2014. Southwest China

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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

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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.


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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


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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.

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