REVIEW PAPER
Factors affecting the removal of organic micropollutants
from wastewater in conventional treatment plants (CTP)
and membrane bioreactors (MBR)
Magdalena Cirja Æ Pavel Ivashechkin Æ
Andreas Scha
¨
ffer Æ Philippe F. X. Corvini
Received: 19 July 2006 / Accepted: 8 March 2007 / Published online: 27 March 2007
Ó Springer Science+Business Media B.V. 2007
Abstract As a consequence of insufficient re-
moval during treatment of wastewater released
from industry and households, different classes of
organic micropollutants are nowadays detected in
surface and drinking water. Among these micro-
pollutants, bioactive substances, e.g., endocrine
disrupting compounds and pharmaceuticals, have
been incriminated in negative effects on living
organisms in aquatic biotope. Much research was
done in the last years on the fate and removal of
those compounds from wastewater. An important
point it is to understand the role of applied treat-
ment conditions (sludge retention time (SRT),
biomass concentration, temperature, pH value,
dominant class of micropollutants, etc.) for the
efficiency of conventional treatment plants (CTP)
and membrane bioreactors (MBR) concerning the
removal of micropollutants such as pharmaceuti-
cals, steroid- and xeno-estrogens. Nevertheless,
the removal rates differ even from one compound
to the other and are related to the physico-
chemical characteristics of the xenobiotics.
Keywords Organic micropollutants Sorption
Biodegradation Wastewater Conventional
wastewater treatment Membrane bioreactor
1 Introduction
Environmental pollution by organic micropollu-
tants is nowadays of great concern, especially,
when it affects the aquatic environment. For
many years, quantification of water pollution was
restricted to monitoring biochemical oxygen
demand (BOD), chemical oxygen demand
(COD), nitrates, phosphates and total suspended
solids (Metcalf and Eddy 2003; EN-ISO-9887
1994). Paralleling the bio/analytical progresses,
the focus on macropollutants related to extensive
industrial and agricultural activities is being enlarged
to micropollutants belonging to diverse classes of
chemicals such as pesticides, pharmaceuticals and
personal care products (PPCP), and industrial
M. Cirja A. Scha
¨
ffer
Institute of Environmental Research – Environmental
Biology and Chemodynamics (BioV), RWTH Aachen
University, Worringerweg 1, 52074 Aachen, Germany
P. Ivashechkin
Institute of Environmental Engineering (ISA),
RWTH Aachen University, Mies-van-der-Rohe-Str.
1, 52074 Aachen, Germany
A. Scha
¨
ffer
IME Fraunhofer Institute, Auf dem Aberg 1, 57392
Schmallenberg, Germany
P. F. X. Corvini (&)
Institute for Ecopreneurship, School of Life Sciences,
University of Applied Sciences Northwestern
Switzerland (FHNW), 40 Gru
¨
ndenstrasse, 4132
Muttenz, Switzerland
e-mail: philippe.corvini@fhnw.ch
123
Rev Environ Sci Biotechnol (2008) 7:61–78
DOI 10.1007/s11157-007-9121-8
chemicals, which are detected in trace amounts
(Daughton and Ternes 1999).
Raising the micropollutants, e.g., pharmaceu-
ticals and personal care products (PPCP), endo-
crine disrupting compounds (EDCs), and natural
estrogens, as topic of interest, analytical chemis-
try technology improved the ability to detect
concentrations of ng/l in aqueous media over the
last decades. Mass spectrometric methods and
techniques combining two chromatographic
separation steps such as GC-MS(MS) and LC-
MS(MS) (Marcomini et al. 1987; Jeannot et al.
2000; Li et al. 2000; Reemtsma et al. 2002; Braun
et al. 2003; Clara et al. 2004a; Kloepfer et al.
2004b; Stehmann and Schro
¨
eder 2004; Einchorn
et al. 2005; Huber et al. 2005; Weber et al. 2005;
Moeder et al. 2006; Ternes 1998; Rychlowska
et al. 2003; Luthje et al. 2004) and the use of
radiolabelled tracers (Ingerslev et al. 2001; Doi
et al. 2002; Lalah et al. 2003; Corvini et al. 2004;
Liebig et al. 2005) are only few examples of
analytical techniques to identify and analyse
organic micropollutants and their degradation
products occurring in wastewater. In parallel,
the importance of detecting micropollutants was
emphasized through the development of biotests
(e.g., specialized to identify compounds with
endocrine disrupting properties), which pointed
out to the high biological activity of some class of
micropollutants. For instance, the ubiquitous
distribution in the environment of EDCs and
their harmful potential was emphasized through
the development of very sensitive biological tests
based on immunological techniques such as
ELISA or on endocrine functions such as yeast
estrogen screen (YES) (Huang and Sedlak 2001;
Aerni et al. 2003; Bringolf and Summerfelt 2003;
Matsunaga et al. 2003) E-SCREEN assay (Soto
et al. 1995), EROD activity assay (Ma et al. 2005)
or combination of bioassays (Oh et al. 2006). Part
of these studies concluded that some of these
compounds [e.g., alkylphenol ethoxylates, bisphe-
nol A (BPA), estrone (E1), 17b-estradiol (E2),
17a-ethinylestradiol (EE2)] can have high
(specific biological) estrogenic activity even at
extremely low concentration (Purdom et al. 1994
;
Jobling et al. 1998). Depending on the dose
exposure, the EDCs are responsible for a wide
range of adverse effects on aquatic organisms,
e.g., feminization of male fish, masculinization of
snails (Desbrow et al. 1998;Ko
¨
rner et al. 2000,
2001; Rajapakse et al. 2002), growth inhibition
(Halling-Sørensen 2000; Cleuvers 2005), immo-
bility (Cleuvers 2004), mutagenicity, mortality
(Robinson et al. 2005), and changes in population
density (Shull and Pennington 1993). Micropol-
lutants are detected in river water world-wide
(Ternes 1998; Kolpin et al. 2002) and wastewater
is identified as substantial release route. Besides,
further contamination occurs via leaching from
solid waste sites, deposition from the air, etc.
Micropollutants are not sufficiently removed
in conventional sewage treatment plants and in
order to prevent the spreading of contamination
to groundwater and soils, the emission of some
micropollutants, which are considered to be pri-
ority compounds, is regulated through the Water
Framework Directive (2000/60/EC). The removal
of micropollutants from wastewater during the
treatment occurs through abiotic transformation,
biological degradation and/or sorption. Among
these mechanisms, sorption to suspended solids
and biodegradation were reported to play pre-
dominant roles. Nevertheless, mechanisms of
removal do not follow a general rule since their
relative contribution depends on the physico-
chemical properties of the micropollutant, the
origin and composition of the wastewater, and the
operational parameters of the wastewater treat-
ment facility.
This article provides an overview on the fate
of representative classes of organic micropollu-
tants, i.e., PPCP and EDCs, during the waste-
water treatment. Conventional treatment process
(CTP) and membrane bioreactor (MBR) are in
the focus of this review since CTP is still
nowadays the most common wastewater treat-
ment process, while MBR is a new promising
technology for municipal wastewater treatment
as well as for industrial one. Due to the lack of
information on MBR and CTP comparative
studies, important classes of micropollutants
such as pesticides were left out of discussion in
the present paper.
CTP and MBR are presented in parallel with
respect to their performance for the removal of
micropollutants. After a short overview on the
fate of micropollutants in CTP and MBR in terms
62 Rev Environ Sci Biotechnol (2008) 7:61–78
123
of bioavailability, sorption, biodegradation, and
abiotic phenomena, the main factors affecting the
removal of organic micropollutants during waste-
water treatment are presented. These factors are
linked to the chemical properties of pharmaceu-
ticals and estrogenic compounds and the opera-
tional parameters of wastewater treatment
process including sludge retention time, biomass
concentration, pH value and temperature of
wastewater.
2 Wastewater treatment plant (WWTP)
The treatment of wastewater aims at the removal
of bulk organic matter (proteins, carbohydrates,
etc.) and nutrients (Clara et al. 2005a). In CTP
and MBR processes, sorption and biological
degradation of organics and assimilation of inor-
ganics by activated sludge, take place. In both
systems, activated sludge consists mainly of floc-
culating microorganisms held in suspension and
contact with wastewater in mixed aerated tanks.
In CTP, the wastewater influent is first submitted
to a mechanical treatment where large particles
are removed from water. After a primary sedi-
mentation stage, where the water flows slowly
through large tanks, wastewater is send to the
biological activated sludge tank. Finally, one
additional separation step is achieved by gravity
sedimentation in an external clarifier. In MBR
process, the mechanical treatment and primary
sedimentation tanks are not carried out. The main
difference between CTP and MBR is the sludge–
liquid separation step. The activated sludge tank
includes a filtration step through micro or nano-
filtration membrane, which retains the solid
particles in the aeration tank. The biomass
separation technique considerably influences the
quality of wastewater effluent (Clara et al.
2005b). Generally, CTPs are operated at 1–5 g/l
mixed liquor suspended solids (MLSS), while in
MBR this concentration is considerably higher,
ranging from 8 to 25 g/l or even more
(Stephenson et al. 2000; Galil et al. 2003;
Ivasheckin et al. 2004a). MBR technology allows
sewage treatment at high MLSS concentration
due to the membrane separation step and 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 perfor-
mance of the wastewater treatment concerning
COD removal. Usually, SRT up to 25 or 80 days
are applied in MBR, while typical 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 and
Williams 2004; Joss et al. 2005). Due to the high
SRT values and complete retention of solids
inside of MBR, biodiversity of the microorgan-
isms is favoured and even slowly growing and free
living bacteria remain in the system (Clara et al.
2005b; Pollice and Laera 2005; Howell et al.
2003). Furthermore, the adaptation of some
microorganisms for the degradation of persistent
compounds contained in sewage, e.g., nonylphe-
nol (NP) and further estrogens, is assumed to be
more likely in MBR than in CTP (De Wever
et al. 2004; Ivasheckin et al. 2004a; Siegrist et al.
2004).
Recognized advantages of MBR are high
effluent quality in terms of COD, nitrogen,
phosphorus, ammonia, retention of suspended
solids and microorganisms, reliable biomass con-
centration, efficient treatment of complex waste
streams and compactness of the installation
(Cicek et al. 1999; Abegglen and Siegrist 2006;
Cornel and Krause 2006). At the opposite, the
high MLSS concentration used in MBR leads to
problems concerning oxygen supply of the micro-
organisms and the membranes require frequent
cleaning (Cicek et al. 2001; Cornel and Krause
2006).
3 Fate of micropollutants in CTP and MBR
The fate of micropollutants during CTP or MBR
treatment depends on physico-chemical proper-
ties of the compound, operational parameters
(biomass concentration, sludge retention time,
Rev Environ Sci Biotechnol (2008) 7:61–78 63
123
hydraulic retention time, temperature and pH) of
wastewater to be treated. In the literature, sorp-
tion and biodegradation are reported to be two of
the most important removal processes of micro-
pollutants from wastewater and both processes
are correlated with the availability of the sub-
strate to the degrading microorganisms (Clara
et al. 2004a; Ivashechkin et al. 2004a; Clara et al.
2005a; Joss et al. 2005).
3.1 Bioavailability
As biodegradation is the primary removal
pathway for organics in the activated sludge
treatment, the degree of bioavailability of a
micropollutant is important (Vinken et al. 2004;
Burgess et al. 2005). In wastewater treatment
plants, 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 acces-
sibility of the substance to microorganisms, while
internal bioavailability is limited to the uptake of
the compounds into the internal cell compart-
ment. 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 enzy-
matic equipment and ability to excrete enzymes
and biosurfactants (Wallberg et al. 2001; Cavret
and Feidt 2005; Del Vento and Dachs 2002;
Ehlers and Loibner 2006). Higher bioavailability
and thus potential for biological degradation of
pollutants depend mostly on the solubility of
these compounds in aqueous medium.
3.2 Sorption
Sorption mainly occurs via absorption and adsorp-
tion mechanisms. Absorption involves hydropho-
bic 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
C
sorbed
(g/l), is usually modelled with a simplified
linear equation (1) (Siegrist et al. 2004).
C
sorbed
¼ K
d
SS C
dissolved
ð1Þ
K
d
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
activated sludge tank, and C
dissolved
(g/l) is the
dissolved concentration of the substance.
3.3 Biodegradation
Biodegradation defines the reaction processes
mediated by microbial activity (biotic reaction).
In aerobic processes, microorganisms can trans-
form organic molecules via the succession of
oxidation reactions to simpler products for
instance other organic molecules or mineralized
to CO
2
(Siegrist et al. 2004; van der Meer et al.
2006). At low concentration, the kinetics of
decomposition of micropollutants occurs mostly
according to a first order reaction (see Eq. 2,
Siegrist et al. (2004).
R
degradation
¼ K
degradation
SS C
micropollutant
ð2Þ
R
degradation
is the degradation rate, K
degradation
is
the degradation constant, SS (g/l) is the concen-
tration of suspended solids and C
micropollutants
(mg/l) is the concentration of micropollutants in
influent supposed to be degraded.
The degradation rates are strongly dependent
upon environmental conditions, such as the redox
potential of the systems and the microbial pop-
ulations present. The acclimatization of microor-
ganisms to the substrate requires time and the
affinity of the bacterial enzymes for the micro-
pollutant in the activated sludge influences the
pollutant transformation or decomposition (Spain
et al. 1980; Matsumura 1989).
3.4 Abiotic degradation and volatilization
Abiotic degradation comprises the degradation of
organic chemicals via chemical (e.g., hydrolysis,
64 Rev Environ Sci Biotechnol (2008) 7:61–78
123
oxidation) or physical (e.g., photolysis) reactions
(Acher 1985; Doll and Frimmel 2003; Bouillon
and Miller 2005; Iesce et al. 2006). Abiotic
processes are not mediated by bacteria and have
been found to be of fairly limited importance in
wastewater compared to the biodegradation of
micropollutants (Stangroom et al. 2000; Lalah
et al. 2003; Ivashechkin et al. 2004a; Katsoyinnis
and Samara 2005; Soares et al. 2006). The
removal of micropollutants by volatilization dur-
ing the activated sludge process depends on
vapour pressure (Henry’s constant) and octanol
water partition coefficient (K
ow
) of the analysed
micropollutant, and becomes significant when the
Henry’s law constant (H) ranges from 10
–2
to 10
–3
(Stenstrom et al. 1989). At very low H/K
ow
ratio,
the compound tends to be retained by particles
(Galassi et al. 1997; Roger 1996). The rate of
volatilization is also affected by gas flow rate and
therefore, high efficiency submerged aeration
systems such as fine bubble diffusers should be
used to minimize volatilization rates in wastewa-
ter treatment plants (Stenstrom et al. 1989).
4 Factors affecting the removal of
micropollutants during wastewater treatment
4.1 Chemical properties of micropollutants
4.1.1 Hydrophobicity and hydrophilicity
Hydrophobicity refers to the physical property of
a molecule that is repelled from a mass of water.
Many of the organic micropollutants found in
wastewater are hydrophobic compounds. Hydro-
phobicity is the main property, which leads to
sorption to the sludge, fat and particulate matter
during the wastewater treatment (Garcia et al.
2002; Ilani et al. 2005; Yu and Huang 2005).
Micropollutants can sorb to suspended solids and
subsequently be removed via the withdrawal of
the excess sludge during the wastewater treat-
ment. Sorption of micropollutants to the solid
phase can be estimated using the K
ow
values,
which reflects the equilibrium of partitioning 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
K
ow
is characteristic for hydrophobic compounds,
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 K
ow
< 2.5, the sorption
to activated sludge is not expected to contribute
significantly to the removal of the pollutants via
excess sludge withdrawal. Between log K
ow
2.5
and 4 moderate sorption is expected and values
higher than 4.0 are synonyms to high sorption
potential (Rogers 1996).
4.1.1.1 The influence of hydrophobicity on the
removal of pharmaceuticals in wastewater
treatment Despite the presence of ionic charges
on antibiotic molecules and their low K
ow
, the
fate of these compounds in wastewater treatment
systems can be influenced by hydrophobic
interactions with the sludge matrix. For instance,
oxytetracycline can sorb to the sludge even if they
are present in the form of zwitterion (Kulshrestha
et al. 2004). Sorption to sewage sludge of
antibiotics in a CTP led to 80–90% removal of
ciprofloxacin and norfloxacin (Giger et al. 2003).
In another study, approximately 80% of
norfloxacin and ciprofloxacin which entered into
the CTP was sorbed to particles in the raw sewage
water (Lindberg et al. 2006). Sorption kinetics of
oxytetracycline to the sludge in a lab scale study
was studied by Kim et al. (2005). At 3.6 g/l MLSS
concentration, 95% of oxytetracycline was
removed from water phase within only 1 h and
the concentration at equilibrium remained
unchanged over 24 h.
In sewage treatment plant, the removal of
some pharmaceuticals (e.g., diazepam, diclofenac,
ibuprofen, naproxen, sulfamethoxazole) was
mainly due to adsorption of those compounds to
sludge present in the biological reactor (aeration
tank) (Carballa et al. 2004). At the end of this
experiment, the removal efficiency varied be-
tween 40 and 60% for the anti-inflammatory
compounds and reached approximately 60% for
sulfamethoxazole. The sorption was even evident
during the primary treatment aiming at fat
separation, whereby the liphophilic properties of
organic pollutants led to removal rates ranging
from 20 to 50%. In another study, Carballa and
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