Pharmaceuticals and personal care products (PPCPs)

Environment International 59 (2013) 208–224

Contents lists available at SciVerse ScienceDirect

Environment International journal homepage: www.elsevier.com/locate/envint

Review

Pharmaceuticals and personal care products (PPCPs): A review on environmental contamination in China Jin-Lin Liu b, Ming-Hung Wong a,b,⁎ a b

School of Environmental and Resource Sciences, Zhejiang Agricultural and Forestry University, Linan, PR China Croucher Institute for Environmental Sciences, and Department of Biology, Hong Kong Baptist University, Hong Kong, PR China

a r t i c l e

i n f o

Article history: Received 15 February 2013 Accepted 18 June 2013 Available online xxxx Keywords: Pharmaceuticals and personal care products (PPCPs) China Contamination Sewage Surface water

a b s t r a c t Pharmaceuticals and personal care products (PPCPs) which contain diverse organic groups, such as antibiotics, hormones, antimicrobial agents, synthetic musks, etc., have raised significant concerns in recently years for their persistent input and potential threat to ecological environment and human health. China is a large country with high production and consumption of PPCPs for its economic development and population growth in recent years. This may result in PPCP contamination in different environmental media of China. This review summarizes the current contamination status of different environment media, including sewage, surface water, sludge, sediments, soil, and wild animals, in China by PPCPs. The human body burden and adverse effects derived from PPCPs are also evaluated. Based on this review, it has been concluded that more contamination information of aquatic environment and wildlife as well as human body burden of PPCPs in different areas of China is urgent. Studies about their environmental behavior and control technologies need to be conducted, and acute and chronic toxicities of different PPCP groups should be investigated for assessing their potential ecological and health risks. © 2013 Published by Elsevier Ltd.

Contents 1.

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Introduction . . . . . . . . . . . . . . . . . . . . . . . . . 1.1. Classification of PPCPs . . . . . . . . . . . . . . . . . 1.2. Sources, fates, and persistence of PPCPs . . . . . . . . . PPCP production and usage in China . . . . . . . . . . . . . Occurrence of PPCPs in different environmental media of China . 3.1. PPCPs in sewage and sludge . . . . . . . . . . . . . . 3.1.1. PPCPs in sewage . . . . . . . . . . . . . . . 3.1.2. Fates of PPCPs in STPs . . . . . . . . . . . . . 3.1.3. Removal efficiencies of PPCPs in STPs . . . . . 3.1.4. PPCPs in sludge . . . . . . . . . . . . . . . . 3.2. PPCPs in surface water and sediments . . . . . . . . . 3.2.1. PPCPs in surface water . . . . . . . . . . . . 3.2.2. Fates of PPCPs in surface water . . . . . . . . 3.2.3. PPCPs in sediments . . . . . . . . . . . . . . 3.2.4. Removal efficiencies of PPCPs in water treatment 3.3. PPCPs in soil . . . . . . . . . . . . . . . . . . . . . 3.4. PPCPs in wild animals . . . . . . . . . . . . . . . . . Human exposure and body loading of PPCPs . . . . . . . . . . 4.1. Human exposure . . . . . . . . . . . . . . . . . . . 4.2. Body loading . . . . . . . . . . . . . . . . . . . . . Toxicity and risk assessments of PPCPs . . . . . . . . . . . . 5.1. Toxicity . . . . . . . . . . . . . . . . . . . . . . . . 5.2. Risk assessments . . . . . . . . . . . . . . . . . . .

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⁎ Corresponding author at: School of Environment and Energy, Peking University, Shenzhen, PR China. Tel.: +852 3411 7746; fax: +852 3411 7743. E-mail address: [email protected] (M.-H. Wong). 0160-4120/$ – see front matter © 2013 Published by Elsevier Ltd. http://dx.doi.org/10.1016/j.envint.2013.06.012

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J.-L. Liu, M.-H. Wong / Environment International 59 (2013) 208–224

6. Regulations and 7. Conclusions . . Acknowledgments . References . . . . .

control strategies of PPCP . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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1. Introduction Pharmaceuticals and personal care products (PPCPs) have received growing attention in recent years as emerging contaminants for their possible threats to aquatic environment and human health. They contain a large and diverse group of organic compounds, including pharmaceutical drugs and integrates of daily personal care products (PCPs), such as soaps, lotions, toothpaste, fragrances, sunscreens, etc., which are widely used in high quantities throughout the world, together with their metabolites and transformation products (Daughton and Ternes, 1999; Kummerer, 2000). As an important group of organic pollutants with intensive studies in recent years, PPCPs have been found to be ubiquitous in the aquatic environment throughout the world (Carballa et al., 2004; Kasprzyk-Hordern et al., 2009; Lishman et al., 2006; Nakada et al., 2006; Tauxe-Wuersch et al., 2005), including in the raw water sources of drinking water treatment plants (Radjenovic et al., 2008; Ternes et al., 2002; Vieno et al., 2007a). China is a large country with high production and consumption of PPCPs owe to its economic development and population growth, which may result in serious pollution by PPCPs. This review aims to summarize the current PPCP contamination status in different environmental media of China and assess their possible threats to ecosystem and human health. 1.1. Classification of PPCPs PPCPs contain diverse groups of organic compounds, such as antibiotics, hormones, anti-inflammatory drugs, antiepileptic drugs, blood

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220 221 221 221

lipid regulators, β-blockers, contrast media, and cytostatic drugs for pharmaceuticals; and antimicrobial agents, synthetic musks, insect repellants, preservatives, and sunscreen UV filters for personal care products (Table 1) (Daughton and Ternes, 1999). Among the pharmaceutical group, antibiotics have received special attention for their wide application in human therapy and livestock agriculture. Persistent exposure of antibiotics can result in the emergence of resistant bacteria strains with public health concerns (Zhang et al., 2009b). Antibiotics contain several subgroups, such as macrolides (e.g. erythromycin, roxithromycin), sulfonamides (e.g. sulfamethoxazole, sulfadimethoxine), and fluoroquinolones (e.g. norfloxacin, ciprofloxacin) (Heberer, 2002). Hormones are another most studied group of pharmaceuticals which are believed to be connected with the endocrine disrupting effects of polluted water bodies (Lai et al., 2002). The most concerned and studied hormones are steroid estrogens, including natural steroid estrogens which are primarily excreted by humans and animals, e.g. estrone (E1), estradiol (E2), estriol (E3), and synthetic steroid estrogens which are used as oral contraceptives, mainly ethinylestradiol (EE2) (Heberer, 2002). Natural steroid estrogens are not actually pharmaceuticals, however, they are usually studied together with synthetic hormones for their endocrine disrupting effects in polluted water. They will be also discussed in the current review. Other pharmaceutical groups include analgesics and anti-inflammatory drugs (such as diclofenac and ibuprofen); antiepileptic drugs (such as carbamazepine and primidone); blood lipid regulators (such as clofibrate and gemfibrozil); β-blockers (such as metoprolol and propanolol); and contrast media (such as iopromide and diatrizoate) (Heberer, 2002). For the groups of personal care products, triclosan and triclocarban are the two typical antimicrobial

Table 1 Classification of PPCPs.

Pharmaceuticals

Subgroups

Representative compounds

Antibiotics

Clarithromycin Erythromycin Sulfamethoxazole Sulfadimethoxine Ciprofloxacin Norfloxacin Chloramphenicol Estrone (E1) Estradiol (E2) Ethinylestradiol (EE2) Diclofenac Ibuprofen Acetaminophen Acetylsalicylic acid Carbamazepine Primidone Clofibrate Gemfibrozil Metoprolol Propanolol Diatrizoate Iopromide Ifosfamide Cyclophosphamide Triclosan Triclocarban Galaxolide (HHCB) Toxalide (AHTN) N,N-diethyl-m-toluamide (DEET) Parabens (alkyl-p-hydroxybenzoates) 2-ethyl-hexyl-4-trimethoxycinnamate (EHMC) 4-methyl-benzilidine-camphor (4MBC)

Hormones

Analgesics and anti-inflammatory drugs

Antiepileptic drugs Blood lipid regulators β-blockers Contrast media Cytostatic drugs Personal Care Products

Antimicrobial agents/Disinfectants Synthetic musks/Fragrances Insect repellants Preservatives Sunscreen UV filters

agents frequently detected in wastewater. The synthetic musks include nitro musks (mainly musk xylene (MX) and musk ketone (MK)) and polycyclic musks (mainly galaxolide (HHCB) and toxalide (AHTN)) with more production and application than the nitro group in recent years. N,N-diethyl-m-toluamide (DEET) is the main active ingredient of insect repellents and regularly detected. Parabens are typical preservatives, and 2-ethyl-hexyl-4-trimethoxycinnamate (EHMC) and 4-methyl-benzilidine-camphor (4MBC) are sunscreen UV filter species (Brausch and Rand, 2011). 1.2. Sources, fates, and persistence of PPCPs

2.0 1.5 1.0 0.5 0.0 2011 1-11

2010

2009

2008

2007

2006

2005

2004

25

(b)

Personal Care Products 20

15

10

5

0 2013 (Estimated)

2012 (Estimated)

2011

2010

2009

2008

China has become a large country with high production and usage volumes of pharmaceuticals as well as a rapid growing rate of personal care product consumption, which may result in significant occurrence of PPCPs in the environment. The production of active pharmaceutical ingredients by China has rapidly increased in recent years, and in fact China has become the largest producer of active pharmaceutical ingredients of the world. As shown in Fig. 1(a), from 2003 to 2011, the production of active pharmaceutical ingredients by China has increased two times, and in 2011 about two million tonnes of pharmaceuticals were produced. The pharmaceutical production by China can account for more than 20% of the total production volume of the world (SERI, 2012). More than 1500 kinds of active pharmaceutical ingredients were produced, and more than 6900 pharmaceutical-manufacturing companies were registered in 2007 (Zhou et al., 2010). China also has the highest production of several antibiotic species, such as penicillin and terramycin (SCIO, 2008). Besides the huge pharmaceutical production volume, the consumption rate is also remarkable, especially for the severe antibiotic abuse in current China. The average usage of antibiotics by Chinese is 10 times more than the usage by Americans (CAST, 2008). Around 75% of the patients and 80% of the inpatients with seasonal influenza are prescribed antibiotics (Heddini et al., 2009; Zheng and Zhou, 2007). For the treatment of animal diseases and promotion of livestock growth, antibiotics are also widely used in livestock agriculture. The antibiotic usage volume in livestock farms is about 97,000 tonnes and can account for 46% of the total volume every year, and most antibiotic species used

(a)

Pharmaceuticals

2.5

2007

2. PPCP production and usage in China

3.0

2003

The main source of PPCP infusion into the environment is through sewage treatment plants (STPs) (Daughton and Ternes, 1999). The presence of PPCPs in wastewater treatment plants had been reported in different countries all over the world, mostly in the levels of ng L−1 to μg L−1, such as USA (Boyd et al., 2004), United Kingdom (Ashton et al., 2004), Spain (Carballa et al., 2004), Finland (Lindqvist et al., 2005), and Japan (Nakada et al., 2006). Studies also reveal that the removal efficiency of PPCPs by the conventional wastewater treatment processes (flocculation, sedimentation, and active sludge treatment) is limited (Castiglioni et al., 2006; Lishman et al., 2006; Paxeus, 2004; Santos et al., 2007). After discharging from STPs, PPCPs in sewage would cause subsequent contamination to the receiving water bodies. For their polar properties, the fraction which enters into the atmospheric environment will be limited. Their distribution will primarily occur in the aquatic environment. PPCPs may also be adsorbed onto the active sludge in STPs and then introduced into the environment through sludge land application (Daughton and Ternes, 1999). As a group of novel emerging contaminants, PPCPs show varied properties with the conventional persistent organic pollutants whose sources may have been banned or limited. The input of PPCPs with STPs as the main source is perpetual, resulting in a steady-state concentration in aquatic systems, which has been described as “pseudo-persistent” (Daughton, 2003). Persistent exposure by PPCPs even at low concentration levels can be significant.

Active pharmaceutical ingredient production (million tonnes)


Market scale (billion USD)

210

Fig. 1. (a) Active pharmaceutical ingredient production of China (2003–2011) (SERI, 2012). (b) Personal care product market scale of China (2007–2013) (BosiData, 2011).

in human therapy have been applied in livestock agriculture (Heddini et al., 2009; Li and Wang, 2009; Zheng and Zhou, 2007). The emergence or persistence of resistant bacteria strains derived from antibiotic abuse may become a serious public health crisis in the future. China is also among the top three countries with the largest personal care product consumption, together with America and Japan (ChinaIRN, 2012). China possesses the fastest growing rate of personal care product market in the world, which will reach 8% between 2010 and 2013. In 2013, the total industry value will reach USD21.3 billion, accounting for about 10% of the world (Fig. 1(b)) (BosiData, 2011; ChinaIRN, 2012). 3. Occurrence of PPCPs in different environmental media of China 3.1. PPCPs in sewage and sludge Sewage treatment plants are considered as the main source of PPCPs. Occurrence of PPCPs in STPs receives great attention, from which the information about environmental loads of PPCPs can be obtained. Their fate in STPs also matters. Whether or not PPCPs can be removed by the wastewater treatment processes determines their final input into the receiving aquatic environment. A number of studies in China focused on the occurrence and fate of PPCPs in STPs, including sewage (Table 2) and sludge. Most investigations were carried out in several hotspot study areas, such as the Pearl River Delta, Yangtze River Delta, and Beijing area, due to their dense population and rapid socio-economic development. 3.1.1. PPCPs in sewage 3.1.1.1. Antibiotics. As the most concerned group of pharmaceuticals, antibiotics had been found ubiquitously in STP sewage of China,


211

Table 2 Occurrence of PPCPs in sewage of STPs. Location

Chemical

Concentration

Media

Reference

Guangzhou

Antibiotics Antibiotics

Guangzhou Guangzhou and Hong Kong

Antifungal drugs Antibiotics

Hong Kong Hong Kong

Antibiotics Antibiotics

Hong Kong

Antibiotics

Shenzhen Jiangsu Hangzhou

Antibiotics Antibiotics Antibiotics

Chongqing

Antibiotics

Beijing

Antibiotics

Beijing

Antibiotics

Beijing

Antibiotics

Beijing Guangzhou

Antibiotics Hormones

Guangzhou

Hormones

Wuhan

Hormones

Beijing Beijing Beijing

Hormones Hormones Hormones

Beijing

Hormones

Beijing

Hormones

Beijing

Hormones

Tianjin

Hormones

Guangzhou Shanghai Beijing

Pharmaceuticals Carbamazepine Pharmaceuticals

Beijing

Pharmaceuticals

Beijing

Pharmaceuticals

Taiwan Guangzhou Guangzhou

Pharmaceuticals Antimicrobial agents Synthetic musks

Guangzhou

Synthetic musks

Shanghai

Synthetic musks

Beijing

Synthetic musks

Beijing

Insect repellents

Tianjin

UV filters

Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent and effluent Sewage influent Sewage effluent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Livestock sewage Sewage influent Sewage effluent Hospital sewage Swine nursery sewage Slaughter house sewage Municipal sewage Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Swine sewage Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage of a controceptives producing factory Sewage influent and effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage effluent Sewage influent and effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent of a cosmetic plant Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent Sewage influent Sewage effluent

Peng et al. (2006)

Guangzhou

1730–7910 ng L−1 ND−9460 ng L−1 b100–5597 ng L−1 b35–1920 ng L−1 1–1834 ng L−1 10–1978 ng L−1 9–2054 ng L−1 ND−7870 ng L−1 3.2–1718 ng L−1 1.3–1176 ng L−1 b4–2900 ng L−1 b4–1800 ng L−1 29–1100 ng L−1 Max. 211000 ng L−1 108–1405 ng L−1 54–429 ng L−1 ND−4240 ng L−1 7–157 ng L−1 11–825 ng L−1 18–2020 ng L−1 Max. 3100 ng L−1 Max. 1200 ng L−1 ND−2900 ng L−1 ND−4000 ng L−1 28–1208 ng L−1 6–503 ng L−1 600–32700 ng L−1 1.2–1368 ng L−1 1.0–23.1 ng L−1 2.4–621 ng L−1 b0.4–77.3 ng L−1 4.8–82.4 ng L−1 ND-31.5 ng L−1 73–155 ng L−1 ND−4100 ng L−1 ND−760 ng L−1 ND−74.2 ng L−1 22.4–911.5 ng L−1 ND−253.8 ng L−1 ND−3700 ng L−1 ND−12 ng L−1 b0.08–120 ng L−1 b0.02–1.9 ng L−1 b0.4–8.1 ng L−1 b0.4–3.4 ng L−1 b0.7–3239 ng L−1 230–1110 ng L−1 38–89500 ng L−1 27–2700 ng L−1 4.4–6600 ng L−1 2.2–320 ng L−1 ND−11400 ng L−1 ND−1020 ng L−1 ND−30967 ng L−1 10.9–342 ng L−1 ND−3190 ng L−1 b10–2050 ng L−1 3730–595480 ng L−1 500–33540 ng L−1 b4–3430 ng L−1 b4–336 ng L−1 111.9–3003.8 ng L−1 47.3–1258.3 ng L−1 124–1520 ng L−1 21.6–235 ng L−1 34–2128 ng L−1 21–1287 ng L−1

ND−19401 ng L−1 ND−21278 ng L−1 b3–10025 ng L−1 b3–3052 ng L−1

Sewage Sewage Sewage Sewage

Other counties Korea (Ulsan) U.K. (South Wales)

Antibiotics Antibiotics Antibiotics

influent effluent influent effluent

Peng et al. (2008a) Peng et al. (2012) Xu et al. (2007a) Minh et al. (2009) Li and Zhang (2011) Gulkowska et al. (2008) Gulkowska et al. (2008) Wei et al. (2011) Tong et al. (2011) Chang et al. (2010)

Gao et al. (2012) Chang et al. (2008) Xiao et al. (2008) Ben et al. (2008) Liu et al. (2012b) Liu et al. (2011b) Jin et al. (2008) Cui et al. (2006) Zhou et al. (2011e) Zhou et al. (2011d) Zhou et al. (2009b) Chang et al. (2011) Chang et al. (2007) Wang et al. (2005) Zhao et al. (2010b) Zhou et al. (2011c) Zhou et al. (2010) Sui et al. (2010) Sui et al. (2011) Lin et al. (2010) Zhao et al. (2010a) Zeng et al. (2007) Chen et al. (2007) Zhang et al. (2008) Zhou et al. (2009a) Sui et al. (2011) Li et al. (2007)

Behera et al. (2011) Kasprzyk-Hordern et al. (2009) Lindberg et al. (2005) (continued on next page)

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Table 2 (continued) Location Sweden (5 cities) Other counties Finland (12 cities) U.S. (New Mexico) U.S. (Colorado) Korea (Ulsan) Japan (Tokyo) Portugal (5 cities) Canada (18 cities) Korea (Ulsan) Japan (Tokyo) U.K. (South Wales) Portugal (5 cities) Finland (3 cities) Norway (Tromsø) U.S. (3 cities) Japan (A town in western Japan)

Chemical

Antibiotics Antibiotics Antibiotics Hormones Hormones Hormones Hormones Pharmaceuticals Pharmaceuticals Pharmaceuticals Pharmaceuticals Pharmaceuticals Pharmaceuticals Pharmaceuticals Synthetic musks UV filters

U.K. (South Wales)

Antimicrobial agents Preservatives UV filters

Portugal (5 cities) Norway (Tromsø) U.S. (New York State)

Synthetic musks Triclosan Synthetic musks

Concentration

Media

ND−2480 ng L−1 ND−1340 ng L−1

Sewage influent Sewage effluent

Reference

ND−4230 ng L−1 ND−130 ng L−1 ND−1400 ng L−1 110–310 ng L−1 220–1140 ng L−1 90–320 ng L−1 ND−802 ng L−1 ND−24 ng L−1 13.3–255 ng L−1 0.3–110 ng L−1 103–2484 ng L−1 ND−25 ng L−1 2.4–78 ng L−1 0.2–96 ng L−1 ND−11239 ng L−1 ND−5911 ng L−1 1.6–19400 ng L−1 0.2–968 ng L−1 ND−482687 ng L−1 ND−97616 ng L−1 32–106490 ng L−1 20–43653 ng L−1 30–1350 ng L−1 b5.8–1070 ng L−1 ND−104000 ng L−1 ND−126000 ng L−1 330–43800 ng L−1 ND−72 ng L−1 280–1400 ng L−1 27–660 ng L−1 b8.7–78 ng L−1 2.1–b 8.7 ng L−1 27–65381 ng L−1 13–888 ng L−1 b2–30688 ng L−1 b0.6–155 ng L−1 9–13248 ng L−1 b2–6325 ng L−1 1–11463 ng L−1 1–889 ng L−1 380–430 ng L−1 160–480 ng L−1 304–12700 ng L−1 495–3730 ng L−1

Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage Sewage

influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent influent effluent

Vieno et al. (2007b) Brown et al. (2006) Yang and Carlson (2003) Behera et al. (2011) Nakada et al. (2006) Salgado et al. (2010) Servos et al. (2005) Behera et al. (2011) Nakada et al. (2006) Kasprzyk-Hordern et al. (2009) Salgado et al. (2010) Vieno et al. (2006) Weigel et al. (2004) Thomas and Foster (2005) Nakata and Shinohara (2010)

Kasprzyk-Hordern et al. (2009)

Salgado et al. (2010) Weigel et al. (2004) Reiner et al. (2007a)

⁎ND: not detected.

with concentration up to several μg L−1 for most reviewed cases (ND (not detected)-7910 ng L−1 for influent samples and ND-9460 ng L−1 for effluent samples). In Pearl River Delta cities located in South China (such as Guangzhou, Shenzhen, and Hong Kong), antibiotic species with high detection frequency and concentration included macrolides (e.g. roxithromycin and erythromycin–H2O (the main degradation product of erythromycin)), fluoroquinolones (e.g. ofloxacin and norfloxacin), and sulfonamides (e.g. sulfamethoxazole) (Gulkowska et al., 2008; Xu et al., 2007a). Azole antifungal drugs, such as fluconazole, clotrimazole, econazole, ketoconazole, and miconazole, also occurred in STPs (Huang et al., 2012; Peng et al., 2012). Different cities have different antibiotic composition patterns for sewage influents, e.g. cephalexin was the dominant species in Hong Kong sewage, while cefotaxim prevailed in Shenzhen STPs (Gulkowska et al., 2008). This can be attributed to the regional difference in antibiotic prescription patterns. The environmental loads of antibiotics can be obtained according to the occurrence investigation results. According to the results from four sewage treatment plants in Guangzhou and Hong Kong, environmental loads of different antibiotic species were estimated to range from 0.5 to 828 g per day (Xu et al., 2007a). From seven sewage treatment plants of Hong Kong, the total input of antibiotics into the receiving Victoria Harbor was estimated to be 14.4 kg per day (Minh et al., 2009).

As another developed area located in eastern China, antibiotics were also present in STPs of Yangtze River Delta. Veterinary antibiotics, including oxytetracycline, tetracycline, sulfamethazine, sulfadiazine, and sulfamethoxazole, were monitored in wastewater of 27 animal farms from Jiangsu Province, with the highest concentration up to 211 μg L− 1 (Wei et al., 2011). This is much higher than the level in municipal wastewater, such as the maximum of 1405 ng L− 1 for sewage influent samples in STPs of Hangzhou located in the same region (Tong et al., 2011), implying the importance of livestock wastewater as the source of antibiotic input into the environment. Groups of antibiotics also existed in STPs of Beijing, the capital city of the nation located in North China. Veterinary antibiotics were found with a concentration range of 0.6–32.7 μg L−1 in swine wastewater (Ben et al., 2008). This is the highest antibiotic level found in the sewage of Beijing, exceeding the levels generally up to several μg L−1 for other cases (Chang et al., 2008; Gao et al., 2012; Xiao et al., 2008). This confirms the heavy loads of antibiotics from livestock farms. Generally, the sewage antibiotic levels of China are comparable with other areas of the world, such as other Asian countries (Behera et al., 2011), Europe (Kasprzyk-Hordern et al., 2009; Lindberg et al.,


2005; Vieno et al., 2007b), and U.S. (Brown et al., 2006; Yang and Carlson, 2003), up to several μg L−1. However, there is some difference between specific cases. Sewage antibiotic concentrations more than 10,000 ng L−1 were found in Korea (Behera et al., 2011) and U.K. (Kasprzyk-Hordern et al., 2009) which are relatively higher than that of most studies in China, while sewage from Guangzhou and Hong Kong (Minh et al., 2009; Peng et al., 2006) appeared to have more antibiotics (ND-9460 ng L−1) than the levels (ND-4230 ng L−1) from Sweden (Lindberg et al., 2005), Finland (Vieno et al., 2007b), and U.S. (Brown et al., 2006; Yang and Carlson, 2003). In China, the predominant and frequently detected antibiotic groups, such as macrolides, fluoroquinolones, and sulfonamides, are related to the therapy use pattern and probably area-specific. Considering the intensive usage of veterinary antibiotics in livestock industry (Heddini et al., 2009; Li and Wang, 2009; Zheng and Zhou, 2007) and their significant occurrence in livestock wastewater as indicated by several studies (Ben et al., 2008; Wei et al., 2011), loads of veterinary antibiotics from livestock agriculture may pose a risk to the receiving water environment. 3.1.1.2. Hormones. Hormones as another important group of organic compounds with public health concerns were also found in STP sewage of China. In most cases their concentrations were below the μg L−1 grade, and therefore relatively lower than that of antibiotics. Steroid estrogens, including the natural E1, E2, E3 and synthetic EE2, were mostly studied (Jin et al., 2008; Zhou et al., 2011d, 2011e). Other novel groups of hormones, such as androgens, progestagens, and glucocorticoids, were also investigated (Chang et al., 2011; Liu et al., 2012b). In Pearl River Delta, steroid estrogen species, such as estrone and 17α-estradiol, had occurrence range of 5.0–66.8 ng L−1 in sewage samples of Guangzhou STPs (Yu et al., 2011b). The investigation about steroid estrogens, androgens, progestagens, and glucocorticoids in STPs of Guangdong Province implied the androgen group had the highest total concentration (13.3–1778 ng L−1), followed by glucocorticoids (2.2–192 ng L−1), estrogens (5.6–60.2 ng L−1), and progestagens (6.9–40.5 ng L−1) (Liu et al., 2012b). In Wuhan located in central China, study about seasonal variations of hormone contents in sewage treatment plants indicated contents of natural E1 and E3 tended to be higher in winter than in summer, probably due to their slow decomposition rate at low temperature. The natural (E1, E2, and E3) and synthetic (diethylstilbestrol (DES)) estrogens were believed to be the main contributors of the total estradiol equivalents (EEQs), the index of estrogenic activity of sewage sample which is calculated directly by comparing the EC50 (concentration for 50% of maximal effect) of E2 and the sewage sample (Jin et al., 2008). However, it has been noted that other environmental hormones, such as bisphenol A and alkylphenols, mainly caused the estrogenic effect of sewage (Sun et al., 2008), indicating the main contributor of sewage estrogenic effect is probably case-specific. The study in Beijing indicated estrogen concentrations in different effluents were in the order of pharmaceutical factory and STP inlets N hospital N hennery N chemical factory N fish pool, suggesting the main sources of hormones may rely on municipal wastewater and pharmaceutical-producing factory wastewater (Zhou et al., 2011e). Investigation about androgens, progestogens, and estrogens in sewage of STPs confirmed androgens dominated in both influent and effluent samples, followed by progestogens and estrogens (Chang et al., 2011). Levels of hormones in sewage samples of China are comparable with other areas, and are generally below the μg L−1 grade (Behera et al., 2011; Nakada et al., 2006; Salgado et al., 2010; Servos et al., 2005). Relatively more steroid estrogens with contents exceeding 1000 ng L−1 were found in several studies, such as Beijing in China (ND-4100 ng L−1) (Zhou et al., 2011e) and five cities in Portugal (ND-2484 ng L−1) (Salgado et al., 2010). Both natural and synthetic hormones were detectable in sewage of China, indicating both natural release and therapy use can contribute to the presence of hormones

213

in wastewater. Besides the mostly studied estrogens, other hormone groups, such as androgens and progestagens, may have much higher contents as indicated by several studies (Chang et al., 2011; Liu et al., 2012b). This can be explained by their higher excretion amounts in human urine than those of estrogens (Shemesh and Shore, 2003) and wide application in human and veterinary therapies, especially synthetic progestogens (Chang et al., 2011). Their high occurrence in sewage, sometimes more than several μg L−1, may pose potential harm to the ecosystem. However, due to the analytical difficulty, studies about these hormone groups are relatively lacking. 3.1.1.3. Other pharmaceuticals. Besides antibiotics and hormones, other pharmaceuticals were also present in STP sewage of China. Pharmaceutical groups, including anti-inflammatory drugs, antiepileptic drugs, blood lipid regulators, anti-hypertensive drugs, anticonvulsants, stimulants, and antipsychotic drugs, had concentrations up to 89,500 ng L−1 in influents and 3239 ng L−1 in effluents (Sui et al., 2010, 2011; Zhao et al., 2010b; Zhou et al., 2010). The species with high detection frequency and concentration include caffeine, ibuprofen, diclofenac, carbamazepine, etc. for their wide application, especially caffeine which can act as a tracer of domestic sewage pollution for its widespread occurrence (Siegener and Chen, 2002). Besides Mainland, Taiwanese consumers also use large quantities of pharmaceuticals, and 3.6 tonnes of them are estimated to be dumped annually (Lin et al., 2010). The study about pharmaceuticals in STP sewage of Taipei, Taichung, and Tainan indicated cephalexin, caffeine, acetaminophen, ibuprofen, and naproxen had the highest influent contents, ranging from 458 to 30,967 ng L−1; and ibuprofen and naproxen had the highest effluent contents, ranging from 20 to 3807 ng L−1 (Lin et al., 2010). These studies imply the existence of diverse pharmaceutical groups in STP sewage of China. Levels of some species with high usage quantities, such as caffeine, may exceed that of antibiotics. This is similar with the situation of the studies in other countries (Salgado et al., 2010; Weigel et al., 2004). Pharmaceutical occurrence in STPs of China is similar with the cases from Korea (Behera et al., 2011), Japan (Nakada et al., 2006), Finland (Vieno et al., 2006), and U.S. (Thomas and Foster, 2005), while higher contents up to 482,687, 106,490, and 126,000 ng L−1 were found in the STP sewage samples of U.K. (Kasprzyk-Hordern et al., 2009), Portugal (Salgado et al., 2010), and Norway (Weigel et al., 2004), respectively. 3.1.1.4. Personal care products. Personal care product groups, e.g. antimicrobial agents, synthetic musks, insect repellents, and UV filters, existed in STP sewage with concentration range of ND-3430 ng L−1 for most reviewed cases in China. Triclosan and triclocarban were the most frequently detected antimicrobial agents (Zhao et al., 2010a), and HHCB and AHTN were the most frequently detected synthetic musks (Zeng et al., 2007). Extremely high contents of polycyclic musks up to 595.48 μg L−1 could be found in the cosmetic plant sewage (Chen et al., 2007). Input quantities of HHCB and AHTN into STPs were estimated to be 0.2 and 0.06 g y−1 per capita of Shanghai, and their environmental loads into receiving water were estimated to be 1.26 and 0.38 t in 2007 (Zhang et al., 2008). The widely used insect repellent DEET together with synthetic musks existed in Beijing STPs (Sui et al., 2011), and the sunscreen UV filters were present in Tianjin STPs in North China (Li et al., 2007). Wastewater levels of studied PCP groups are up to the μg L−1 grade for most cases in China, while extremely high contents can be found in manufacturing factories. Some groups of PCPs have received more attention, e.g. polycyclic musks and triclosan, for their wide application in the personal care products. The PCP levels in STP sewage of several other countries, such as U.K. (Kasprzyk-Hordern et al., 2009), Portugal (Salgado et al., 2010), and U.S. (Reiner et al., 2007a), could reach more than 10,000 ng L−1, exceeding the levels

214


in China, which may be attributed to the low per capita consumption of personal care products by the Chinese (Wang et al., 2012). 3.1.2. Fates of PPCPs in STPs Mainly two processes are responsible for the fates of PPCPs in STPs, including biodegradation and adsorption by active sludge. Biodegradation performance of PPCPs by active sludge can be affected by several factors, such as initial substrate concentration, temperature, and biodegradation process type. Biodegradation of four antibiotics decreased when their initial influent concentrations increased, probably due to the inhibition effect of microbial activities by antibiotics (Yu et al., 2009). Biodegradation of synthetic musks showed better performance in summer than in fall which may be attributed to the effect of temperature (Lv et al., 2010). Aerobic and anaerobic biodegradation have varied influence on different types of PPCPs, such as diclofenac diminished mainly through anaerobic biodegradation, while aerobic biodegradation governed the removal of anti-inflammatory drugs (ibuprofen, naproxen, indomethacin) and lipid regulators (bezafibrate, clofibric acid, and gemfibrozil) (Huang et al., 2011). Performance comparison of three kinds of biological treatment processes, including conventional active sludge (CAS), biological nutrient removal (BNR), and membrane bioreactor (MBR), revealed MBR resulted in better performance for the biodegradable species, e.g. caffeine and bezafibrate (Sui et al., 2011). Active sludge can also act as adsorbent for PPCPs in STPs. Majority of the antimicrobial agents (66–82% for triclosan and 82–86% for triclocarban) were absorbed onto sludge (Yu et al., 2011b). Partitioning may act as the dominant mechanism for adsorption, such as for the case of 17β-estradiol (Yu et al., 2011b), while lipophilicity governed the adsorption of antifungal drugs, e.g. clotrimazole, econazole, and miconazole (Peng et al., 2012). Adsorption process can be affected by pH, such as for the case of sulfonamide antibiotics (Yang et al., 2011). Meanwhile, desorption was observed reversibly from the sludge, implying sorbed PPCPs can be introduced into the environment if no further treatment was employed to remove them from the biomass (Yang et al., 2011). Besides the two processes above, it was also found disinfection process could eliminate some kinds of antibiotics, such as ampicillin and cefalexin (Li and Zhang, 2011). 3.1.3. Removal efficiencies of PPCPs in STPs PPCPs can be removed in STPs mainly through the two processes, biodegradation and adsorption. Studies about PPCP removal in STPs suggest different groups of PPCPs perform differently. Antibiotics had wide removal range up to 81%. Fluoroquinolones could be partially removed with adsorption onto active sludge and suspended particulates as the possible mechanism, while majority of macrolides, sulfonamides, penicillin, and imidazole could not be affected by the treatment (Lin et al., 2009; Peng et al., 2011; Xu et al., 2007a). For azole antifungal drugs, fuconazole passed through the treatment unchanged, ketoconazole could be readily biotransformed, while clotrimazole, econazole, and miconazole were more likely to be adsorbed onto the sludge (Peng et al., 2012). As a group of compounds with biodegradable properties, hormones can be eliminated efficiently. The most studied natural and synthetic steroid estrogens, including E1, E2, E3, and EE2, had high removal rates ranging from 83.2% to 98.8% (Zhou et al., 2011d), and the more abundant androgens and progestogens also showed removal efficiency of 91–100% (Chang et al., 2011). Studies in other countries confirmed norfloxacin (the fluoroquinolone) as well as 17β-estradiol and estriol (the steroid estrogens) had removal rates of N 80% in STPs (Miège et al., 2009). For other groups of pharmaceuticals, caffeine, the pharmaceutical with abundant occurrence in many cases, could be removed almost completely (Zhou et al., 2010), iopromide (one kind of X-ray contrast agents) could be greatly removed, β-blockers were moderately eliminated, while carbamazepine, the antiepileptic drug, was not affected by the treatment (Yu et al., 2011b). Studies in other countries also found the low removal efficiency, less

than 20%, for carbamazepine during active sludge treatment (Miège et al., 2009). As to PCPs, disinfectants (triclocarbon, triclosan, and 2-phenylphenol) and preservatives (parabens) were greatly removed (Yu et al., 2011b), while polycyclic musks were partially removed by 26% to 70% with adsorption onto sludge as the dominant mechanism (Zhou et al., 2009a). Besides the species-specific removal performance of PPCPs, different STPs had varied removal for individual compounds owe to their different operation conditions (Zhou et al., 2010). Different treatment processes also perform differently. Sui et al. (2010) found the target compounds, pharmaceuticals and DEET, could not be eliminated by sand filtration, but could be effectively removed by the advanced processes—ozonation and microfiltration/reverse osmosis (MF/RO). From the studies above, it can be concluded some groups of PPCPs, such as fluoroquinolones, hormones, caffeine, antimicrobial agents, and preservatives, can be removed efficiently with biodegradation and adsorption as the main mechanisms. However, some other groups, such as majority of macrolides and sulfonamides, penicillin, fluconazole, and carbamazepine, are refractory to the conventional sewage treatment processes and subsequently discharged unchanged to the receiving water bodies. For the treatment of these PPCPs, advanced technologies, such as oxidation (ozonation) and filtration (microfiltration and ultrafiltration) methods, may need to be introduced. 3.1.4. PPCPs in sludge Through adsorption, groups of PPCPs can be found in sludge samples of STPs in China, including antibiotics (ND-21,000 ng g−1 dry weight (dw)) (Gao et al., 2012; Xu et al., 2007a), azole antifungal drugs (ND-1442 ng g−1 (dw)) (Huang et al., 2010, 2012), hormones (1.6–372 ng g−1 (dw)) (Liu et al., 2011b), other pharmaceuticals (1.7–33.7 ng g−1 (dw)) (Yu et al., 2011a), antimicrobial agents (200.1–5088.2 ng g−1 (dw)) (Yu et al., 2011a), polycyclic musks (700–17,000 ng g−1 (dw)) (Zhou et al., 2009a), and sunscreen UV filters (ND-24,700 ng g−1 (dw)) (Ruan et al., 2012), with concentrations generally between ng g−1 and μg g−1 (dw). These studies reveal PPCPs can be removed by adsorption onto sludge without being completely metabolized or degraded. This also demonstrates the need to describe the composition and fate of PPCPs in sludge, since about half of the sludge is land applied and thus becomes a potentially ubiquitous nonpoint source of PPCPs into the environment (Kinney et al., 2006). To summarize the studies on PPCPs in STPs, groups of PPCPs can be found in STP sewage as well as in sludge samples, including the normally detected antibiotics and hormones as well as other pharmaceutical species (such as caffeine, ibuprofen, and carbamazepine) and personal care products (such as triclosan, HHCB, and AHTN) in regions such as Pearl River Delta, Yangtze River Delta, and Beijing area. In China, large volume of wastewater containing PPCPs flows into sewage treatment plants (SIC, 2011). However, due to the limited removal efficiencies of some groups of PPCPs in STPs, they will be discharged into the receiving water and thus pose a potential risk to the aquatic environment. 3.2. PPCPs in surface water and sediments After discharging from STPs, PPCPs in sewage would result in subsequent contamination to the receiving aquatic environment. Groups of PPCPs have been detected in surface water bodies and sediments from South to North China, including rivers, lakes, and sea, even in the raw water sources of drinking water treatment plants (Table 3). 3.2.1. PPCPs in surface water 3.2.1.1. Antibiotics. Antibiotics occurred in surface water bodies with concentrations generally below the μg L−1 grade (ND-776 ng L−1) for most reviewed cases in China. In Pearl River Delta, antibiotics existed in Dongjiang River which serves as the drinking water source


215

Table 3 Occurrence of PPCPs in surface water. Location

Chemical

Concentration

Media

Reference

Dongjiang River Pearl River Pearl River Pearl River Victoria Harbor

Antibiotics Antibiotics Antifungal drugs Antibiotics Antibiotics

Drinking water source River water River water River water Seawater

Zhang et al. (2012a) Peng et al. (2008a) Huang et al. (2010) Xu et al. (2007b) Xu et al. (2007b)

Sea near Hong Kong Beibu Gulf Jiulongjiang Huangpu River Qiantang River Yellow River and its tributaries Rivers near Beijing Hai River Bo Sea Bay Bo Sea and Yellow Sea Laizhou Bay Dongjiang River Pearl River Pearl River Pearl River Danshui River Yundang Lagoon

Antibiotics Antibiotics Antibiotics Antibiotics Antibiotics Antibiotics Antibiotics Antibiotics Antibiotics Antibiotics Antibiotics Hormones Hormones Hormones Hormones Hormones Hormones Hormones Hormones

Seawater Seawater River water River water River water River water River water River water Seawater Offshore seawater Seawater Drinking water source River water River water River water River water Lake water Pore water River water River water

Gulkowska et al. (2007) Zheng et al. (2012) Zhang et al. (2012c) Jiang et al. (2011) Tong et al. (2011) Xu et al. (2009b) Xiao et al. (2008) Luo et al. (2011) Zou et al. (2011) Zhang et al. (2013) Zhang et al. (2012b) Gong et al. (2009) Gong et al. (2009) Peng et al. (2008b) Zhao et al. (2009) Liu et al. (2011b) Zhang et al. (2010)

Licun River near Jiaozhou Bay Wulo Creek (Taiwan) Pearl River Pearl River Pearl River Yangtze River Rivers near Beijing Yellow River Hai River Liao River (North China) Dongjiang River Pearl River Pearl River Pearl River Pearl River

0.9–67.4 ng L−1 b35–510 ng L−1 b1–6.6 ng L−1 ⁎ND−636 ng L−1 Below the limit of quantification (LOQ) b2–486 ng L−1 ND*−50.9 ng L−1 0.05–775.5 ng L−1 ND−313.4 ng L−1 7.0–51.6 ng L−1 b1–327 ng L−1 1.3–535 ng L−1 26–210 ng L−1 ND−6800 ng L−1 ND−16.6 ng L−1 ND−330 ng L−1 ND−3.6 ng L−1 2.5–8.2 ng L−1 ND−65 ng L−1 ND−75 ng L−1 ND−55.3 ng L−1 0.3–6.8 ng L−1 0.7–6.4 ng L−1 ND−180 ng L−1 ND−1267 ng L−1

Pharmaceuticals Pharmaceuticals Pharmaceuticals Carbamazepine Pharmaceuticals Pharmaceuticals

11.2–102 ng L−1 ND−490 ng L−1 ND−2717 ng L−1 ND−1090 ng L−1 b24–7051 ng L−1 ND−416 ng L−1 ND−127 ng L−1 ND−717 ng L−1

River River River River River River

Zhao et al. (2010b) Zhao et al. (2009) Peng et al. (2008b) Zhou et al. (2011c) Zhou et al. (2010) Wang et al. (2010a)

Antimicrobial agents Antimicrobial agents Antimicrobial agents Triclosan Antimicrobial agents Preservatives Polycyclic musks

b0.7–269 ng L−1 1.5–478 ng L−1 b3.9–478 ng L−1 0.6–347 ng L−1 # NQ−2506 ng L−1 NQ−3142 ng L−1 8–93 ng L−1

Drinking water source River water River water River water River water

Zhao et al. (2013) Zhao et al. (2013) Zhao et al. (2010a) Zhao et al. (2009) Peng et al. (2008b)

River water

Zhang et al. (2008)

Antimicrobial agents

ND−64.7 ng L−1 ND−117 ng L−1 b0.8–404 ng L−1

River water

Zhao et al. (2013)

Antibiotics

7–360 ng L−1

River water

Managaki et al. (2007)

Antibiotics

ND-544 ng L−1

River water

Tamtam et al. (2008)

River water


Suzhou Creek (Shanghai) Yellow River Hai River Liao River (North China) Other counties Vienam (Mekong Delta) France (Seine River) U.K. (Taff and Ely River) Finland (Vantaa River) U.S. (Choptank River) U.S. (Streams in Iowa) U.S. (139 streams) Korea (Youngsan River) Spain (Llobregat River) Netherlands (Scheldt estuary) U.S. (Choptank River) U.S. (139 streams)

−1

water water water water water water

Zhou et al. (2011b) Chen et al. (2010)

Antibiotics

b0.5–183 ng L

Antibiotics

b1.6–36 ng L−1

Drinking water source

Vieno et al. (2007a)

Antibiotics

ND-694 ng L−1

River water

Arikan et al. (2008)

Antibiotics

ND-300 ng L−1

River water

Kolpin et al. (2004)

River water


−1

Antibiotics

ND-1900 ng L

Hormones

1.7–5.0 ng L−1

River water

Kim et al. (2007)

Hormones

2–5 ng L−1

Hormones

River water

Brix et al. (2009)

−1

River water

Noppe et al. (2007)

−1

River water

Arikan et al. (2008)

0.4–10 ng L

Hormones

ND-20 ng L

Hormones

ND-872 ng L−1

River water


Hormones

0.03–18.9 ng L−1

River water

Ferguson et al. (2013) (continued on next page)

216


Table 3 (continued) Location Australia (Little River) Other counties Korea (Youngsan River) Japan (37 rivers and Tamagawa estuarine) U.K. (Taff and Ely River) U.K. (5 rivers) Finland (Vantaa River) U.S. (Streams in Iowa) U.S. (139 streams) Brazil (Rivers in Rio de Janeiro State) India (Kaveri, Vellar, and Tamiraparani River) U.K. (Taff and Ely River) U.S. (Michigan Lake) U.S. (Streams in Iowa)

Chemical

Concentration

Media

Reference

Pharmaceuticals

1.1–361 ng L−1

River water

Kim et al. (2007)

Pharmaceuticals

ND–749 ng L

−1

River water

Nakada et al. (2008)

Pharmaceuticals

ND–5970 ng L−1

River water


Pharmaceuticals

b1–928 ng L−1

River water

Thomas and Hilton (2004)

Drinking water source

Vieno et al. (2007a)

River water


Pharmaceuticals

3–107 ng L

−1

−1

Pharmaceuticals

ND–1950 ng L

Pharmaceuticals

ND–10000 ng L−1

River water


Pharmaceuticals

20–500 ng L−1

River water

Stumpf et al. (1999)

River water

Ramaswamy et al. (2011)

−1

Triclosan

Max. 5160 ng L

Antimicrobial agents Preservatives UV filters Synthetic musks

b1–358 ng L−1 b0.2–305 ng L−1 b0.3–323 ng L−1 0.03–4.7 ng L−1

River water


Lake water

Peck and Hornbuckle (2004)

−1

River water


HHCB and ANTH Triclosan DEET

ND–1200 ng L ND–140 ng L−1 ND–130 ng L−1

⁎ND: not detected. # NQ: detected but. too low to be quantified.

for Hong Kong and Guangdong Province in southern China, with concentration range of 0.9–67.4 ng L−1. Macrolides in Dongjiang River were probably related to domestic sewage discharge, while sulfonamides and trimethoprim might be attributed to livestock agriculture wastewater (Zhang et al., 2012a). Xu et al. (2007b) compared the antibiotic levels in Pearl River of Guangzhou and its nearby Victoria Harbor of Hong Kong in Pearl River Delta. The antibiotic concentrations in the coastal water of Victoria Harbor were generally below the limit of quantification (LOQ) while target analytes were all detectable in Pearl River, implying possible more serious contamination of Pearl River by antibiotics. Antibiotic presence in the vicinity of aquaculture in Beibu Gulf suggested the contribution of aquaculture activities to antibiotic contamination in coastal water (Zheng et al., 2012). For Huangpu River in Yangtze River Delta, more veterinary antibiotics, such as tetracyclines, were monitored in the suburban sampling sites compared with the urban sampling sites, confirming the role of livestock wastewater as antibiotic contamination source (Jiang et al., 2011). The Hai River in North China contained veterinary antibiotics owe to the release of swine farms and fishponds (Luo et al., 2011). In the nearby coastal water of Bo Sea Bay, high density of human activities in the north bay resulted in more serious antibiotic contamination than the south bay (Zou et al., 2011). Antibiotics also occurred in the offshore water of Bo Sea and Yellow Sea with concentration of ND-16.6 ng L−1, and the risk assessment based on risk quotients (RQs) showed medium to low ecological toxicity (0.01 b RQ b 1) derived from sulfamethoxazole, dehydration erythromycin, and clarithromycin to some sensitive aquatic organisms, including Synechococcus leopoliensis and Pseudokirchneriella subcapitata (Zhang et al., 2013). These studies confirm that wastewater discharge is the main source of surface water pollution by antibiotics. Similar with the situation of sewage, contribution from livestock agriculture wastewater discharge is important for antibiotic contamination. Surface water in other countries seemed to be contaminated to a similar extent compared with the situation of China, such as Vienam (Managaki et al., 2007), France (Tamtam et al., 2008), U.K. (Kasprzyk-Hordern et al., 2009), Finland (Vieno et al., 2007a), and U.S. (Arikan et al., 2008; Kolpin et al., 2004), generally

below the μg L−1 grade (ND-694 ng L−1). In China antibiotics up to 6800 ng L−1 were found in seawater of Bo Sea Bay (Zou et al., 2011). Kolpin et al. (2002) measured 95 kinds of organic wastewater contaminants in water samples from a network of 139 streams across 30 states of U.S. during 1999 and 2000, and antibiotic contents in these streams could reach 1900 ng L−1. 3.2.1.2. Hormones. Hormones occurred in surface water of China with content level generally below μg L−1 (ND-180 ng L−1) for most reviewed studies, and also relatively lower than the antibiotic group. They were also found in the drinking water source—Dongjiang River with concentration up to 3.6 ng L−1 (Gong et al., 2009). In two reservoirs of Shenzhen, the total estrogen levels were in the range of 3–11 ng L−1 (Lee et al., 2006). The Wulo Creek in Taiwan had hormone occurrence up to 1267 ng L−1 due to livestock feedlot discharge (Chen et al., 2010). In Beijing, hormone groups, e.g. estrogens, androgens, progestogens, glucocorticoids, and mineralocorticoids, were analyzed in 45 urban rivers, and it was found similar with the situation of sewage, androgens dominated among these groups, followed by glucocorticoids, progestogens, and estrogens. The principal component analysis for investigating the main contributor of urban river water contamination suggested 62.7% of the mean summed hormones resulted from untreated sewage, 29.4% from treated sewage and/or naturally attenuated untreated sewage, and 7.9% from an unknown source, possibly pharmaceutical manufacturing plants (Chang et al., 2009). These studies reveal the occurrence of hormones in surface water bodies as well as drinking water sources by the effect of wastewater discharge, while untreated wastewater is probably the main contributor in some cases. For other countries, rivers in most reviewed cases had registered less than 20 ng L−1 of hormones (Arikan et al., 2008; Brix et al., 2009; Ferguson et al., 2013; Kim et al., 2007; Noppe et al., 2007), except for the nationwide investigation in U.S. (ND-872 ng L−1) (Kolpin et al., 2002). 3.2.1.3. Other pharmaceuticals. Pharmaceutical groups, including antiinflammatory drugs (e.g. salicylic acid, ibuprofen, diclofenac,


mefenamic acid, naproxen), blood lipid regulators (e.g. clofibric acid, gemfibrozil), and antiepileptic drugs (e.g. carbamazepine), showed widespread occurrence in the rivers from South to North China, such as Pearl River (Peng et al., 2008b), Yangtze River (Zhou et al., 2011c), Yellow River, Hai River, and Liao River (Wang et al., 2010a), with concentration range of ND-7051 ng L− 1. Several cases found pharmaceuticals more than 1000 ng L− 1 in the water bodies (Peng et al., 2008b; Zhou et al., 2010, 2011c). Similar with the situation of sewage, caffeine dominated in some cases (Zhou et al., 2010). Detection frequency and median concentration of pharmaceuticals appeared higher during the low-flow season than the high-flow season, probably due to the dilution effect by rainfall (Peng et al., 2008b). Sites near metropolitan areas appeared more contaminated (Wang et al., 2010a), and the pharmaceutical pollution may be closely associated with the direct discharge of untreated wastewater (Yu et al., 2011b). The contamination level of surface water by pharmaceuticals in other countries seemed comparable with China, such as Korea (Kim et al., 2007), Japan (Nakada et al., 2008), U.K. (Kasprzyk-Hordern et al., 2009; Thomas and Hilton, 2004), U.S. (Kolpin et al., 2002; Kolpin et al., 2004), and Brazil (Stumpf et al., 1999), with concentration range of ND-10,000 ng L− 1. Pharmaceuticals up to 107 ng L− 1 were found in Vantaa River as drinking water source in Finland (Vieno et al., 2007a). 3.2.1.4. Personal care products. Personal care products in surface water of China had an occurrence range of ND-3142 ng L−1, while contents of most cases were below the μg L−1 level. Urban rivers of Guangzhou in Pearl River Delta contained disinfectants (triclosan, triclocarban, 2-phenylphenol) and preservatives (methyparaben, propylparaben) with high detection frequency and concentration (Peng et al., 2008b; Zhao et al., 2009). Triclosan and triclocarban in rivers, including Dongjiang River, Pearl River, Yellow River, Hai River, and Liao River, had high occurrence in sites located near the urban area and were significantly influenced by the factors such as the total or untreated urban domestic sewage discharge at river basin scale (Zhao et al., 2013). Polycyclic musks occurred in Suzhou Creek of Shanghai in Yangtze River Delta with HHCB and AHTN as the two main species (Zhang et al., 2008). When looking at the PCP studies worldwide, Pearl River in China with antimicrobial agents up to 2506 ng L−1 and preservatives up to 3142 ng L−1 (Peng et al., 2008b) and Kaveri, Vellar, and Tamiraparani River in India with triclosan maximum concentration of 5160 ng L−1 (Ramaswamy et al., 2011) appeared to be more contaminated by PCPs. 3.2.2. Fates of PPCPs in surface water In natural surface water bodies, PPCPs may undergo several environmental processes, mainly adsorption by sediments and photolysis. Majority of antimicrobial agents, such as triclosan, appeared to be adsorbed onto sediments. Pores of sediments can capture triclosan, while the sediment organic matter may bound triclosan and reduce its solubility. The adsorption process can be affected by pH (Lin et al., 2011). Adsorption can remove PPCPs from the water bodies, however, sediments may also become a sink for PPCPs which may be released back to the aquatic environment (Zhao et al., 2013). Aqueous photolysis was observed for tetracycline, and the process could be enhanced by high pH (Jiao et al., 2008). The nonsteroidal anti-inflammatory drug diclofenac can also be degraded by sunlight. pE value of the aqueous environment may affect the forms of N and further influent the photolysis process (Zhang et al., 2011a). Further photolytic mechanism study for the different dissociation species of ciprofloxacin indicated five dissociation species had dissimilar photolytic kinetics, including stepwise cleavage of the piperazine ring, defluorination, and oxidation. For accurate ecological risk assessment of ionizable emerging pollutants, study

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about the environmental fates of different dissociation species is therefore suggested (Wei et al., 2013). 3.2.3. PPCPs in sediments PPCPs can be found in sediments of surface water through adsorption, but with much lower levels compared with sludge samples. Antibiotics (ND-1560 ng g− 1 (dw)) (Yang et al., 2010), antifungal drugs (1–35 ng g− 1 (dw)) (Huang et al., 2010), hormones (b0.9– 10.9 ng g− 1 (dw)) (Gong et al., 2011), antimicrobial agents (b0.6– 2633 ng g− 1 (dw)) (Yu et al., 2011a; Zhao et al., 2010a), polycyclic musks (b 0.3–121 ng g− 1 (dw)) (Zeng et al., 2008) existed in the sediments of Pearl River and coastal water in southern China, antibiotics (ND-5770 ng g− 1 (dw)) occurred in the sediments of Yellow River, Hai River, and Liao River in northern China (Zhou et al., 2011a), and UV filters (0.3–7.1 ng g− 1 (dw)) were present in sediments of Songhua River in northern China (Zhang et al., 2011c), suggesting the PPCP adsorption on sediments as well as the possible desorption. For better understanding PPCPs in sediments as a potential pollution source to surface water, more studies about PPCPs in sediments are needed. 3.2.4. Removal efficiencies of PPCPs in water treatment plants About the removal efficiency of PPCPs in water treatment plants, Qiao et al. (2011) stated conventional water treatment processes (coagulation and flocculation) removed the type and concentration of PPCPs by 30% and 50% respectively, while advanced treatment processes could achieve the reduction ratios of 50% and 90% respectively. Study in Korea confirmed that conventional treatment had limited PPCP removal, however, granular activated carbon adsorption could result in efficient control (Kim et al., 2007). Another study also found the conventional coagulation and sedimentation showed much marginal removal of studied pharmaceuticals by 3%, sand filtration could achieve more additional removal of 10%, while the advanced oxidation method, ozonation, reduced most pharmaceuticals from up to 107 ng L− 1 to below the detection limits of 0.2–7.0 ng L− 1 (Vieno et al., 2007a). To summarize the studies about PPCPs in surface water bodies, different groups of PPCPs occur ubiquitously in aquatic environment from South to North China, owing to the impact of municipal and agriculture wastewater discharge. Antibiotics and steroid estrogens can even be found in source water of drinking water treatment plants. The polluted surface water can be attributed to the huge wastewater discharge volume and limited wastewater treatment rate in current China. The discharge volume of municipal, industry, and agriculture wastewater is huge as well as with a rapid growth rate. In 2009, 59 billion tonnes of wastewater were discharged into the receiving aquatic environment of China. From 2003 to 2009, the municipal sewage discharge volume increased by 40% (SIC, 2011). However, the wastewater treatment rate is relatively low due to the lack of facilities, compared with the developed countries. Wastewater treatment rate in rural regions of China was as low as 60% in 2010 (MOHURD, 2012), meaning that up to 40% of the wastewater was discharged directly into the receiving water bodies without any treatment. The current wastewater discharge and treatment condition finally results in the pollution of surface water by PPCPs in China. Water pollution is a large challenge for the developing China, and numerous efforts need to be made to solve this problem in the future. Removal efficiency of PPCPs in water treatment plants is not satisfactory by the conventional treatment processes, and thus advanced steps, e.g. ozonation and activated carbon adsorption, need to be introduced for the control of PPCP problem in source water. Besides surface water as drinking water source, the studies about PPCPs in groundwater are urgent. Meanwhile, more studies about PPCPs in seawater which possesses varied physical, chemical, and biological properties with river water are needed.

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3.3. PPCPs in soil

4. Human exposure and body loading of PPCPs

PPCPs can be introduced into soil through sludge land application or landfill, use of livestock wastes as fertilizers, and reclaimed water irrigation. Pollutants in soil may be accumulated in plants or migrate through soil intact or transformed and reach groundwater, finally resulting in pollution to the drinking water source. Veterinary antibiotics existed in the soil from organic vegetable farmland fertilized with livestock wastes in Tianjin (Hu et al., 2010a), with concentrations up to 2683 ng g− 1 (dw). Azole biocides were detected in biosolid-amended soil and showed persistence with dissipation half lives up to 440 days (Chen et al., 2013). Chen et al. (2011) found irrigation with wastewater could result in more PPCPs in soil compared irrigation with groundwater. The study about sorption and degradation behavior of drugs in soil discovered some drug species, such as diclofenac and ibuprofen, showed poor adsorption as well as with persistence under anaerobic condition (Lin and Gan, 2011). The simulated turf grass irrigation for the study of PPCP leachability discovered estrone, ibuprofen, naproxen, clofibric acid, and triclosan in the 30-cm soil profiles, indicating the potential to contaminate groundwater. The leachability of PPCPs may be affected by the chemicals' characteristics (such as pKa values), properties of soil (such as soil organic matter and clay content), and salinity of the irrigation water (Xu et al., 2009a; Xu et al., 2010). These studies reveal PPCPs can be introduced into soil by these routes. Some groups of PPCPs may perform poor adsorption in soil and be refractory to biodegradation, suggesting their possible leaching risk. Therefore, reclaimed water irrigation and sludge and livestock waste land application need to be evaluated carefully for the potential risk of PPCP introduction.

4.1. Human exposure

3.4. PPCPs in wild animals The studies about PPCPs in wild animals of China are limited. Fluoroquinolones were found in fish species from marine aquaculture regions of the Pearl River Delta, including rabbitfish, black porgy, and gray snapper, with liver concentration up to 254.6 ng g− 1 wet weight (ww) (He et al., 2012). Steroid estrogens had contents up to 11.3 ng g− 1 (dw) in wild fish species from Dianchi Lake in Southwest China, e.g. crucian carp, carp, and silvery minnow. Liver had the highest estrogen accumulation, following by gill and muscle (Liu et al., 2011a). Wan et al. (2007) investigated the distribution of musk fragrances in different organs of Chinese sturgeons and found higher concentration of fragrances in adipose tissues (1.0–62.1 ng g− 1 (ww)), indicating their distribution was associated with the affinity to lipid. The lipid-corrected concentrations increased when the age of Chinese sturgeons increased, which is consistent with the situation of fish, but different from that of mammals. Hu et al. (2011) discovered the biota-sediment accumulation factors (BSAFs) higher than 1.7 for two common polycyclic musks, HHCB and AHTN, in crucian carp, common carp, and silver carp, revealing the possible bioaccumulation of these fragrances. The contents of synthetic musks in wild fish from China are comparable with that from Japan (b1.9–149 ng g− 1 (ww)) (Nakata, 2005) and Denmark (ND-52.6 ng g− 1 (ww)) (Duedahl-Olesen et al., 2005). The study in Denmark also revealed the decreasing tendency of synthetic musks in wild fish according to time from 1992 to 2004 (Duedahl-Olesen et al., 2005). These studies suggest the presence and possible bioaccumulation of PPCPs in wild animals. However, the limited studies about PPCPs in wildlife of China so far are mainly focused on antibiotics, hormones, and synthetic musks. Occurrence and behavior of other PPCP groups in wild animals need to be investigated in the future, and the potential adverse effects need to be evaluated.

Humans can be exposed to PPCPs mainly through daily use of personal care products, inhalation, and dietary intake. More than 80% of the investigated daily personal care products, including tooth pastes, hair care products, body washes, soaps, skin lotions, and makeup, contained synthetic musks with total contents as high as 1.02 mg g− 1. HHCB was the predominant species in all of the personal care products analyzed, accounting for 52% of the total musk concentrations on average. The dermal exposure rates in adults were calculated to be 3.38 mg d− 1 for musks (Lu et al., 2011). For the low volatility of PPCPs, their participation into atmospheric environment would be limited. However, some PPCPs, mainly groups with relatively high volatility, were still found in indoor dusts or air. Siloxanes which are widely used in personal care and other consumer products were found in the indoor dust samples with total concentration range of 21.5 to 21,000 ng g−1. Their concentrations in indoor dusts were associated with the number of electrical/electronic appliances and number of occupants and smokers living in the house. Toddlers had higher estimated daily intake than adults (Lu et al., 2010b). Another study investigated the occurrence of two polycyclic musks, three nitro musks, and one metabolite of HHCB (HHCB lactone) in indoor dust samples. The total concentrations of musks ranged from 4.4 to 688 ng g−1 with HHCB as the predominant species (Lu et al., 2010a). In the polycyclic musk investigation from a typical cosmetic plant in Guangzhou, except for sewage and sludge, polycyclic musks were found to be mostly contained in the gas phase at the percentage of 86.4–97.7% (Chen et al., 2007). Parabens in indoor dusts from four countries, including U.S., China, Japan, and Korea, were investigated, and paraben species with contents up to 110,800 ng g−1 were detected. Geometric mean concentrations of the investigated parabens in indoor dusts from four countries were in the order of Korea (2320 ng g−1) N Japan (2300 ng g−1) N U.S. (1390 ng g−1) N China (418 ng g−1). The relatively lower parabens in dusts from China is probably related to the low per capita consumption of personal care products by the Chinese (Wang et al., 2012). PPCPs existed in drinking water and food as the dietary sources for humans. Four kinds of PPCPs existed in the drinking water from distribution network with contents around 1 ng L−1 in South China (Qiao et al., 2011). Fluoroquinolone antibiotics also occurred in the tap water of Guangzhou and Macau (Wang et al., 2010b). Triclosan was present in both tap water and bottled water with concentrations up to 14.5 and 9.7 ng L−1 respectively. Triclosan may also be released from baby bottles. Daily intakes of triclosan by adults and infants were estimated to be 10 and 5 ng d−1 respectively (Li et al., 2010a). Vegetables contained antibiotics in the organic farm bases with livestock wastes as fertilizers in Tianjin of North China, with concentration up to 532 ng g−1. The antibiotics accumulated in vegetables probably through water transport and passive absorption, and the distribution in the plant was in the order of leaf N stem N root (Hu et al., 2010a). 4.2. Body loading Through the different exposure pathways, PPCPs were found in kinds of human samples, e.g. human milk, blood, and urine. Synthetic musks existed in the human milk samples from the mothers living in Chengdu of Southwest China, with concentration of b 1.4–16.5 ng g−1 lipid weight. Their occurrence was related with the frequent use of personal care products, such as hand-cleaning agents, body lotions, shampoos, hair dyes, and hair gels (Yin et al., 2012). Four synthetic musks, MX, MK, HHCB, and AHTN, were also present in the human milk samples from three cities of eastern China with contents of 4–63 ng g−1


lipid weight. The dominant species was HHCB, followed by MX. Daily intake of synthetic musks by infants from milk was estimated to be 277–7391 ng d−1 (Zhang et al., 2011b). The occurrence of synthetic musks in human milk of China seems relatively lower compared with the studies of U.S. (b2–917 ng g−1 lipid weight) (Reiner et al., 2007b), Denmark (38–422 ng g−1 lipid weight) (Duedahl-Olesen et al., 2005), and Sweden (b2–268 ng g−1 lipid weight) (Lignell et al., 2008). Two nitro musk fragrances and five polycyclic musk fragrances were measured in human blood samples from 11 cities of China. HHCB had the highest content of 0.85 ng g−1 and detection frequency of 91%, followed by AHTN with the highest content of 0.53 ng g−1 and detection frequency of 77% (Hu et al., 2010b). The study on synthetic musks in human blood plasma of young adults from Austria discovered the dominant HHCB and MX with median concentrations of 420 and 11 ng g−1, respectively (Hutter et al., 2009), which are significantly higher than the previous study in China. Triclosan was present in the urine samples of 287 children and students aged from 3 to 24 years old in Guangzhou, with detection rate of 93% and mean concentration of 3.55 μg g−1 creatinine (3.77 μg g−1) (Li et al., 2013). The occurrence of parabens and the metabolites were investigated in the urines of U.S. and Chinese people. Medium concentrations of the sum of 6 investigated parabens in urine of U.S. and Chinese children were 54.6 and 10.1 ng mL−1 respectively, while the concentration for Chinese adults was 33.2 ng mL−1, which was relatively lower than the reported value for U.S. adults. Elevated urinary paraben concentration (1000– 10,000 ng mL−1) were found for Chinese adult females (Wang et al., 2013). It is believed PPCPs pose no significant risk to human health based on current environmental exposure and human body burden data (Touraud et al., 2011). However, the studies above demonstrate the accumulation of PPCPs in human bodies, and their occurrence in human milk implies the possible impact to infants. Studies about PPCP human exposure and body loads are mainly about synthetic musks, antimicrobial agents, and parabens. Investigation about other PPCP groups is urgent. Meanwhile, for the full understanding of potential risks derived from PPCP exposure, a comprehensive investigation about the contribution of different exposure routes is needed. 5. Toxicity and risk assessments of PPCPs 5.1. Toxicity The wide occurrence of PPCPs in different environmental media raises concerns about their potential harm to ecosystem and human health. Based on the toxicological data and environmental concentration levels, risks of acute toxic effects from these PPCPs are believed to be not likely (Brausch and Rand, 2011; Fent et al., 2006). However, as indicated by the case that residues of veterinary diclofenac caused the significant decline of vulture population in Pakistan, severe adverse effects such as wildlife population decline may occur (Lindsay and Martin, 2004). Meanwhile, chronic effects could not be excluded. The continuous input of PPCPs into the environment may accumulate and cause irreversible harm to wildlife and human beings (Brausch and Rand, 2011; Fent et al., 2006). Environmental exposure to antibiotics may accelerate the persistence or emergence of antibiotic resistance genes (ARGs) which pose potential harm to ecosystem and human health (Kemper, 2008; Zhang et al., 2009b). Antibiotic resistance genes encoding resistance to a broad range of antibiotic species, such as macrolides, sulfonamides, fluoroquinolones, and tetracyclines, occur ubiquitously in hospital and livestock feeding effluents, municipal wastewater, surface water, as well as drinking water resources (Pruden et al., 2006; Zhang et al., 2009b). On the other hand, multiple antibiotic resistant (MAR) superintegrons, which may contain over 100 ARG cassettes, have been also discovered (Pruden et al., 2006). Horizontal

219

gene transfer (HGT) is a major mechanism for sharing ARGs between microorganisms, and antibiotic resistance determinants can be transmitted to other animals and humans via consumption of polluted water as well as agricultural and aquacultural products (Kobayashi et al., 2007). The situation with respect to antibiotic resistance derived from antibiotic abuse in China is severe (Reynolds and McKee, 2009). In 2010, the national Ministry of Health collected more than 270,000 isolated bacteria samples from 128 hospitals across the country to study the antibiotic resistance situation. Results showed the resistance rates had reached nearly 80%, creating a severe situation for public health (Chinadaily, 2012). The level of antibiotic resistance in China was compared with Kuwait and U.S., revealing China had the highest value. China also had the most rapid growth rate of resistance (22% average growth in a study period of 1994 to 2000), while Kuwait was second (17% average growth from 1999 to 2003), and U.S. was the lowest (6% average growth from 1999 to 2002) (Zhang et al., 2006). ARGs were widely found in the aquatic environment of China. Tetracycline resistance genes existed in sewage of wastewater treatment plants in Hong Kong and Shanghai (Zhang et al., 2009a). Escherichia coli resistant to antibiotics, mainly sulfonamides, tetracycline, and ampicillin, existed in Wenyu River Basin in Beijing, while multiantibiotic resistance was also discovered (Hu et al., 2008). Sulfonamide resistance genes occurred in Hai River. Their concentrations in sediments were 120–2000 times higher than that in water, indicating sediments are probably an important reservoir for ARGs in surface water (Luo et al., 2010). Hormones may result in the adverse effects of endocrine disruption. These endocrine disruptors can have a great range of reproductive and development effects, e.g. reduced fertility, feminization of males, induction of vitellogenesis in male (plasma vitellogenin induction), and intersex phenomena (Lai et al., 2002). These hormones possess 10,000–100,000 times higher estrogenic potency compared with exogenous endocrine disrupting chemicals (Khanal et al., 2006), and are believed to be the most significant estrogenic compounds present in many sewage effluents (Jobling et al., 2006). A number of aquatic species, for example, crucian carp, trout, minnow, and turtle, have been reported to be sexually inhibited or reversed by the presence of estrogens at environmental levels (Irwin et al., 2001; Liu et al., 2012a; Tabata et al., 2001; Zha et al., 2008). For the personal care product group, parabens and UV filters can also act as endocrine disruptors (Gomez et al., 2005), while triclosan is suspected to result in the endocrine disrupting effects (Foran et al., 2000). Table 4 lists some examples of adverse effects from different kinds of PPCPs. Besides the antibiotic resistance by antibiotics and endocrine disruption by hormones, it was found antibiotics (Yang et al., 2008) and gemfibrozil (blood lipid regulator) (Rosal et al., 2009) as well as triclosan and triclocarban (antimicrobial agents) (Yang et al., 2008) could inhibit the growth of algae. Caffeine (stimulant drug) may result in endocrine disruption of goldfish (Li et al., 2012), and propranolol (beta-blocker) can reduce the viable eggs of Japanese medaka (Huggett et al., 2002). Carbamazepine (antiepileptic drug) and HHCB (polycyclic musk) may result in oxidation stress to rainbow trout (Li et al., 2010b) and goldfish (Chen et al., 2012). Diclofenac, the nonsteroidal anti-inflammatory drug, may cause renal lesions and gill alterations to rainbow trout (Schwaiger et al., 2004). Furthermore, synergistic effects of these toxic chemicals can be a concern. Tests with combinations of various pharmaceuticals (carbamazepine, diclofenac, and ibuprofen) revealed stronger effects than expected from the effects measured singly to the target aquatic organism—Daphnia magna (Cleuvers, 2003). Another concern about PPCPs is their potential bioaccumulation and biomagnification in aquatic food web. Several groups of PPCPs, such as UV filters, disinfectants, and synthetic musks, have been approved to be able to bioaccumulate and probably biomagnify and may finally cause adverse effects to human beings (Brausch and Rand, 2011).

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Table 4 Adverse effects of PPCPs. Chemical

Adverse effect

Exposure dose

Exposure span

Reference

Roxithromycin, clarithromycin, tylosin (Antibiotics) Caffeine (Stimulant drug) Diclofenac (Nonsteroidal anti-inflammatory drug) Carbamazepine (Antiepileptic drug) Gemfibrozil (Blood lipid regulator) Propranolol (β-blocker) HHCB (Synthetic musk) Triclosan and triclocarban (Antimicrobial agents)

Growth inhibition of algae (Pseudokirchneriella subcapitata)

40–64 μg L−1

3d

Yang et al. (2008)

Endocrine disruption of goldfish (Carassius auratus)

2000 μg L−1

7d

Li et al. (2012)

28 d

Schwaiger et al. (2004)

Renal lesions and gill alterations of rainbow trout (Oncorhynchus mykiss)

5 μg L

−1

−1

Oxidation stress of rainbow trout (Oncorhynchus mykiss)

200 μg L

42 d

Li et al. (2010b)

Growth inhibition of algae (Anabaena sp.)

4420 μg L−1

1d

Rosal et al. (2009)

Reduction of viable eggs of Japanese medaka (Oryias latipes)

0.5 μg L−1

28 d

Huggett et al. (2002)

−1

Oxidation stress of goldfish (Carassius auratus)

1.5 μg L

7–21 d

Chen et al. (2012)

Growth inhibition of algae (Pseudokirchneriella subcapitata)

0.4–10 μg L−1

3d

Yang et al. (2008)

5.2. Risk assessments

6. Regulations and control strategies of PPCP contamination

Many studies about risk assessments of PPCPs base on the European Medicines Agency (EMA) guidelines, in which the risk quotient is calculated as the ratio between predicted environmental concentrations (PEC) and predicted no-effect concentrations (PNEC). If risk quotient is higher or equal to one, it suggests that the particular substances could cause potential adverse ecological effects (EMA, 2006). Stuer-Lauridsen et al. (2000) evaluated the risks of 25 most used pharmaceuticals in aquatic environment of Denmark, and PEC/PNEC ratios of ibuprofen, acetylsalicylic acid, and paracetamol were found to exceed one, and the ratios of estrogens approached one. The risk assessment for most used drugs in aquatic environment of England found the PEC/PNEC ratios of paracetamol, amoxycillin, oxytetracycline, and mefenamic acid exceeded one (Jones et al., 2002). Besse and Garric (2008) evaluated 120 pharmaceuticals and 30 metabolites in the aquatic environment of France according to their biological and physicochemical properties as well as predicted environmental concentrations, and 40 pharmaceuticals and 14 metabolites were suggested to need prioritized attention. Prioritization and ranking methods are needed for screening target compounds for their potential risks. Cooper et al. (2008) developed a risk assessment database for commonly used pharmaceuticals based on their environmental behavior and ecological toxicity focusing on the marine and esturarine environment, and anti-infectives were found to pose the highest risk. A prioritization approach based on number of prescriptions and toxicity information as well as human metabolism and wastewater treatment removal information was performed on 200 most-prescribed drugs in U.S. in 2009, finding levothyroxine and montelukast sodium received the highest scores (Dong et al., 2013). A ranking system based on the pharmaceutical consumption, removal performance in the wastewater treatment plants, and potential ecological effects was applied to the situation of China. Erythromycin, diclofenac acid, and ibuprofen had the high priority among 39 pharmaceuticals. Antibiotics contributed most to the pharmaceuticals with priority. However, only 32% of the antibiotics were listed as priority drugs, compared with the 71% and 100% for anti-inflammatory and antilipidemic drugs. Thus, the author stated antibiotics might be overanxiously considered (Sui et al., 2012). In general, based on the current environmental contamination levels of PPCPs, it is believed the acute toxicity caused by PPCPs has low possibility, however, chronic toxicity could not be excluded. The high level and rapid growth of antibiotic resistance in China derived from antibiotic abuse need a special attention. Current risk assessments of PPCPs are mainly based on the acute toxicity data, more accurate evaluation considering chronic toxicity exposed at environmental concentration levels is urgent. Moreover, more risk assessments for personal care product groups should be conducted.

Regulations are established for PPCP management in several countries. In U.S. the assessment of environmental risks from pharmaceuticals has been required by U.S. Food and Drug Administration (U.S.FDA) under the National Environmental Policy Act (NEPA) since 1969 (U.S.FDA, 1969), and in 1998 the guidance for a tiered risk assessment method was published by Center for Drug Evaluation and Research (CDER) of U.S.FDA (CDER, 1998). In the same year U.S. Environmental Protection Agency (U.S.EPA) promulgated revised regulations for the pharmaceutical industry to control both effluent discharges and air emissions (U.S.EPA, 1998). For European Union (EU), in 2006, the first Guideline for Environmental Risk Assessment of Human Medicines was published by the European Medicines Agency (EMA, 2006). Switzerland, a non-member of EU, requires only ecotoxicity information for human pharmaceuticals and both ecotoxicity and potential environmental risks for veterinary drugs (AMZV, 2001). In Australia, pharmaceutical environmental risk assessment is required with new medicine registration by Australian Therapeutic Goods Administration (TGA, 2008). In China, the environmental management regulations for hazardous chemicals include the Environmental Management Method for New Chemicals promulgated in 2010 (MEP, 2010) and Environmental Management and Registration Method for Hazardous Chemicals which comes into force recently in March 2013 (MEP, 2012). The environmental risk assessments of drugs and cosmetics are not included in these regulations, however, their ingredients and intermediates during production are regulated under these rules. The human health risk assessments of new drugs are required under the Pharmaceutical Administration Law (CFDA, 2001), however, environmental risk assessments of drugs and cosmetics are not regulated currently. The livestock agriculture waste discharge volume is regulated by the national Ministry of Environmental Protection (MEP) of China (MEP, 2001). Wastewater discharge volume of pharmaceutical manufacturers producing 16 kinds of drugs, e.g. caffeine, ibuprofen, and sulfadiazine, is also regulated (MEP, 2008). In the future, regulations about environmental risk assessment as well as the assessment guidelines for PPCPs should be established in China. Source control is a direct and efficient strategy. Expired drugs should be collected through the establishment of a drug management and reclamation scheme. Removal performance of PPCPs in STPs as well as in drinking water treatment plants needs to be improved by the adoption of advanced treatment technologies. In addition, sewage sludge landfill, livestock waste land use, and reclaimed water irrigation need more careful evaluation considering possible introduction of PPCPs into soil and groundwater environment. For the specific situation of China, the antibiotic abuse needs to be improved in the future for the control of current serious antibiotic resistance. More wastewater


treatment facilities need to be established to increase the current sewage treatment rate and avoid direct sewage discharge into aquatic environment, especially in the rural areas of China. Protection of drinking water sources is not sufficient in current China. Management and supervision of drinking water sources need to be strengthened. Any possible pollution source of PPCPs, e.g. wastewater treatment plants, livestock agriculture, hospitals, and cosmetic plants, should not be located in the vicinity of drinking water sources, and any illegal wastewater discharge to the source water should be strictly forbidden. 7. Conclusions PPCPs are present ubiquitously in different environmental media of China. The contamination levels of PPCPs in sewage and surface water are in the range of ng L−1 to μg L−1, while the levels in sludge, sediments, and soil are in the range of ng g−1 to μg g−1 (dw). With huge wastewater discharge volume in China, the environmental loads of PPCPs are tremendous. The surface water pollution by PPCPs may be attributed to the huge wastewater discharge volume and limited wastewater treatment rate in current China. For the protection of surface water, there are great needs to improve the wastewater treatment rate to reduce the direct discharge. In some cases, PPCPs can be detected in source water of drinking water treatment plants or even tap water. It is urgent to protect the water sources from PPCP contamination. There is severe lack of information about other environmental media contamination status of PPCPs in China, such as surface and ground drinking water sources, wildlife, as well as the human body burden. Areas other than the several hotspots all over China need to be investigated about their PPCP pollution status. Meanwhile, most studies available so far mainly focused on the occurrence of PPCPs. The transportation and transformation behavior in different environmental media as well as control technologies of PPCPs need to be investigated. Acute and chronic toxicity studies for different groups of PPCPs need to be conducted in the future for the assessments of their potential ecological and health risks. Acknowledgments Financial support from Seed Collaborative Research Fund from the State Key Laboratory in Marine Pollution (SCRF0003) and National Natural Science Foundation of China/Research Grants Council (N_HKBU210/11) are gratefully acknowledged. References AMZV (Arzneimittel-Zulassungsverordnung). Verordnung des Schweizerischen Heilmittelinstituts vom 9. November 2001 über die Anforderungen an die Zulassung von Arzneimitteln (Arzneimittel-Zulassungsverordnung, AMZV), Stand am 12. September 2006. SR 812.212.22. Berne: The Swiss Confederation; 2001. [http://www.admin.ch/ch/d/sr/c812_212_22.html. Accessed by May 2013]. Arikan OA, Rice C, Codling E. Occurrence of antibiotics and hormones in a major agricultural watershed. Desalination 2008;226:121–33. Ashton D, Hilton M, Thomas KV. Investigating the environmental transport of human pharmaceuticals to streams in the United Kingdom. Sci Total Environ 2004;333: 167–84. Behera SK, Kim HW, Oh JE, Park HS. Occurrence and removal of antibiotics, hormones and several other pharmaceuticals in wastewater treatment plants of the largest industrial city of Korea. Sci Total Environ 2011;409:4351–60. Ben W, Qiang Z, Adams C, Zhang H, Chen L. Simultaneous determination of sulfonamides, tetracyclines and tiamulin in swine wastewater by solid-phase extraction and liquid chromatography-mass spectrometry. J Chromatogr A 2008;1202:173–80. Besse JP, Garric J. Human pharmaceuticals in surface waters. Toxicol Lett 2008;176: 104–23. BosiData (Bosi Data Research Center). 2010–2013 personal care product market of China and investment prospect analysis; 2011. Boyd GR, Palmeri JM, Zhang S, Grimm DA. Pharmaceuticals and personal care products (PPCPs) and endocrine disrupting chemicals (EDCs) in storm water canals and Bayou St. John in New Orleans, Louisiana, USA. Sci Total Environ 2004;333:137–48. Brausch JM, Rand GM. A review of personal care products in the aquatic environment: environmental concentrations and toxicity. Chemosphere 2011;82:1518–32.

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