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Occurrence and fate of pharmaceutically active compounds in the largest
municipal wastewater treatment plant in Southwest China: Mass
balance analysis and consumption back-calculated model
Qing Yan a,b,1, Xu Gao c,⇑, Lei Huang a, Xiu-Mei Gan a, Yi-Xin Zhang a, You-Peng Chen a,d, Xu-Ya Peng a, Jin-Song Guo a,d
a Key Laboratory of the Three Gorges Reservoir Region’s Eco-Environments of Ministry of Education, Chongqing University, Chongqing 400045, PR Chinab College of Geography Science and Tourism, Chongqing Normal University, Chongqing 400047, PR Chinac Chongqing Water Group, Co., Ltd., Chongqing 400015, PR China
d Chongqing Institute of Green and Intelligent Technology, Chinese Academy of Sciences, Chongqing 401122, PR China
h i g h l i g h t s
21 and 18 Target PhACs were detected in the wastewater and sludge.
Mass loads of PhACs per person were calculated and compared with other countries.
Biotransformation/biodegradation was the main removal mechanism for the PhACs.
Construct the back-calculated PhAC consumption model based on influent concentration.
a r t i c l e i n f o
Article history:
Received 12 June 2013
Received in revised form 16 October 2013Accepted 22 October 2013
Available online 21 November 2013
Keywords:
Pharmaceutically active compound
Wastewater
Sludge
Mass balance analysis
Pharmaceutical consumption back-
calculated model
China
a b s t r a c t
The occurrence and fate of twenty-one pharmaceutically active compounds (PhACs) were investigated in
different steps of the largest wastewater treatment plant (WWTP) in Southwest China. Concentrations of
these PhACs were determined in both wastewater and sludge phases by a high-performance liquid chro-
matography coupled with electrospray ionization tandem mass spectrometry. Results showed that 21
target PhACs were present in wastewater and 18 in sludge. The calculated total mass load of PhACs
per capita to the influent, the receiving water and sludge were 4.95 mg d1 person1, 889.94 lg d1 per-
son1 and 78.57lg d1 person1, respectively. The overall removal efficiency of the individual PhACs ran-
ged from ‘‘negative removal’’ to almost complete removal. Mass balance analysis revealed that
biodegradation is believed to be the predominant removal mechanism, and sorption onto sludge was a
relevant removal pathway for quinolone antibiotics, azithromycin and simvastatin, accounting for
9.35–26.96% of the initial loadings. However, the sorption of the other selected PhACs was negligible.
The overall pharmaceutical consumption in Chongqing, China, was back-calculated based on influent
concentration by considering the pharmacokinetics of PhACs in humans. The back-estimated usage
was in good agreement with usage of ofloxacin (agreement ratio: 72.5%). However, the back-estimated
usage of PhACs requires further verification. Generally, the average influent mass loads and back-calcu-
lated annual per capita consumption of the selected antibiotics were comparable to or higher than those
reported in developed countries, while the case of other target PhACs was opposite.
2013 Elsevier Ltd. All rights reserved.
1. Introduction
Recently, overwhelming interest about the presence of pharma-
ceutically active compounds (PhACs) as ‘‘pseudopersistent’’
contaminants in the environment have been shown due to their
potential negative effects on aquatic ecosystems and terrestrial
wildlife (Pomati et al., 2007; Martinez, 2008; Dirany et al., 2011).
A significant fraction of parent PhACs are excreted either as
unmetabolized or as transformation products, via urine and feces
of human body or veterinary, and are introduced into the sewer
systems, which have become the principal entry pathway of PhACs
residues into the aquatic environment (Leung et al., 2012).
Municipal wastewater treatment plants (WWTPs) are regarded
as major barriers that can prevent contaminants in wastewater
0045-6535/$ - see front matter 2013 Elsevier Ltd. All rights reserved.http://dx.doi.org/10.1016/j.chemosphere.2013.10.062
⇑ Corresponding author at: Chongqing water group, Co., Ltd. Chongqing, 400015,
PR China. Tel.: +86 13508351373; fax: +86 2363860805.
E-mail addresses: [email protected], [email protected] (X. Gao).1 Tel.: +86 13883570863.
Chemosphere 99 (2014) 160–170
Contents lists available at ScienceDirect
Chemosphere
j o u r n a l h o m e p a g e : w w w . e l s e v i e r . c o m / l o c a t e / c h e m o s p h e r e
http://dx.doi.org/10.1016/j.chemosphere.2013.10.062mailto:[email protected]:[email protected]:[email protected]://dx.doi.org/10.1016/j.chemosphere.2013.10.062http://www.sciencedirect.com/science/journal/00456535http://www.elsevier.com/locate/chemospherehttp://www.elsevier.com/locate/chemospherehttp://www.sciencedirect.com/science/journal/00456535http://dx.doi.org/10.1016/j.chemosphere.2013.10.062mailto:[email protected]:[email protected]://dx.doi.org/10.1016/j.chemosphere.2013.10.062http://crossmark.crossref.org/dialog/?doi=10.1016/j.chemosphere.2013.10.062&domain=pdf
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from entering the receiving environment. However, WWTPs were
not originally designed to deal with complex PhACs. These sub-
stances and their metabolites enter into WWTPs where some of
them may not be completely removed or transformed during the
treatment process leading them into the receiving environment.
Even higher concentrations were found in effluent than in influent
for some recalcitrant PhACs. In the last decade, numerous pub-
lished literatures have investigated and documented the behavior
and fate of PhACs from various therapeutic classes in the WWTPs
in North American, Europe and Australia (Verlicchi et al., 2012).
However, only a few papers were concerned with the situation in
China, perhaps because of the difficulties of analysis and the
expensive trial costs. To date, main one specific therapeutic class,
antibiotics, has been investigated by limited previous studied in
China (Xu et al., 2007; Gulkowska et al., 2008; Gao et al., 2012a;
Zhou et al., 2013).
Extensive literatures on the concentration levels of PhACs in
aqueous phases such as wastewater or surface water is available
(Li and Zhang, 2011; Leung et al., 2012; Aydin and Talinli, 2013);
however, the presence of PhACs is much less explored in sewage
sludge than in wastewater or surface water because of the great ef-
fort required in analyzing this difficult matrix. An aqueous phase
removal percentage, which is based on the concentrations of PhACs
in the influent and the effluent of WWTPs, is often used as the only
parameter available for calculating the PhAC removal efficiency in
WWTPs (Leung et al., 2012). The sorption onto sludge is a relevant
removal pathway for certain PhACs ( Jelic et al., 2011, 2012; Jia
et al., 2012; Zhou et al., 2013). Thus, an aqueous phase removal
percentage cannot comprehensively assess the removal of PhACs
in WWTPs accurately. Mass balance analysis approach would be
an effective way to understand the fate of PhACs in WWTPs and
their mass loading to the receiving environments.
The level of PhACs in the influent depends on their consump-
tion. Scheurer et al. (2009) reported the occurrence of the widely
used metformin in surface waters in German and concluded that
the high concentrations of metformin in aquatic environment were
in agreement with the consumption data. ter Laak et al. (2010)stressed the potential of using pharmaceutical sales data for the
prediction of concentrations in the aqueous environment.
Kasprzyk-Hordern et al. (2009) estimated the pharmaceutical
usage in local communities based on their concentrations in waste-
water influent. Besse et al. (2008) calculated the predicted environ-
mental concentrations (PECs) of PhACs using drug consumption
data and found that the calculated PECs were consistent with the
field measurements. Rowney et al. (2009) predicted the concentra-
tions of cytotoxic drugs in the catchment area of the Thames River
by considering the consumption data. Sum up above mentioned
studies, we can come to the conclusion that there is a good corre-
lation between PhAC consumption and the residual loads in the
influent for different therapeutic classes and that the analysis of
post-therapeutic residual concentrations in the influent after hu-man administration can be an alternative method to back-calculate
PhAC usage by considering the pharmacokinetics of the target
PhACs in humans.
To date, a few studies reported the occurrence and behavior of
the antibiotics in WWTPs in the fastest developing cities of China
such as Beijing, Guangzhou and Hong Kong (Xu et al., 2007;
Gulkowska et al., 2008; Sui et al., 2010; Gao et al., 2012a; Zhou
et al., 2013), and the results of these studies indicated that the
contamination level of antibiotics varied among cities in China.
However, no information was available in other wide areas in
China. During the past two decades, the Chongqing region in the
southwestern China, having a population of 3.3 million inhabitants,
has become one of the fastest growing economies and most
densely urbanized areas in the world. The pharmaceutical con-sumption in hospitals in the region is about 1.361 billion RMB. It
is supposed that the occurrence of PhACs in the aquatic environ-
ment of Chongqing is of particular interest and may be higher con-
centrations than other regions. All the WWTPs in Chongqing,
which include only two treatments steps (physical and biological)
and do not use tertiary treatment or an advanced sewage treat-
ment (e.g. ultrafiltration, flocculation, ozonation, advanced oxida-
tion, or osmosis), were not originally designed for removal of the
PhACs. Therefore, it is imperative to obtain accurate information
on the elimination of PhACs in these WWTPs to supply the scien-
tific data for the WWTP upgrades and also to provide treatment
alternatives for those PhACs refractory to elimination.
Probably due to the lack of regulations of all kinds of drugs, the
PhACs, especially the antibiotics, were misused seriously in China.
According to statistics, the annual per capita consumption of anti-
biotics is 138 g in China and the figure is 10 times as much as that
of the United States. However, the information on annual pharma-
ceutical consumption in various cities in China is unavailable be-
cause establishing a collective record system for all practitioners,
public and private hospitals, as well as over-the-counter PhACs is
complicated and costly.
Here, based on measured pharmaceutical concentrations in
influent of the WWTP and the pharmacokinetics of PhACs in hu-
mans, we tried to create a back-calculated model for the prediction
of the loads of PhACs to provide a reference for improving current
statutory regulation on pharmaceutical consumption. In addition,
the average mass loads of PhACs per person reported in developed
countries and in this study were calculated and a comparison anal-
ysis was made to have a better understanding of pharmaceutical
occurrence and mass inputs into the environment. Lastly, dewa-
tered sludge was collected to determine the concentrations and
to assess the sorption of target PhACs onto sludge. Based on the
data obtained, mass balance analysis was used to explore their po-
tential removal mechanisms.
2. Materials and methods
2.1. Chemicals and reagents
Eight classes of 21 PhACs were selected for this study: analge-
sics, sulfonamide antibiotics (SAs), macrolide antibiotics (MAs),
quinolone antibiotics (QAs), antiepileptics, cholesterol lowering
statin drugs, lipid regulators and antihypersensitives. The 21 target
PhACs were ibuprofen (IBP), diclofenac (DCF), clofibric acid (CA),
bezafibrate (BZB), simvastatin (SVT), atorvastatin (ATT), carbamaz-
epine (CBZ), erythromycin-H2O (ERY), roxithromycin (ROX), azith-
romycin (AZM), Amlodipine (ALP), moxifloxacin (MOX),
Acetaminophen (ACM), gemfibrozil (GFB), metoprolol (MTP), sulfa-
methoxazole (SMZ), sulfadiazine (SDZ), sulfamethazine (SM1), tri-
methoprim (TMP), ofloxacin (OFX) and norfloxacin (NOR). These
compounds were selected because of their high consumption inChongqing and their being frequently detected in surface and
wastewater. Chemical structures, CAS numbers and physicochem-
ical properties of the 21 target PhACs are shown in Supplementary
Information. Internal standards simatone (SMT), dihydrocarb-
amazepine (DCBZ), caffeine-13C3 (CF-13C) and mecoprop-D3 were
purchased from Accustandard (New Haven, CT, USA), Sigma–Al-
drich, C/D/N Isotopes (Quebec, Canada) and Dr. Ehrenstoefer
(Augsburg, Germany), respectively. Oasis hydrophilic–lipophilic
balanced (HLB, 6 cc, 200 mg) cartridges were purchased from
Waters (Milford, MA, USA). Syringe filters with 0.45 mm pore size
were purchased from Pall Corp., United States. Milli-Q water was
used throughout the study. HPLC-grade methanol was provided
by Merck (Germany).
The individual and internal standard solutions were prepared atconcentrations of 500 mg L 1 by dissolving appropriate amounts of
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PhACs in methanol and were then stored in the dark at20 C. The
dehydrated form of erythromycin (ERY-H2O) is frequently detected
in the environment; therefore, ERY-H2O was measured in this
study. The standard solution of ERY-H2O was prepared following
the method described in our previous study (Gulkowska et al.,
2007).
2.2. Sample collection
The studied WWTP, which is the fourth largest WWTP in China
and the largest WWTP in Southwest China, serves a population of
1540000 equivalent inhabitants and treats up to 600000 m3 d1 of
municipal wastewater. Treatments consists of pretreatment
(screening), primary (settling) and secondary (a cyclic activated
sludge system) (HRT: 10.5 h) treatments. The secondary effluent
is further treated with chlorination before being discharged as final
effluent. The total hydraulic retention time (HRT) and solid reten-
tion time (SRT) were 15.8 h and 21.4 d, respectively. Time-inte-
grated sampling was started at different times of the day in the
WWTP to approximately compensate for the HRT of each treat-
ment step. Sampling started in STP A, influent sewage, at 7:00.
STP B samples of water that had undergone grit chamber and pri-mary sedimentation, were collected starting at 8:30 while STP C
samples after biological treatment and secondary clarifier were
collected starting at 22:30. STP D samples after chlorine disinfec-
tion treatment had a collection delay of 30 min with respect to
STP C samples. Fig. 1 shows a diagram scheme of wastewater with
sampling locations marked in bold.
Grab sludge samples (E ) were taken from the outlet of the
dewatering system (dewatered sludge) as five grab samples with
same quantity, with equal time intervals at the sampling site dur-
ing every sampling campaign. The samples were immediately put
into ice-packed cooler, and mixed to give a single sample and
wrapped in silver paper. The resulting samples were chilled to
4 C for transportation, freeze-dried at 50 C, and stored at
20 C prior to analysis.
Wastewater and sludge samples were collected in four sam-
pling campaigns between November 2012 and January 2013, with
intervals of two weeks. The samples were analyzed in triplicate.
2.3. Analytical methods
Water samples were extracted and cleaned by solid phase
extraction (SPE) by using an Oasis HLB cartridge (6 cc; 200 mg;
Waters Corp., Milford, USA). Sludge samples were extracted by
ultrasonic technology and then cleaned by SPE with an HLB
cartridge. The detailed pretreatment information is listed in
Supplementary Information (Text S1).
All target compounds were separated and quantified by using a
1200 binary liquid chromatography (LC) system coupled with a
6410 QQQ LC/MS equipped with an electrospray ion source (ESI)
(Agilent, USA). The separation of the analytes was conducted on
the ZORBAX Eclipse Plus C18 column (4.6 mm 150 mm;
3.5 lm; Agilent, USA) at a flow rate of 0.25 mL min1. The identifi-
cation of the target PhACs was accomplished by comparing the
retention time (within 2%) and the ratio (within 20%) of the two se-
lected multiple-reaction monitoring (MRM) ion transitions with
the standards. The target PhACs had different physicochemical
characteristics. Therefore, the target compounds were classified
into two groups, namely, group A (detected by positive ionization
mode) and group B (detected by negative ionizationmode), by con-
sidering the analytical conditions (i.e., polarity of produced ion and
mobile phase solvent). The LC conditions and the electrospray ion-
ization tandem mass spectrometry for determining the 21 target
PhACs and the 4 internal standards are summarized in Tables S2,
S3 and S4 in the Supplementary Information.
2.4. Method validation
The quantification of the target PhACs was performed by using
the internal standard calibration approach to eliminate the influ-
ence of the matrix effect. The calibration samples in triplicate withconcentrations ranging from 0.1 lg L 1 to 1000 lg L 1 were pre-
pared by spiking the working solutions in Milli-Q water. The corre-
lation coefficients (r 2) of the calibration curves exceeded 0.99
except for ATT (r 2 = 0.986). Standard solutions were spiked into
tap water to identify the recovery, accuracy and precision of the
analytical method. The samples were extracted and then analyzed.
In the analytical method, the recovery, accuracy and precision of all
target PhACs ranged from 43% to 104%, from 96% to 118% and from
2% to 10%, respectively. To confirm the matrix effects, all target
analytes and internal standards were spiked into three effluent,
three influent wastewater and three sludge samples (extracted
by ultrasonic-assisted technology). The samples were then ana-
lyzed by using the same analytical method. For each type of sam-
ples, recoveries were determined by comparing the concentrationsobtained after the whole SPE procedure, calculated by internal
standard calibration, with the initial spiking levels. As spiked sur-
face, wastewaters and sludge samples already contained target
compounds, blanks (non-spiked samples) were analyzed to deter-
mine their concentrations, which were afterward subtracted from
the spiked samples. Results show that except for ATT, all target
compounds achieved a recovery rate ranging from 41% to 140%
for the influent, from 44% to 158% for the effluent, and from 46%
to 139% for the sludge, with relative standard deviation values un-
der 15%. Only ATT exhibited lower recovery, which is primarily
attributed to the unsuitable conditions chosen for the compounds.
However, the low recovery was not considered an obstacle for
reliably determining ATT in the environmental waters because its
sensitivity was relatively good. The accuracy and precision of the
Fig. 1. Schematic diagram of the studied WWTP and the sampling site locations.
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instrumental analysis over the experimental period was monitored
by the replicate injections of standard solutions at 10 and
100 ng mL 1. The instrumental intra-day and inter-day precisions
for all PhACs ranged from 3% to 15%, and from 5% to 18%, respec-
tively. In our study, the limit of quantification (LOQ) was estimated
for a signal-to- noise (S/N) ratio 10 from the sample chromato-
grams at the lowest validation level tested, using the quantification
transition. The LOQs of the target compounds ranged from
0.03 ng L 1 to 3.4 ng L 1, from 0.2ngL 1 to 17.5ng L 1, from
0.2ng L 1 to 5.6 ng L 1, and from 0.17 lg kg1 to 5.83 lg kg1 for
surface water, influent wastewater, effluent wastewater, and
sludge samples, respectively (Table S4 in the Supplementary
Information).
2.5. Mass balance analysis
The average mass flow of each target compound was calculated
by multiplying the concentration with the corresponding flow. The
equation can be expressed as follows:
M aqueous ¼ Q C aqueous 106 ð1Þ
M sludge ¼ P sludge C sludge 106: ð2Þ
The percentage of each compound (Ri, %) in the effluents from
different treatment units can be calculated based on the following
equation:
Ri ¼ M i=M influent ð3Þ
where M aqueous and M sludge (g d1) are the mass flux of the pharma-
ceutical calculated in aqueous and sludge phase, respectively;
C aqueous (ng L 1) and C sludge (ng g
1) were the measured concentra-
tions in the aqueous phase and sludge, respectively; Q (m3 d1)
and P sludge (kg d1) are the flow rate of wastewater and the produc-
tion rate of sludge, respectively; M i (g d1) is the mass flux in
effluents from different units, M influent (g d1) is the mass influent.
The overall removal efficiency (Roverall) of the selected PhACs
during wastewater treatment was calculated based on Eq. (4):
Rov erallð%Þ¼ðC influent Q þC absorbedQ T ss=10
3ÞðC effluent Q þC sludgeP sludgeÞ
C influent Q
100
ð4Þ
The aqueous removal efficiency (Raqueous) was calculated based
on Eq. (5):
Raqueousð%Þ ¼C influent Q C effluent Q
C influent Q 100 ð5Þ
where C influent (ng L 1), C effluent (ng L
1) and C absorbed (ng g1) were
the measured concentrations in the influent, the effluent and the
suspended solids phase, respectively; T ss (mg L 1) is the concentra-
tion of total suspended solids.
2.6. Calculation of pharmaceutical consumption
In this study, the studied WWTP was considered representative
of the general situation in Chongqing because the plant treats over
50% of the municipal wastewater in the entire area. The PhAC
usage back-estimated from the influent in Chongqing was calcu-
lated according to the following equation:
U ¼ C influent Q 10
9 365:25
ð1 RdisposedÞ ðRabs Rexcreted þ 1 RabsÞP T P S
ð6Þ
where U represents the back-estimated usage of the target PhACs(kg year1); C influent , (ng L
1) and Q (m3 d1) refer to their previous
definitions in Eqs. (4) and (5); Rexcreted, Rdisposed and Rabs refer to
the percentage of the parent PhACs excreted, the ratio of the dis-
posed drugs in the municipal solid waste to the drug sales, and
the absorption rate of the drugs (i.e., bioavailability), respectively;
P S and P T refer to the served population of the studied WWTP and
the total population of Chongqing, respectively. Rexcreted and Rabswere respectively assumed to be 50% and 100% in the absence of
data for certain PhACs. The Rdisposed
in developing countries was
set at 0.2 on the basis of previous literature ( Zhang and Geissen,
2010). The back-calculation model assumed that all PhACs were
prescribed evenly across the territory and that no sorption and
transformation occurred during their conveyance to WWTPs and
before sampling so as not to underestimate the actual PhAC
consumption.
3. Results and discussion
3.1. Occurrence, average daily load and global comparison of the
target PhACs in wastewater
The concentrations of the target PhACs in wastewater taken at
various stages of the treatment and the dewatered sludge samples
are summarized in Table 1. The average wastewater flow during
the sampling periods in the WWTP was 600000 m3 d1. The daily
average dewatered sludge production was approximately
125600kg dw.
All target PhACs were quantified in the wastewater samples
from the studied WWTP (Table 1), with concentrations ranging
from 1.5 ± 0.5 ngL 1 to 7111.7 ± 322.9 ng L 1 in the influent and
0.5ng L 1 to 1147.9 ± 65.1 ng L 1 in the final effluent. Fig. 2 shows
that the composition profiles between the effluents and the influ-
ents were different because of the different removal efficiencies
of the target PhACs in the studied WWTP. In particular, ACM, which
is a widely used over-the-counter analgesic (pain reliever) and
antipyretic (fever reducer), contributed more than 50% to the se-
lected PhACs loads in the influent of the studied WWTP. However,ACM exhibited values below or very close to the LOQ in most of the
effluent samples primarily because ACM is easily biodegraded in
the water phase (Behera et al., 2011). Comparatively higher levels
of antibiotics in effluent were measured in the studied WWTP
(Fig. 2 and Table 1).
The average daily mass load per person (Li, lg d1 person1) of
individual PhACs (i) was obtained by multiplying the concentration
in the sewage (C i, n g L 1) and the average treated flow rate (Q ,
m3 d1) during the sampling period and normalizing this value to
the population served by the corresponding WWTP. The equation
can be expressed as Li = C i Q /Served population. The daily mass
loads per person of the selected PhACs in the current study and
from the literature were obtained by using the above formula,
and the obtained values were divided into groups according tothe study location. The average values are summarized in Fig. 3
(full details are presented in Table S5).
The influent mass loads of the target PhACs can indicate the in-
put into WWTP and reflect the usage pattern of PhACs in the ser-
vice area to a certain extent, assuming that attenuation of a
compound during transportation from toilets to WWTPs was the
same in different areas. By contrast, the effluent mass loads can
be used to estimate the contribution of PhACs to the receiving
water. The sum of the average daily mass flow for all the selected
disposed PhACs was estimated to be 4.95 mg d1 person1 for the
studied WWTP. The calculated total mass load per person to the
receiving water on the basis of the effluent concentration data
from the WWTP was 889.94 lg d1 person1. The data on average
daily mass flows of PhACs reported by other studies are relativelylimited, and the available information is heterogeneous. The high-
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holm and Kalmar in Sweden (influents: 7.13 lg d1 person1 and
effluents: 2.87 lg d1 person1) (Castiglioni et al., 2006), but were
five times lower than the average mass load values detected in Cas-
tellon de la Plana in Spain (influents: 50.73lg d1 person1) (Kart-
hikeyan and Meyer, 2006). By contrast, the levels of GFB, CBZ and
ATT were much lower than those reported in the United States and
Europe (Karthikeyan and Meyer, 2006; Gracia-Lor et al., 2012). To
the best of our knowledge, this study is the first report on the mass
loads of MTP and SVT in wastewater. The average daily mass loadper capita of the target PhACs varied in different countries, reflect-
ing different pharmaceutical usage consumed by human.
3.2. Concentrations of PhACs in sludge
Table 1 shows that 18 out of the 21 analyzed PhACs were de-
tected in the sludge samples. GFB, ACM and IBP were lower than
the LOQs in sludge. In all the samples, the composition profile of
the target PhACs in sludge shows that the mass loads of MAs and
QAs accounted for more than 80% of the total amount of PhACs
found (Fig. 2), which was consistent with those obtained by Nieto
et al. (2010). The concentration levels of PhACs in sludge are clearly
determined by different factors such as the physicochemical prop-
erties, usage and removal percentages (Radjenovic et al., 2009; Xueet al., 2010; Jia et al., 2012). The total load of the detected PhACs
that left the plant unmodified through sludge was 121 g d1 in
the WWTP. Only 1.6% of the total mass load of the target PhACs
was retained by dewatered sludge.
SMZ, which has a 100% detection frequency in wastewater, was
also found in the sludge samples, albeit with relatively low concen-
trations of 9.51 lg kg1 dw in the dewatered sludge. In addition,
SDZ was detected in all collected sludge samples with an average
concentration of 3.57 lg kg1 dry weight (dw) in the dewatered
sludge. SM1 was also found in the sludge, but at extremely low
concentrations (0.92lg kg1 dw). SAs are generally presented as
neutral and anionic species at wastewater with pH 7 and have
low log K ow values (LogK ow < 2.5 high hydrophilic compounds,
see Table S1). Therefore, sorption to sludge is expected to be weakfor SAs because of the electrostatic repulsion from the negatively
charged functional groups in the activated sludge. Other studies
have also found low concentrations of SAs in the sludge of WWTPs.
Gobel et al. (2005) determined different SAs in sewage sludge from
different WWTPs and the authors found that SMZ and TMP were
detected in activated sludge with average concentrations of 68
and 41 lg kg1 dw, respectively. However, another study by Gobel
et al. (2007) found lower concentrations of SMZ and TMP in sewage
sludge from Germany and Switzerland. Nieto et al. (2010) found
that SMZ, SM1 and TMP appeared in only a few samples and al-ways at a concentration lower than LOQs. Gao et al. (2012a) found
that the concentrations of SAs in the sludge samples were rela-
tively low, and SM1 and SMZ were present in 23% and 77% of the
analyzed samples, respectively. Zhou et al. (2013) reported that
the sorption onto sludge was negligible and biotransformation is
believed to be the predominant removal mechanism for SAs.
Conversely, all target QAs contain nitrogen as positively charged
moiety, whereas AZM possesses positively charged dimethylamino
groups in its molecules ( Jelic et al., 2012; Jia et al., 2012). Thus, the
higher sorption potential of QAs and AZM compared with SAs is
possibly caused by the electrostatic interactions involved with
the positively charged locations of these compounds. AZM
(466.76 lg kg1) had the highest sorption to sludge in this study.
Similar results have also been observed in sludge from otherWWTPs in China ( Jelic et al., 2012; Jia et al., 2012; Gao et al.,
2012a) and other countries such as Switzerland (Golet et al.,
2003) and Spain ( Jelic et al., 2012). A study by Zhou et al. (2013)
came to the conclusion that the aqueous removals for QAs mainly
were attributed to the adsorption onto sludge. SVT is a neutral
compound with the highest LogDow of all the studied PhACs,
namely hydrophobic interactions (logDow 4.46 for pH 6–8) (cal-
culated by ACD/logP ow ver. 1.0, Advanced Chemistry Development,
Inc.), thereby resulting in its comparatively high sorption onto
sludge. Jelic et al. (2011) reported that ATT was ubiquitous in
sludge samples in average concentrations from 30 to 60 lg kg1 dw
and sorption contributed to its elimination from aqueous phase
with more than 20% related to the amount in influent. CBZ is
reportedly a recalcitrant to biodegradation and abiotic transforma-tion during wastewater treatment processes ( Jelic et al., 2011). In
0
1
10
100
1000
10000
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
I n f l u e n t
E f f l u e n t
NOR OFX MOX SMZ SDZ SM1 TMP ERY-
H2O
ROX AZM IBP DCF ACM BZB CA GFB MTP ALP ATT SVT CBZ
M a s s L o a d s ( µ g / d a y / p e r s o n )
In thhis study
Liede, Guangzhou, China
Qinghe STP, Beijing,
China
Shatin, Hong Kong
Stanley, Hong Kong
Tai Po, Hong Kong
Brisbane, Australia
Two STPs, Wisconsin,
USA
Six STPs, Italy
Kloten-Opfikon,
Switzerland
Stockholm and Kalmar,
Sweden
Castellon de la Plana,
Spain
Athens WWTP, Greece
Fig. 3. Daily mass loads (lg d1 person1) of PhACs in Chongqing and global comparison with other countries.
Q. Yan et al. / Chemosphere 99 (2014) 160–170 165
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this study, CBZ was detected with relatively low level. Previous
some studies also showed that the K d values for the CBZ were neg-
ligibly low, indicating that this anti-epileptic agent does not adsorb
onto the sludge to an appreciable degree ( Jones et al., 2002; Ternes
et al., 2004; Urase and Kikuta, 2005; Carballa et al., 2008; Jelic
et al., 2012). The results of the EU POSEIDON project (http://posei-
don.bafg.de) also indicated that CBZ does not adsorb onto sludge.
However, Radjenovic et al., 2009 found that the average concentra-
tion of CBZ in primary sludge was more 100 ng L 1 in some sludge
samples, and one possible reason was the difference in the compo-
sition and pH of the sludge. Acidic drugs such as IBP (pK a = 4.9), CA
(pK a = 3), GFB (pK a = 4.7), BZB (pK a = 4) and DCF (pK a = 4) were rel-
atively low or below the method’s LOQs in sludge because of the
considerable ionization present at wastewater pH (averagepH = 7.5). Limited reports are available on the concentrations of
acidic drugs in sludge. Two previous studied by Jelic et al. (2011,
2012) showed that DCF was detected frequently in sludge with
concentrations between 30 and 60 lg kg1 and that BZB and GFB
were found in very low concentration. Martin et al. (2012) found
that DCF and GFB were mainly detected in wastewater instead of
in sludge and their concentrations in sludge were below their
LOQs. Nieto et al. (2010) found concentrations of some tens
lg kg1 IBP in dewatered sludge samples, originating from two
Spanish WWTPs. Recently, Yu et al. (2013) showed that the mean
concentrations of CA, DCF, GFB and IBP in sludge were 36.4, 48.4,
93.3 and 109 lg kg1, respectively. Samaras et al. (2013) reported
that DCF and IBP were found at relatively low concentrations and
that the concentrations of these drugs were below the LOQs in sev-eral cases. However, these were not similar to the ones reported in
a study of Martin et al. (2012), which detected IBP at the highest
average concentration with above 1000 lg kg1. This study is the
first report on the presence of ALP in sludge.
3.3. Mass balance analysis
Mass flux and mass balance analysis of the individual PhACs
were conducted to assess their potential removal mechanisms in
the WWTP (Table 2). The input mass load for all target PhACs
was7627 g d1. For the filtered disinfection effluent, the total mass
load was 1371 g d1, and the mass loads of the individual PhACs
varied from 1.74 g d1 (GFB) to 688.75 g d1 (SMZ). The sorbed
amounts were negligible because of the relatively low concentra-tions of the suspended solids in the effluent. The total mass load
of the selected PhACs in the dewatered sludge (without further
digestion) was 121 g d1. QAs and MAs were the predominant
residual PhACs, accounting for more than 80% of all PhACs ana-
lyzed in the dewatered sludge.
Fig. 4 shows the mass loads of the PhACs discharged from the
plant with the effluent and the treated sludge normalized on the
influent mass loads. The overall removal efficiencies (Roverall) of
the target PhACs ranged from ‘‘negative removal’’ to almost com-
plete removal. ACM and IBP, which have excellent aqueous re-
moval efficiencies (Raqueous) (i.e., 99.87% and 94.53%, respectively),
were removed and did not accumulate in the sludge, thereby sug-
gesting that the aqueous removal of ACM and IBP can be attributed
to the degradation process. Table 2 shows that the aqueous re-
moval of ACM and IBP mainly occurred in the biological treatmentunit, indicating that ACM and IBP are mainly biodegraded in the
WWTP. Similar removals of the two compounds from the aqueous
phase with conventional treatment processes were observed in
other studies (Gomez et al., 2007; Kasprzyk-Hordern et al., 2009;
Zorita et al., 2009; Behera et al., 2011). By contrast, MTP
(32.22%) and CBZ (15.18%) showed ‘‘negative removal rates’’
(referring to an increase in the concentration of a detected PhACs
during treatment), which might be explained by the sampling pro-
tocols (Ort et al., 2010) and the formation of unmeasured products
of human metabolism and pharmaceutical conjugate that pass
through the plant and become converted back to the parent com-
pounds (Miao et al., 2005; Gobel et al., 2007; Radjenovic et al.,
2009). The results concerning the persistence of MTP and CBZ
match with those from previous studies (Miao et al., 2005; Gaoet al., 2012b). No significant overall removal was observed for
the antibiotics SDZ, ROX and TMP. The incomplete removal of these
PhACs during conventional treatment has been previously reported
as well (Roberts and Thomas, 2006; Gobel et al., 2007; Ghosh et al.,
2009; Gao et al., 2012b; Zhou et al., 2013). ATT, DCF, GFB and ALP
were detected with concentrations below or close to the LOQs.
Thus, no reliable conclusion can be made regarding the behavior
and fate of these compounds. The calculated Roverall for the other
PhACs ranged from 39.11% to 80.02%. QAs, AZM and SVT were
found to be apparently distributed in the sludge, accounting for
9.35–26.96% of the initial loadings. This finding indicates that sorp-
tion via sludge is a relevant removal pathway for these compounds
in the WWTP. For the other selected PhACs, the mass proportions
in the sludge only accounted for a minor part in the removal. Pre-vious studies have reported that QAs are easily adsorbed onto
Table 2
Mass flux of the target PhACs at different treatment units.
PhACs Mass flux (g d1) Discharged with effluent (%) Sorbed to sludge (%) Removed
(%)Influent Primary
treatment
Secondary
treatment
Disinfection Dewatered
sludge
SDZ 137.94 125.84 98.49 92.98 0.45 67.41 0.33 32.27
SM1 90.14 76.3 23.88 23.92 0.12 26.53 0.13 73.34
SMZ 1761.22 1083.75 704.15 688.75 1.19 39.11 0.07 60.83
TMP 46.42 48.89 28.4 31.56 1.04 67.98 2.25 29.77
OFX 207.54 163.58 61.02 34.76 19.4 16.75 9.35 73.9
NOR 121.79 91.1 17.77 18.21 13.54 14.95 11.12 73.92
MOX 11.93 11.2 4.17 3.93 2.02 32.95 16.9 50.15
ERY 152.54 154.76 87.24 91.8 1.09 60.18 0.71 39.11
ROX 242.38 231.4 184.2 208.53 5.5 86.04 2.27 11.7
AZM 217.47 157.66 53.14 48.91 58.63 22.49 26.96 50.55
IBP 160.82 138.4 11.45 8.79 0 5.47 0 94.53
ACM 4267.03 4046.35 3.99 5.44 0 0.13 0 99.87
BZB 75 89.36 38.16 41.53 0.24 55.37 0.33 44.3
CA 16.51 12.45 10.64 9.7 0.17 58.76 1.01 40.23
GFB 8.7 4.92 2.19 1.74 0 19.98 0 80.02
MTP 30.11 32.94 36.01 38.84 0.67 129.01 2.21 31.22
SVT 70.52 75.7 10.83 11.86 15.91 16.81 22.56 60.62
CBZ 8.7 9.48 10.01 9.92 0.1 113.98 1.2 15.18
166 Q. Yan et al./ Chemosphere 99 (2014) 160–170
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sludge and that adsorption is an important aqueous removal mech-
anism (Lindberg et al., 2005; Jia et al., 2012; Zhou et al., 2013).
Table 2 shows the distribution of the detected PhACs across the
treatment plant. The obtained results showed that NOR, OFX and
AZM decreased by 21.18%, 25.20% and 27.5%, respectively, during
the primary treatment step, whereas the other target PhACs pre-
sented either no obvious reduction or a slight increase. A signifi-
cant reduction (>40%) was observed for most of target PhACs
during the biological treatment, although the elimination was
not apparent or was negative for MTP, SDZ, ROX and CBZ in the
WWTP. The removal of PhACs during the disinfection treatment
step was negligible with the exception of OFX.
3.4. Estimation of PhACs consumption
Table 3 shows pharmacokinetics and the results for the back-
estimated usage of the target PhACs in Chongqing by using Eq.(6). The back-estimated usage of OFX was 0.416 ton year1 which
was in good agreement with its consumption (0.574 ton) reported
in 2012 with a model agreement ratio of 72.5%. Thus, there must be
other environmental sinks for this antibiotic, which may underes-
timate the real situation and consequently result in a usage ratio
lower than 100%. Additionally, not all PhACs that are sold are actu-
ally consumed by the user because of numerous reasons (Ruhoy
and Daughton, 2008). No reliable quantitative data on the fraction
of unused medication could be found in Chongqing. This difference
also may be attributed to environmental loss factors. In the present
situation, it is difficult to verify the correct for other PhACs due to
the absence of consumption data. Further verification of the back-
estimated usage of target PhACs should be conducted if enough
information is available. However, the good agreement ratio ob-tained for OFX proved the feasibility of the back-estimated usage
using the established back-calculated model in other areas where
pharmaceutical consumption records are not available.
As a way of ground truthing these values, we can calculate the
medication consumption in hospitals for AZM and OFX in Chongq-
ing main urban district, based on the total prescription amounts
provided by Shanghai food and drug administration (Table S6 of
Supplementary Information). The total sales of OFX and AZM were
294 kg and 179 kg in 2012 in Chongqing main urban district
(Table S6 in the Supplementary Information). The annual con-
sumption of OFX and AZM determined using back-calculation
(Eq. (6)) were 416 and 525 kg (Table 3). In partial explanation for
this discrepancy between back-calculation and prescription
amounts, it is known that other medical sources, primarily numer-ous private practitioners, veterinary consumption and amounts ap-
plied by other modes of administration including ocular treatmentand topical use, are excluded.
To make a global comparison of PhAC usage, all back-estimated
usage values in this study and the official usage values reported in
previous literatures were normalized by the total population (in
g year1 person1) in each studied region. Different consumption
patterns were found by contrasting the local consumption with
the available levels calculated in developed countries. As shown
in Table 4, all antibiotics have higher levels of usage in Chongqing
compared with Australia by 1–3 orders of magnitude ( Watkinson
et al., 2007). Generally, the annual per capita consumption of the
antibiotics were comparable to or even higher than those reported
in developed countries, while the case of other target PhACs was
opposite with exception of CA. The estimated annual per capita
consumption of OFX, NOR, ERY and ROX in this study was lowerthan those reported in Hong Kong of China, whereas SMZ
-40%
-20%
0%
20%
40%
60%
80%
100%
120%
140%
160%
S D Z
S M 1
S M Z
T M P
O F X
N O R
M O X
E R Y
R O X
A Z M
I B P
D C F
A C M
B Z B
C A
G F B
M T P
A T T
S V T
C B Z
Removed Sorbed to sludge Discharged with effluent
Fig. 4. Mass loads of the selected PhACs normalized to the influent mass load: fraction discharged with effluent, sorbed to sludge and removed during treatment.
Table 3
Pharmacokinetics and corresponding back-calculated consumption in Chongqing.
PhACs C influent Pharmacokinetics (%) PredictedPhACs consumption
(tonyear1) in ChongqingAbsorption Excreted
unchanged
SDZ 229.9 n.a. n.a. 0.443
SM1 150.23 n.a. n.a. 0.426
SMZ 2935.37 70b,c 20a 6.432
TMP 77.37 95b,c 45b,c 0.156
OFX 345.9 99b,c 80b,c 0.416
NOR 202.98 40b,c 60b,c 0.233
MOX 19.88 90b,c 45b,c 0.038
ERY 254.24 45b,c 25b,c 0.370
ROX 403.96 50d 66d 0.469
AZM 362.45 38b,c 12b,c 0.525
IBP 268.03 85b,c 30a 0.638
DCF 4.24 55b,c 16a 0.008
ACM 7111.72 95b,c 6b,c 64.080
BZB 125 100b,c 51a 0.236CA 27.52 100b,c 1b,c 2.653
GFB 14.5 n.a. 76e 0.018
MTP 50.18 n.a. 11a 0.440
ATT 1.48 70b,c 1b,c 0.005
SVT 117.53 70b,c 1f 0.369
CBZ 14.5 90b,c 3e 0.110
a Lienert et al. (2007).b WebMD (2011).c MERCK (2011).d Carballa et al. (2008).e Sui et al. (2012).f Kosjek et al. (2009).
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(1.187 g year1 person1) is relatively higher than in Hong Kong
(0.186 g year1 person1) (Leung et al., 2012).
As the present study used grab sampling data from one WWTP
to back-estimate the overall pharmaceutical consumption in Chon-
gqing of China, these results needs to be used with caution and
should be seen as a pilot study aimed at providing preliminary
consumption data, and should be improved through long-termmonitoring work. However, it should be noted that current back-
calculation based on influent concentration in WWTPs can supply
more consumption information due to the wide coverage of multi-
ple pharmaceutical sources and can work as a reference for
improving current statutory regulation on pharmaceutical
consumption.
4. Conclusions
As a result of the incomplete removal of the target PhACs during
wastewater treatment, all the 21 analyzed PhACs were detected in
the effluent. Apart from IBP, GFB and ACM, the selected PhACs also
were found in the dewatered sludge. SAs and MAs were the pre-
dominant PhACs in the effluents, while in the sludge MAs andQAs were dominant with much higher concentrations than the
other groups of the target PhACs. The sum of the average daily
mass load for all selected PhACs is estimatedto be 4.95 mgd1 per-
son1 in the influent, and the average mass load per capita to the
receiving water and sludge were 889.94 and 78.57 lg d1 per-
son1, respectively. The highest mass loads in influent, effluent
and sludge were ACM (2771 lg d1 person1), SMZ (477 lg d1 -
person1) and AZM (39 lg d1 person1), respectively. Generally,
the average influent mass loads and back-calculated annual per ca-
pita consumption of the selected antibiotics were comparable to or
higher than those reported in developed countries, while the case
of other target PhACs was opposite. Mass balance analysis reveals
the different behavior and fate of various pharmaceutical classes
and that their removal is mainly attributed to their biodegradationstep. QAs, AZM and SVT are apparently distributed in the sludge,
accounting for 9.35–26.96% of the initial loadings. This finding
indicates that sorption via sludge is a relevant removal pathway
for these compounds in the WWTP. For the other selected PhACs,
the sorption onto sludge is negligible. In summary, the results of
this research will help improve our understanding on the behavior,
fate and transport of PhACs in WWTPs. The analysis of post-thera-
peutic residual concentrations in the influent after human admin-istration can be an effective method to back-calculate PhAC usage
by considering the pharmacokinetics of PhACs in humans. The
back-calculation can supply more consumption information and,
if necessary, can work as a reference for improving current statu-
tory regulation on pharmaceutical consumption. However, the
back-estimation of PhAC usage requires further verification if en-
ough information is available.
Acknowledgements
We would like to thank the Chongqing Water Group for their
help during sample collection. This study was funded by the
National Science and Technology Supporting Program (Grant Nos.
2012BAJ25B09 and 2012BAJ25B06), and the Chongqing Scienceand Technology Commission (Grant Nos. CSTC2012jjA0775 and
CSTC2012ggB20001).
Appendix A. Supplementary material
Supplementary data associated with this article can be
found, in the online version, at http://dx.doi.org/10.1016/
j.chemosphere.2013.10.062.
References
Adams, M., Tavakoli, H., 2006. Gatifloxacin-induced hallucinations in a 19-year-old
man. Psychosomatics 47, 360.
Alder, A.C., Bruchet, A., Carballa, M., Clara, M., Joss, A., Loffler, D., McArdell, C.S.,
Miksch, K., Omil, F., Tuhkanen, T., Ternes, T.A., 2006. Consumption andoccurrence. In: Ternes, T.A., Joss, A. (Eds.), Human Pharmaceuticals, Hormones
Table 4
Global comparison consumption of the selected PhACs.
Studied region(g year1 person1) SDZ SM1 SMZ TMP OFX NOR MOX ERY ROX AZM IBP DCF ACM BZB CA GFB MTP ATT SVT CBZ
Chongqing in this study 0.082 0.079 1.187 0.029 0.077 0.043 0.007 0.068 0.087 0.097 0.118 0.001 11.823 0.044 0.489 0.003 0.081 0.001 0.068 0.020
Spaina 0.231 0.002 0.116 0.092 0.039 0.001 0.104 4.647 0.370 31.054 0.133 0.007 0.138 0.287 0.438
Netherlandsb 0.440 1.630 0.330
Hong Kongc 0.186 0.031 0.191 0.281 1.888 0.228
Australiad 0.004 0.002 0.002 0.019 0.002
Swedene 0.104 0.104 0.157 0.014Switzerlandf 0.349 0.070 0.030 0.020
Germanyg 0.065 0.015 0.009 0.304 0.095 0.048 0.079 0.101
Switzerlandg 0.315 0.071 0.020 3.078 0.934 0.216 0.522 0.858
Franceg 0.299 0.352 0.071 0.997 0.387 0.060 0.215 0.570
Austriah 0.120 0.048 0.837 0.768 0.559 0.792
Finlandi 0.077 11.610 0.154 0.115 1.019
Francei 0.383 0.159 2.841 0.255 0.590 0.602
Germanyi 0.571 0.075 1.553 0.595 0.316 0.631 0.947
Polandi 0.829 0.053 1.518 0.541 0.021 1.072
Swedeni 0.160 0.002 7.864 0.376 0.067 0.820
Spain j 0.294 0.188 0.009 6.391 0.748 0.093 0.053 0.463
Switzerlandk 0.352 0.020 2.153 0.533 0.216 0.557
a Ortiz de Garcia et al. (2013).b Oosterhuis et al. (2013).c Leung et al. (2012).d Watkinson et al. (2007).e Lindberg et al. (2005).f Gobel et al. (2005).
g ter Laak et al. (2010).h Kreuzinger et al. (2004).i Alder et al. (2006).
j Carballa et al. (2008).k Huber et al. (2004).
168 Q. Yan et al./ Chemosphere 99 (2014) 160–170
http://dx.doi.org/10.1016/j.chemosphere.2013.10.062http://dx.doi.org/10.1016/j.chemosphere.2013.10.062http://refhub.elsevier.com/S0045-6535(13)01496-3/h0005http://refhub.elsevier.com/S0045-6535(13)01496-3/h0005http://refhub.elsevier.com/S0045-6535(13)01496-3/h0005http://refhub.elsevier.com/S0045-6535(13)01496-3/h0010http://refhub.elsevier.com/S0045-6535(13)01496-3/h0010http://refhub.elsevier.com/S0045-6535(13)01496-3/h0010http://refhub.elsevier.com/S0045-6535(13)01496-3/h0010http://refhub.elsevier.com/S0045-6535(13)01496-3/h0010http://refhub.elsevier.com/S0045-6535(13)01496-3/h0010http://refhub.elsevier.com/S0045-6535(13)01496-3/h0005http://refhub.elsevier.com/S0045-6535(13)01496-3/h0005http://dx.doi.org/10.1016/j.chemosphere.2013.10.062http://dx.doi.org/10.1016/j.chemosphere.2013.10.062
-
8/8/2019 2014. Southwest China
10/11
and Fragrances: The Challenge of Micropollutants in Urban Water Management.
IWA Publishing, London, UK, pp. 15–54.
Aydin, E., Talinli, I., 2013. Analysis, occurrence and fate of commonly used
pharmaceuticals and hormones in the Buyukcekmece Watershed, Turkey.
Chemosphere 90, 2004–2012.
Behera, S.K., Kim, H.W., Oh, J.E., Park, H.S., 2011. Occurrence and removal of
antibiotics, hormones and several other pharmaceuticals in wastewater
treatment plants of the largest industrial city of Korea. The Science of the
Total Environment 409, 4351–4360.
Besse, J.-P., Kausch-Barreto, C., Garric, J., 2008. Exposure assessment of
pharmaceuticals and their metabolites in the aquatic environment:application to the french situation and preliminary prioritization. Human and
Ecological Risk Assessment: An International Journal 14, 665–695.
Carballa, M., Omil, F., Lema, J.M., 2008. Comparison of predicted and measured
concentrations of selected pharmaceuticals, fragrances and hormones in
Spanish sewage. Chemosphere 72, 1118–1123.
Castiglioni, S., Bagnati, R., Fanelli, R., Pomati, F., Calamari, D., Zuccato, E., 2006.
Removal of pharmaceuticals in sewage treatment plants in Italy. Environmental
Science & Technology 40, 357–363.
Dirany, A., Efremova Aaron, S., Oturan, N., Sires, I., Oturan, M.A., Aaron, J.J., 2011.
Study of the toxicity of sulfamethoxazole and its degradation products in water
by a bioluminescence method during application of the electro-Fenton
treatment. Analytical and Bioanalytical Chemistry 400, 353–360.
Gao, L., Shi, Y., Li, W., Niu, H., Liu, J., Cai, Y., 2012a. Occurrence of antibiotics in eight
sewage treatment plants in Beijing, China. Chemosphere 86, 665–671.
Gao, P., Ding, Y., Li, H., Xagoraraki, I., 2012b. Occurrence of pharmaceuticals in a
municipal wastewater treatment plant: mass balance and removal processes.
Chemosphere 88, 17–24.
Ghosh, G.C., Okuda, T., Yamashita, N., Tanaka, H., 2009. Occurrence and
elimination of antibiotics at four sewage treatment plants in Japan and
their effects on bacterial ammonia oxidation. Water Science and Technology
59, 779–786.
Gobel, A., Thomsen, A., McArdell, C.S., Joss, A., Giger, W., 2005. Occurrence and
sorption behavior of sulfonamides, macrolides, and trimethoprim in activated
sludge treatment. Environmental Science & Technology 39, 3981–3989.
Gobel, A., McArdell, C.S., Joss, A., Siegrist, H., Giger, W., 2007. Fate of sulfonamides,
macrolides, and trimethoprim in different wastewater treatment technologies.
The Science of the Total Environment 372, 361–371.
Golet, E.M., Strehler, A., Alder, A.C., Giger, W., 2002. Determination of
fluoroquinolone antibacterial agents in sewage sludge and sludge-treated soil
using accelerated solvent extraction followed by solid-phase extraction.
Analytical Chemistry 74, 5455–5462.
Golet, E.M., Xifra, I., Siegrist, H., Alder, A.C., Giger, W., 2003. Environmental exposure
assessment of fluoroquinolone antibacterial agents from sewage to soil.
Environmental Science & Technology 37, 3243–3249.
Gomez, M.J., Martinez Bueno, M.J., Lacorte, S., Fernandez-Alba, A.R., Aguera, A.,
2007. Pilot survey monitoring pharmaceuticals and related compounds in a
sewage treatment plant located on the Mediterranean coast. Chemosphere 66,993–1002.
Gracia-Lor, E., Sancho, J.V., Serrano, R., Hernandez, F., 2012. Occurrence and removal
of pharmaceuticals in wastewater treatment plants at the Spanish
Mediterranean area of Valencia. Chemosphere 87, 453–462.
Gulkowska, A., He, Y., So, M.K., Yeung, L.W., Leung, H.W., Giesy, J.P., Lam, P.K.,
Martin, M., Richardson, B.J., 2007. The occurrence of selected antibiotics in Hong
Kong coastal waters. Marine Pollution Bulletin 54, 1287–1293.
Gulkowska, A., Leung, H.W., So, M.K., Taniyasu, S., Yamashita, N., Yeung, L.W.,
Richardson, B.J., Lei, A.P., Giesy, J.P., Lam, P.K., 2008. Removal of antibiotics from
wastewater by sewage treatment facilities in Hong Kong and Shenzhen, China.
Water Research 42, 395–403.
Huber, M., 2004. Elimination of Pharmaceuticals during Oxidative Treatment of
Drinking Water and Wastewater. Ph.D. Thesis. Swiss Federal Institute of
Technology, Zurich.
Jelic, A., Gros, M., Ginebreda, A., Cespedes-Sanchez, R., Ventura, F., Petrovic, M.,
Barcelo, D., 2011. Occurrence, partition and removal of pharmaceuticals in
sewage water and sludge during wastewater treatment. Water Research 45,
1165–1176.
Jelic, A., Fatone, F., Di Fabio, S., Petrovic, M., Cecchi, F., Barcelo, D., 2012. Tracingpharmaceuticals in a municipal plant for integrated wastewater and organic
solid waste treatment. The Science of the Total Environment 433, 352–361.
Jia, A., Wan, Y., Xiao, Y., Hu, J., 2012. Occurrence and fate of quinolone and
fluoroquinolone antibiotics in a municipal sewage treatment plant. Water
Research 46, 387–394.
Jones, O.A., Voulvoulis, N., Lester, J.N., 2002. Aquatic environmental assessment of
the top 25 English prescription pharmaceuticals. Water Research 36, 5013–
5022.
Karthikeyan, K.G., Meyer, M.T., 2006. Occurrence of antibiotics in wastewater
treatment facilities in Wisconsin, USA. The Science of the Total Environment
361, 196–207.
Kasprzyk-Hordern, B., Dinsdale, R.M., Guwy, A.J., 2009. Illicit drugs and
pharmaceuticals in the environment–forensic applications of environmental
data. Part 1: Estimation of the usage of drugs in local communities.
Environmental Pollution 157, 1773–1777.
Kosjek, T., Andersen, H.R., Kompare, B., Ledin, A., Heath, E., 2009. Fate of
carbamazepine during water treatment. Environmental Science & Technology
43, 6256–6261.
Kreuzinger, N., Clara, M., Strenn, B., Kroiss, H., 2004. Relevance of the sludge
retention time (SRT) as design criteria for wastewater treatment plants for the
removal of endocrine disruptors and pharmaceuticals from wastewater. Water
Science and Technology: A Journal of the International Association on Water
Pollution Research 50, 149–156.
Leung, H.W., Minh, T.B., Murphy, M.B., Lam, J.C., So, M.K., Martin, M., Lam, P.K.,
Richardson, B.J., 2012. Distribution, fate and risk assessment of antibiotics in
sewage treatment plants in Hong Kong, South China. Environment International
42, 1–9.
Li, B., Zhang, T., 2011. Mass flows and removal of antibiotics in two municipal
wastewater treatment plants. Chemosphere 83, 1284–1289.Lienert, J., Gudel, K., Escher, B.I., 2007. Screening method for ecotoxicological hazard
assessment of 42 pharmaceuticals considering human metabolism and
excretory routes. Environmental Science & Technology 41, 4471–4478.
Lindberg, R.H., Wennberg, P., Johansson, M.I., Tysklind, M., Andersson, B.A., 2005.
Screening of human antibiotic substances and determination of weekly mass
flows in five sewage treatment plants in Sweden. Environmental Science &
Technology 39, 3421–3429.
Martin, J., Camacho-Munoz, D., Santos, J.L., Aparicio, I., Alonso, E., 2012. Occurrence
of pharmaceutical compounds in wastewater and sludge from wastewater
treatment plants: removal and ecotoxicological impact of wastewater
discharges and sludge disposal. Journal of Hazardous Materials 239–240, 40–
47.
Martinez, J.L., 2008. Antibiotics and antibiotic resistance genes in natural
environments. Science 321, 365–367.
MERCK, 2011. Manual MERCK. (accessed 08.12.11).
Miao, X.S., Yang, J.J., Metcalfe, C.D., 2005. Carbamazepine and its metabolites in
wastewater and in biosolids in a municipal wastewater treatment plant.
Environmental Science & Technology 39, 7469–7475.
Nieto, A., Borrull, F., Pocurull, E., Marcé, R.M., 2010. Occurrence of pharmaceuticals
and hormones in sewage sludge. Environmental Toxicology and Chemistry 29,
1484–1489.
Oosterhuis, M., Sacher, F., Ter Laak, T.L., 2013. Prediction of concentration levels of
metformin and other high consumption pharmaceuticals in wastewater and
regional surface water based on sales data. The Science of the Total
Environment 442, 380–388.
Ort, C., Lawrence, M.G., Reungoat, J., Mueller, J.F., 2010. Sampling for PPCPs in
wastewater systems: comparison of different sampling modes and optimization
strategies. Environmental Science & Technology 44, 6289–6296.
Ortiz de Garcia, S., Pinto Pinto, G., Garcia Encina, P., Irusta Mata, R., 2013.
Consumption and occurrence of pharmaceutical and personal care products in
the aquatic environment in Spain. The Science of the Total Environment 444,
451–465.
Pomati, F., Cotsapas, C.J., Castiglioni, S., Zuccato, E., Calamari, D., 2007. Gene
expression profiles in zebrafish (Danio rerio) liver cells exposed to a mixture of pharmaceuticals at environmentally relevant concentrations. Chemosphere 70,
65–73.Radjenovic, J., Petrovic, M., Barcelo, D., 2009. Fate and distribution of
pharmaceuticals in wastewater and sewage sludge of the conventional
activated sludge (CAS) and advanced membrane bioreactor (MBR) treatment.
Water Research 43, 831–841.
Roberts, P.H., Thomas, K.V., 2006. The occurrence of selected pharmaceuticals in
wastewater effluent and surface waters of the lower Tyne catchment. The
Science of the Total Environment 356, 143–153.
Rowney, N.C., Johnson, A.C., Williams, R.J., 2009. Cytotoxic drugs in drinking water:
a prediction and risk assessment exercise for the Thames catchment in the
United Kingdom. Environmental Toxicology and Chemistry/SETAC 28, 2733–
2743.
Ruhoy, I.S., Daughton, C.G., 2008. Beyond the medicinecabinet:an analysis of where
and why medications accumulate. Environment International 34, 1157–1169.
Samaras, V.G., Stasinakis, A.S., Mamais, D., Thomaidis, N.S., Lekkas, T.D., 2013. Fate
of selected pharmaceuticals and synthetic endocrine disrupting compounds
during wastewater treatment and sludge anaerobic digestion. Journal of
Hazardous Materials 244–245, 259–267.
Scheurer, M., Sacher, F., Brauch, H.J., 2009. Occurrence of the antidiabetic drug
metformin in sewage and surface waters in Germany. Journal of EnvironmentalMonitoring: JEM 11, 1608–1613.
Sui, Q., Huang, J., Deng, S., Yu, G., Fan, Q., 2010. Occurrence and removal of
pharmaceuticals, caffeine and DEET in wastewater treatment plants of Beijing,
China. Water Research 44, 417–426.
Sui, Q., Wang, B., Zhao, W., Huang, J., Yu, G., Deng, S., Qiu, Z., Lu, S., 2012.
Identification of priority pharmaceuticals in the water environment of China.
Chemosphere 89, 280–286.
ter Laak, T.L., van der Aa, M., Houtman, C.J., Stoks, P.G., van Wezel, A.P., 2010.
Relating environmental concentrations of pharmaceuticals to consumption: a
mass balance approach for the river Rhine. Environment International 36, 403–
409.
Ternes, T.A., Herrmann, N., Bonerz, M., Knacker, T., Siegrist, H., Joss, A., 2004. A rapid
method to measure the solid-water distribution coefficient (Kd) for
pharmaceuticals and musk fragrances in sewage sludge. Water Research 38,
4075–4084.
Urase, T., Kikuta, T., 2005. Separate estimation of adsorption and degradation of
pharmaceutical substances and estrogens in the activated sludge process.
Water Research 39, 1289–1300.
Q. Yan et al. / Chemosphere 99 (2014) 160–170 169
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