim of this study was to evaluate the efficiency of bioaugmentation of a membrane
bioreactor in order to improve the degradation of recalcitrant nonylphenol during the
wastewater treatment. The 14C-labelled NP isomer 4-[1-ethyl-1,3-dimethylpentyl]phenol was
applied as single pulse to a membrane bioreactor bioaugmented with the bacterium
Sphingomonas sp. strain TTNP3. The effects of five successive inoculations of the membrane
bioreactor with this strain able to degrade NP were investigated in comparison to a non-
bioaugmented reactor. Results showed that the radioactivity spiked in the bioaugmented
system was retrieved mostly in the effluent (56%), followed by fractions sorbed to the system
(25%), associated with the excess sludge (9%) and collected from the gas phase as CO2
resulting from mineralization (2.3%). The degradation products identified in the treated
effluent and in the MLSS were specific metabolites of catabolism of the NP by
Sphingomonas, e.g. hydroquinone resulting from ipso-substitution. The capacity of this
bacterium to excrete biosurfactants and to increase nonylphenol bioavailability was
investigated. The presence and persistence of the strain in the membrane bioreactor was
examined by performing polymerase chain reaction–denaturing gradient gel electrophoresis
(PCR-DGGE).
The results concerning the study on the bioaugmentation of MBR with Sphingomonas
sp. strain TTNP3 for the degradation of NP were partly submitted for publication in a
scientific journal (Cirja et al., submitted).
94 Discussion
6.4.2. Pre-studies on the degradation of nonylphenol in presence of Sphingomonas sp.
strain TTNP3
Nonylphenol showed an increased interest in the last decades mostly due to its persistence and
resistance to biodegradation. The effort to isolate specialists able to degrade the compound
was successful by identifying degraders belonging to the group of sphingomonads (Tanghe et
al., 1999, Fujii et al., 2001). Sphingomonas sp. strain TTNP3 is one of those bacteria
recognized for the degradation of NP to the corresponding alcohol of the alkyl side-chain and
the formation of hydroquinone via ipso-substitution mechanism (Corvini et al., 2004, Corvini
et al., 2006b). Until now, Sphingomonas sp. strain TTNP3 was investigated mostly in pure
culture (Tanghe et al., 1999, Gabriel et al., 2005, Corvini et al., 2006a).
Beside physico-chemical methods, bioaugmentation is a practical solution to enhance
the removal of recalcitrant pollutants from wastewater (van Limbergen et al., 1998). In the
present study, we examined the suitability of bioaugmentation of an MBR with NP-degrading
sphingomonads to improve the elimination of this endocrine disrupter by inoculating
repeatedly freeze-dried cells of pure Sphingomonas strain TTNP3.
Although the degradation efficiency of exogeneous microorganisms can be affected by
the autochthonous microflora present in high cell concentration in the MLSS, repeated
inoculations of low amounts of the bacterium presented the advantage of maintaining the
degradation of NP without affecting the wastewater treatment process. Often,
bioaugmentation studies are carried out by adding high amounts of microorganisms to the
medium to be decontaminated (e.g. 5% used for groundwater bioaugmentation) with good
results but low chances for a technical applicability (de Wildeman et al., 2004).
In order to suit the degradation process to the real conditions, pre-studies are
necessary. In a control carried out with pure cultures, Sphingomonas was able to degrade NP
with a rate of 90% but only very low amounts of the pure culture strain had to be added to the
MLSS of the MBR in order to avoid the disturbance of the system. This aspect could be
improved by successive bioaugmentation of the MBR with Sphingomonas to keep the
degradation rate of NP and to maintain always fresh bacteria, assuming that they have a short
life, are washed out or can be inhibited by the endogenous biocenosis present in the activated
sludge (Stephenson & Stephenson 1992).
Discussion 95
6.4.3. Radioactivity monitoring
In the present study the fate of NP was different in the MBR in comparison to the non-
amended system when Sphingomonas strain TTNP3 was amended even at low amounts (1%
w/w). The first relevant observation during the monitoring of radioactivity was the fast
decrease of radioactivity inside of the bioaugmented MBR. In comparison, in the similar test
experiment carried out as control the radioactivity in the MBR was exhausted over 30 days.
The monitoring of 14C-CO2 emphasized the capacity of the added bacteria to
mineralize NP comparing to the control experiment with only insignificant 14C-CO2
production. The mineralization was doubled when the bacteria were added and this fits to the
fact that Sphingomonas strain TTNP3 is known for its capacity of mineralizing the aromatic
ring of NP (Corvini et al., 2004).
6.4.4. Sludge analyses
The addition of Sphingomonas led to differences concerning the formation of non-extractable
residues. The amount of radioactivity in the non-extractable residues was higher in the study
with the non-amended MBR, while more radioactivity could be recovered with the solvent
extraction fraction and in the water soluble fraction in the bioaugmented system over the
whole period. In the non-bioaugmented system, entrapment of NP in sludge particles may
explain the formation of non-extractable residues, while bound residues in the bioaugmented
system might mainly result from the covalent binding of hydroquinone to the sludge matrix.
Three times more NP residues was discharged with the excess sludge from the control
MBR compared to the bioaugmented system. This may also represent a higher risk for the
environment in the case of agricultural use of sludge since physically entrapped xenobiotics
can be released (Gevao et al., 2005).
6.4.5. Detectability of Sphingomonas sp. strain TTNP3 in bioaugmented MBR
PCR-DGGE
The presence of Sphingomonas sp. strain TTNP3 in the bioreactor was monitored by means of
PCR-DGGE using a primer set specific for Sphingomonas species. Even after stopping the
addition of strain TTNP3 the corresponding specific bands could be detected. The presence of
16S rDNA from Sphingomonas strain TTNP3 did not allow drawing a definitive conclusion
on the ability of the strain in the long term to resist in the reactor. Details on kinetics of cell
96 Discussion
lysis and DNA degradation were not assessed, but differences observed between the two
MBR were assigned to the physiology of the bacterium.
Metabolite pattern in effluent extracts
The degradation products recovered in the effluent proved the presence of Sphingomonas in
the MBR. The residues recovered in the effluent treated by the control MBR consisted mainly
of oxidized alkylphenol products, while in the bioaugmented system the degradation pattern
was drastically different. Only NP and one characteristic product of the degradation of NP by
Sphingomonas strain TTNP3, i.e. hydroquinone, were present, demonstrating the shift of the
degradation pathways. This metabolic pattern was in agreement with previous studies of the
metabolism of NP by this strain, which reported hydroquinone as the central metabolite of NP
degradation processing via type II ipso-substitution (Corvini et al., 2006a). This pathway
leads to approximately 70% mineralization of NP in pure cultures of Sphingomonas strain
TTNP3 (Corvini et al., 2004, Corvini et al., 2006). The large difference in the extent of
mineralization in the present study may be based on an increased bioavailability of NP in
presence of of sludge-amended microorganisms.
Another recent study also demonstrated that hydroquinone is a reactive metabolite,
which rapidly binds covalently to humic acids (Li et al., 2007). From this study, it can be
reasonably assumed that hydroquinone which is formed in situ remained partly immobilized
in the sludge and thus was less available for further mineralization processes. This also
implies that the amount of toxic hydroquinone can be reduced by binding to the sludge before
reaching surface waters.
6.4.6. Bioavailability of NP in presence of Sphingomonas sp. strain TTNP3
An increased amount of NP was recovered in the effluent treated by bioaugmented MBR.
Sphingomonads are known for their ability to produce biosurfactants as part of their cell
membrane or in excreted form (Johnsen & Karlson, 2004). Measurments of the surface
tension helped to explain the increased bioavailability of NP in the presence of Sphingomonas
and to undertake the challenge of bioaugmentation and removal strategy during MBR
treatment.
In order to utilize the production of biosurfactants by amended microorganisms during
the bioaugmentation steps, successive inoculations of the MBR in combination with an
increased HRT could represent an advantageous solution. Benefits of such approach should
Discussion 97
be an increased bioavailability, enhanced biodegradation rates, reduced effluent
concentrations of NP and limited wash out of the inoculated microorganisms.
6.4.7. Radioactivity balance and benefits of the study
Bioaugmentation has been applied to improve the removal of various xenobiotics (e.g.
phenols, chloroanilines, and aromatic hydrocarbons) from wastewater or even to improve the
treatment of specific wastewaters (e.g. paper mill milk, fat degradation or olive mill
wastewater) (Boon et al., 2002; Heilei, et al., 2006; Loperena et al., 2006; McLaughlin et al.,
2006; Olaniran et al., 2006). Nevertheless, various reasons are still driving bioaugmentation
strategies to be inefficient in practical applications.
a) The substrate concentration may be too low to support growth of specialized
bacteria (Stephenson & Stephenson, 1992; de Wildeman et al., 2004). In our study, the initial
amount of NP added in the bioaugmentation test was 2.5 mg/L. Part of it was immediately
distributed to the component parts of the MBR via sorption and the rest remained available to
inoculated Sphingomonas. Beside NP, Sphingomonas is known to be able to utilize other
substrates (e.g. saccharin) (Schleheck et al., 2004) and the synthetic wastewater used in the
present study was enriched with organic carbon sources. Thus, it can be assumed that the
growth support was not a limitation for the bioaugmented species during the present
experiment.
b) The growth of the added microorganisms can be inhibited by other substances
contained in the media to be treated or by the applied operational conditions (e.g.
temperature). It was obvious from the pre-studies carried out before starting the
bioaugmentation experiment in the MBR (see Chapter 5.4.1), that the addition of sludge
inoculums to the Sphingomonas had an impact on the degradation rates of NP. The pure
culture possessed much higher degradation efficiency than mixtures with different
concentrations of sludge. It can be assumed that the Sphingomonas was partly inhibited by the
presence of endogeneous bacteria. Therefore, the bioaugmentation in several steps should
help to counteract the inhibition of specialist degraders already present in the system.
The growth of Sphingomonas is inhibited by the presence of hydroquinone, the
product of NP degradation (Corvini et al., 2006a). As the MBR is a dynamic system, the
produced hydroquinone was expected to be washed out with the effluent. Therefore, higher
concentrations of hydroquinone in the MBR and thus toxic effects to Sphingomonas were not
expected.
98 Discussion
The degradation of NP by pure Sphingomonas strain takes place at 37oC under aerobic
conditions (Corvini et al., 2004). The MBR in the present bioaugmentation experiment was
operated at room temperature (20-22oC). Thus, it has to be expected that Sphingomonas has a
lower activity under such conditions.
c) The competition with autochthonous microorganisms for the substrate may also
affect the biodegradation process.
From the HPLC analysis it could be concluded that the Sphingomonas was not in competition
with the MBR biocenosis to the NP substrate, as the degradation pathway of NP in the
presence of the specialist degraders was completely different in comparison to the control
experiment (see Chapter 5.4.4).
d) Furthermore, the degradation of a target substance sometimes requires long
acclimatization periods before occurence of efficient degradation. In the present study seems
that Sphingomonas started to degrade NP immediately after the first bioaugmentation step
was fulfilled. This could be observed from the fast decrease of radioactivity inside the MBR
and from the hydroquinone residues recovered in the effluent after the first amendment of
Sphingomonas. In this context, Corvini et al. (2006a) reported that a pre-cultivation of
Sphingomonas on NP is not a mandatory need for initiating the degradation.
Based on the radioactivity balance, the fast release of NP from the MBR or the
contaminated wastewater after bioaugmentation can be drawn as main result. Thus, the
highest amount of the initially applied NP was recovered in the effluent as metabolites and as
NP itself. The removal of NP via excess sludge and sorption to the sludge was visibly reduced
comparing to the non-amended system, while the sorption to the component parts of the MBR
could not be avoided.
From the perspectives of a practical application, on the one hand bioaugmentation is
extremely cost-effective compared to other treatment technologies. Furthermore, the
optimization of optimal parameters for the biological treatment including bioaugmentation
strategies may lead to an elimination of the most organic contaminants, e.g. NP, as harmless
or non-persistent products (e.g., hydroquinone). Thus, any future environmental risks or
liabilities could be eliminated in such cases. Bioaugmentation can be a useful tool in the
operation of a modern biological wastewater treatment facility in industrial applications
where the water can be most probably contaminated in waves. Here, bioaugmentation
strategies are able to contribute to cost saving and to operational efficiency.
99
7. Conclusions and outlook
In the present work different fate studies of hydrophobic micropollutants in a lab-scale MBR
have been studied and discussed. The studies were carried with radiolabelled substances and
the MBR was adapted to industrial as well as to municipal wastewater conditions. In the
present chapter some answers will be given to the questions addressed in Chapter 3 of this
work.
• Design, development and optimization of a small-scale MBR with specific functionality
of a real scale system and implementation of operational parameters that provide good
results on wastewater treatment.
Due to the use of radiolabelled compounds in the studies of the present work, it was necessary
to develop and optimize a small-size MBR system in order to avoid the use of high amounts
of radioactivity, to produce high volumes of radioactive waste and to exclude the risk of a
contamination of the working place. The system was developed following the design of a
real-scale MBR. For an effective wastewater treatment operational parameters as a SRT of 25
days, a HRT of 10-15 hours, an activated sludge concentration of 10 g/L and a system
temperature of 20-22oC were implemented. The MBR was adapted for 120 days with
synthetic wastewater at laboratory-conditions. The effluent of the MBR fulfilled the quality
standards required by the surface water quality standards. The designed MBR proved to be
suitable for laboratory fate studies of organic micropollutants.
• Fate and balancing of nonylphenol in an MBR simulating the treatment of municipal
wastewater; one pulse spiking strategy investigations.
For this study a NP isomer was used. This isomer was chosen according to its concentration
in the tNP mixture and due to its recognized endocrine disrupting effect to wild life. The fate
study with 14C-NP emphasised the fast distribution of the compound into the different
compartments of the MBR system. The initially applied 14C-NP did not leave the system
completely within the investigated period of 30 days after the application of NP. The highest
amount of radioactivity was recovered as metabolites of NP in the effluent (42.2%), followed
by the sorption to the component parts of the MBR (33.7%) and the excess sludge collected
over the experimental period (21.3%). The volatilization and mineralization of 14C-NP
seemed to be insignificant in the final radioactivity balancing. The accumulation and release
of metabolites during the MBR process indicates a strong persistence of these compounds and
merits ecotoxicological consideration.
100 Conclusions and outlook
• Fate and balancing of 17α-ethinylestradiol in adapted activated sludge of a
pharmaceutical factory; continuous adaptation of the system to EE2.
A-ring and ethinyl-chain labelled EE2 were used in this study in order to follow the fate of the
targeted compound and eventually to identify degradation compounds in a MBR simulating a
wastewater treatment system of a pharmaceutical factory.
The activated sludge matrix showed a high sorption capacity for EE2. The endocrine
disrupting compound showed resistance to biodegradation even if the system was previously
adapted to EE2. The removal of EE2 from the effluent occurred mostly by a retention inside
of the activated sludge matrix and by sorption to the component parts of the MBR (31-37%),
while the low concentration of metabolites confirmed the biological resistance of EE2 to
biodegradation (5%).
Further investigations are necessary to identify the EE2-degrading microorganism(s)
leading to the production of the discovered metabolite and to determine the optimal
conditions required for its growth and catabolism in order to achieve an improved
biodegradation of EE2.
• Attempts on the isolation and quantification of first metabolites formed by real
degradation of targeted compounds in a MBR system.
Until now only metabolites of NP and EE2 resulting from incubation with the isolated
degrading microorganisms are known. In the present studies, degradation products of NP have
been identified as alkyl chain oxidation compounds. A whole range of compounds was
measured and three of NP metabolites have been identified by means of HPLC-MS/MS as 3-
(4-hydroxyphenyl)-3,5-dimethylheptan-4-one, 5-(4-hydroxyphenyl)-3,5-dimethylheptan-2-
one and 4-(4-hydroxyphenyl)-4-methylhexan-3-one. Attempts on the identification of an
degradation product of EE2 indicated the predisposition to the formation of hydroxylated
derivatives of EE2 but this assumption needs further investigations.
• Implementation of a bioaugmentation strategy for a MBR with specialized degrading
bacteria in order to enhance the biodegradation of NP.
Bioaugmentation of a lab-scale MBR with Sphingomonas sp. strain TTNP3 was investigated
in this study in order to improve and accelerate the degradation of NP from wastewater.
The effects of five successive inoculations of a MBR using this strain specialized to
degrade nonylphenol were investigated in comparison to a non-bioaugmented reactor. Results
showed that the radioactivity spiked in the bioaugmented system was contained mostly in the
Conclusions and outlook 101
effluent (56%), sorbed to the system (25%), associated to the excess sludge (9%) and
contained in the gas phase as CO2 resulting from mineralization (2.3%). The degradation
products identified in the treated effluent and in the MLSS were specific metabolites of the
NP catabolism by Sphingomonas, resulting from ipso-substitution, e.g. hydroquinone. The
presence of Sphingomonas cells in the activated sludge matrix modified the surface tension
and increased the bioavailability of NP in the water phase. These aspects have to be taken into
account for further study optimizations or system up-scalings.
103
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Aerni HR, Kobler B, Rutishauser BV, Wettstein FE, Fischer R, Giger W, Hungerbuhler A,
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9. Appendix
0.1
S. parapaucimobilis JCM 7510 (X72721)S. sanguinis IFO 13937 (D13726)
823
S. paucimobilis GIFU 2395 (D16144)996
S. adhaesiva GIFU 11458 (D16146)507
S. trueperi LMG 2142 (X97776)
428
S. mali IFO 1550 (Y09638)
S. pruni IFO 15498 (Y09637)890
S. asaccharolytica IFO 10564 (Y09639)999
S. echinoides ATCC 14820 (AB021370)613
760
Sphingobium amiensis (AB047364)
S. cloacae (AB040739)539
S. chlorophenolica ATCC 33790 (X87161)362
S. yanoikuyae GIFU 9882 (D16145)998
Sphingomonas wittichii RW1 (AB021492)Sphingomonas sp. TTNP3
1000
488
635
S. aromaticivorans IFO 16086 (U20756)
S. subterranea IFO 16086 (U20773)947
S. capsulata GIFU 11526 (D16147)977
S. stygia IFO 16085 (U20775)975
S. subarctica KF-3 (X94103)
876
S. rosa IFO 15208 (D13945)
857
S. ursincola KR-99 (Y10677)
S. natatoria DSM 3183 (Y13774)1000
S. suberifaciens IFO 15211 (D13737)610
451
S. macrogoltabidus IFO 15033 (D13723)
S. terrae IFO 15098 (D13727)996
843
E. coli K12
Scheme A1. Phylogenetic tree (Neighbour-joining tree including bootstrap values)
126 Appendix
Sequence A2. 1399 bp 16S-rDNA sequence of Sphingomonas sp. strain TTNP3.
ACGCTGGCGGCATGCCTAACACATGCAAGTCGAACGAAGGCTTCGGCCTTAGTGG
CGCACGGGTGCGTAACGCGTGGGAATCTGCCCTTGGGTACGGAATAACTCACCGA
AAGGTGTGCTAATACCGTATGACGACGTAAGTCCAAAGATTTATCGCCTGAGGAT
GAGCCCGCGTTGGATTAGCTAGTTGGTGGGGTAAAGGCCTACCAAGGCGACGAT
CCATAGCTGGTCTGAGAGGATGATCAGCCACACTGGGACTGAGACACGGCCCAG
ACTCCTACGGGAGGCAGCAGTGGGGAATATTGGACAATGGGCGAAAGCCTGATC
CAGCAATGCCGCGTGAGTGATGAAGGCCTTAGGGTTGTAAAGCTCTTTTACCCGG
GAAGATAATGACTGTACCGGGAGAATAAGCCCCGGCTAACTCCGTGCCAGCAGC
CGCGGTAATACGGAGGGGGCTAGCGTTGTTCGGAATTACTGGCGTAAAGCGCAC
GTAGGCGGCTTTGTAAGTTAGAGGTGAAAGCCCGGGGCTCAACCCCGGAATAGC
CTTTAAGACTGCATCGCTTGAACGTCGGAGAGGTGAGTGGAATTCCGAGTGTAGA
GGTGAAATTCGTAGATATTCGGAAGAACACCAGTGGCGAAGGCGGCTCACTGGA
CGACTGTTGACGCTGAGGTGCGAAAGCGTGGGGAGCAAACAGGATTAGATACCC
TGGTAGTCCACGCCGTAAACGATGATAACTAGCTGTCCGGGCACTTGGTGCTTGG
GTGGCGCAGCTAACGCATTAAGTTATCCGCCTGGGGAGTACGGCCGCAAGGTTAA
AACTCAAAGAAATTGACGGGGGCCTGCACAAGCGGTGGAGCATGTGGTTTAATT
CGAAGCAACGCGCAGAACCTTACCAACGTTTGACATCCCTAGTATGGATCGTGGA
GACACTTTCCTTCAGTTCGGCTGGCTAGGTGACAGGTGCTGCATGGCTGTCGTCA
GCTCGTGTCGTGAGATGTTGGGTTAAGTCCCGCAACGAGCGCAACCCTCGCCTTT
AGTTGCCATCATTTAGTTGGGTACTCTAAAGGAACCGCCGGTGATAAGCCGGAGG
AAGGTGGGGATGACGTCAAGTCCTCATGGCCCTTACGCGTTGGGCTACACACGTG
CTACAATGGCAACTACAGTGGGCAGCAATCCCGCGAGGGCGAGCTAATCTCCAA
AAGTTGTCTCAGTTCGGATTGTTCTCTGCAACTCGAGAGCATGAAGGCGGAATCG
CTAGTAATCGCGGATCAGCATGCCGCGGTGAATACGTTCCCAGGCCTTGTACACA
CCGCCCGTCACACCATGGGAGTTGGATTCACCCGAAGGCGCTGCGCTAACCGCAA
GGAGGCAGGCGACCACGGTGGGTTTAGCGACTG
127
Acknowledgments
This work has been carried out at the Institute for Biology V - Environmental Research, RWTH Aachen University, in the frame of AQUAbase Marie Curie Training site, under the supervision of Prof. Dr. Andreas Schäffer between October 2004 and October 2007. My sincerest gratitude and greatest thanks to my supervisor Prof. Dr. Andreas Schäffer for giving me the chance to carry out the research work at his Institute, for his support and correction of this thesis. I would like to express my deep sense of gratitude to my second supervisor Prof. Dr. Philippe Corvini for his inspiring guidance, encouragement and support. Working with him has been a memorable experience. His enthusiasm sustained my research project in spite of many disappointments and difficulties. I would like to thank Prof. Dr. Ing. Thomas Melin for the interest on my research and evaluation of this work as co-assessor. I am thankful to Dr. Pavel Ivashechkin for his great support and advice concerning the MBR and fruitful discussions of the experiments and results. I am grateful to Dr. Sebastian Zühlke (INFU Dortmund University) and Prof. Dr. Juliane Hollender (EAWAG, Switzerland) for their help with the LC-MS/MS analyses and characterisation of the metabolites. Many thanks to Dr. René Ostrowski for his support and help with the characterisation of MBR activated sludge. I am also very grateful to Prof. Dr. Ulrich Klinner for the help concerning the surface tension measurements. I would like to acknowledge my colleagues Gregor Hommes for the help with PCR-DGGE analyses and Boris Kolvenbach for introducing me in the mystery of Sphingomonas sp. Strain TTNP3. A special thank to Rita Hochstrat, the manager of AQUAbase project for her unconditional help on any scientific and social problems appeared during the “AQUAbase time”. I would like to deeply thank my AQUAbase colleagues and friends for being such a warm international family, for all good and bad moments we have shared and for memorable discussions: Lubomira Kovalova, Vassilis Kouloumpos, Paola Cormio, Liang Yu, Hanna Maes, Maxime Favier, Candida Shinn, Konstantinos Plakas, Karolina Nowak, Li Chengliang, Radka Zounkova and Verónica García Molina. I am thankful to all my friends for their encouraging words and to all the people who helped me, and I did not mention them here. I dedicate this thesis to my mother Emilia, my father Haralambie (who did not have the patience to stay and see this book), my brother Catalin and my best Ralph. They gave me the strength to live, the power to go on, the reason to love. Thanks for everything!
129
Curriculum vitae
Personal Information Name: Magdalena Cirja Date of birth: 21. February 1977 Birthplace: Borca-Neamt, Romania Education and training 10. 2004 – 11. 2007 PhD at RWTH Aachen University, Institute of Biology V,
Environmental Biology and Chemodinamics 09. 2003 – 08. 2004 Post-graduated fellowship at University of Milan, Physical
Chemistry and Electrochemistry Department; Marie Curie fellow in ECSE Training Site
10. 2002 – 08. 2003 Assistant researcher at the Technical University “Gh. Asachi”
Iasi, Faculty of Industrial Chemistry, Department of Inorganic Chemistry
09. 2001 – 07. 2002 Master in Environmental Management at the Technical
University “Gh. Asachi” Iasi, Faculty of Industrial Chemistry, Environmental Engineering Department
09. 1996 – 06. 2001 Diploma at the Technical University “Gh. Asachi” Iasi, Faculty
of Industrial Chemistry, Environmental Engineering Department
09. 1991 – 06. 1996 Mihail Sadoveanu Borca – Neamt, Romania, main study:
Chemistry and Biology Publications Journals: Cirja M, Hommes G, Ivashechkin P, Prell J, Schäffer A, Corvini PFX (2008) Bioaugmentation of Membrane Bioreactor with Sphingomonas sp. strain TTNP3 for the Degradation of Nonylphenol (submitted) Cirja M, Ivashechkin P, Schäffer A, Corvini PFX (2008) Factors affecting the removal of organic micropollutants from wastewater in conventional treatment plants (CTP) and membrane bioreactors (MBR) (Review). Reviews in Environmental Science and Biotechnology 7 (1): 61-78 Cirja M, Zühlke S, Ivashechkin P, Hollender J, Schäffer A, Corvini PFX (2007) Behaviour of two differently radiolabelled 17α-ethinylestradiols continuously applied to a lab-scale membrane bioreactor with adapted industrial activated sludge. Water Research 41: 4403-4412
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Cirja M, Zühlke S, Ivashechkin P, Schäffer A and Corvini PFX (2006) Fate of a 14C-Labeled Nonylphenol Isomer in a Laboratory Scale Membrane Bioreactor; Environmental Science and Technology 40 (19): 6131-6136 Fiori G, Rondinini S, Sello G, Verteva A, Cirja M, Conti L (2005) Electroreduction of volatile organic halides on activated silver cathodes. Journal of Applied Electrochemistry 35 (4): 363-368 Conferences & Proceedings: Cirja M, Schäffer A, Corvini PFX (2007) Behaviour of radiolabelled organic micropollutants in a laboratory scale membrane bioreactor. AQUAbase workshop in mitigation technologies, 27-28 November 2007, Aachen, Germany (Oral presentation) Cirja M, Ivashechkin P, Schaeffer A, Corvini PFX (2007) Fate studies of radiolabeled organic micropollutants in a lab scale membrane bioreactor treating wastewater, MICROPOL&ECOHAZARD, 17-20 June 2007, Frankfurt, Germany (Proceeding & Oral presentation) Salehi F, Wintgens T, Melin T, Corvini P, Cirja M, Schäffer A (2007) Einfluss der Wassermatrix auf den Stofftransport von organischen Spurenschadstoffen in NF-Membranen, Aachener Konferenz Wasser und Membranen 2007, Aachen, Germany (Oral presentation) Cirja M (2006) Studies on the behaviour of organic micropollutants in laboratory scale MBR, Entsorga Entega Exhibition, 24-27.10.2006, Köln, Germany (Poster) Cirja M, Ivashechkin P, Pinnekamp J, Ostrowski R, Schäffer A and Corvini PFX (2006) Development of a lab scale membrane bioreactor (MBR) for fate studies with radiolabeled micropollutants; Wasserwirtschaftsinitiative NRW ,“Membrane technology-An essential key for the world wide water protection", 19th June 2006, Brussels, Belgium (Poster) Cirja M, Ivashechkin P, Schäffer A, Pinnekamp J, Ostrowski R, Corvini PFX (2005) Optimization of a lab-scale membrane bioreactor designed for fate studies of radio-labelled hydrophobic micropollutants during the wastewater treatment; National Young Researchers Conference, 27-28 October 2005, Aachen, Germany (Oral presentation) Pharmaceuticals and Hormones in the Environment – Electronic book following the RECETO Summer School, 14-18 August 2005, Brorfelde, Denmark Cirja M, Ivashechkin P, Schäffer A, Pinnekamp J, Corvini PFX (2005) Study on the behaviour of hydrophobic micropollutants during wastewater treatment using a lab-scale membrane bioreactor (MBR) and a radiolabelled single isomer of nonylphenol – 3. Dresdner Tagung Endokrin aktive Stoffe in Abwasser und Klärschlamm, Dresden, Germany (Poster) Cirja M, Forlini A, Rondinini S, Vertova A (2004) Electroreduction of volatile polychlorohydrocarbons on silver electrocatalyst; Giornate dell’Elettrochimica Italiana GEI 2004. 5-9 September 2004, Padua, Italy (Poster) Doubova L, Rondinini S, Vertova A, Cirja M, Forlini A (2004) Metal-substrate Interactions in the Electrocatalytic Reduction of Volatile Polychlorohydrocarbons; International Society of
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Electrochemistry, 55th Annual Meeting, 19-24 September 2004, Thessaloniki, Greece (Oral Presentation)
