Endocrine disrupting compounds in municipal and industrial wastewater treatment plants in Northern Greece

Chemosphere 73 (2008) 1716–1723

Contents lists available at ScienceDirect

Chemosphere journal homepage: www.elsevier.com/locate/chemosphere

Endocrine disrupting compounds in municipal and industrial wastewater treatment plants in Northern Greece Paraskevi Pothitou, Dimitra Voutsa * Environmental Pollution Control Laboratory, Chemistry Department, Aristotle University of Thessaloniki, 541 24 Thessaloniki, Greece

a r t i c l e

i n f o

Article history: Received 10 July 2008 Received in revised form 9 September 2008 Accepted 10 September 2008 Available online 26 October 2008 Keywords: Alkylphenols Estrogens Mass spectrometry Sewage effluents Tannery Textile

a b s t r a c t The occurrence and fate of endocrine disrupting compounds (EDCs) in a sewage treatment plant and two industrial wastewater treatment plants from textile and tannery factories were investigated. EDCs of interest are 4-nonylphenol, 4-octylphenol, their ethoxylate oligomers (mono- and di-ethoxylates of nonylphenol and octylphenol), bisphenol A, triclosan and steroid estrogens. Target compounds were determined in dissolved fraction, total suspended solids and sludge by employing solid phase extraction and ultrasonication followed by gas chromatography–mass spectrometry. Nonylphenols and oligomers with one or two ethoxy groups were the most abundant compounds in raw wastewater as well as in effluents from all the treatment stages of sewage treatment plant, followed by triclosan and bisphenol A. Steroids were found at very low concentrations. Almost all phenolic EDCs compounds were predominantly associated to suspended solids in influents whereas the dissolved fraction dominated the treated effluents. High removal rates, ranging from 86% to 99%, were observed throughout the whole treatment process. Biodegradation was the main removal pathway of EDCs. Tannery wastewaters exhibited high concentrations of nonylphenolic compounds. This type of wastewaters could pose a significant risk to the aquatic and terrestrial environment. Ó 2008 Elsevier Ltd. All rights reserved.

1. Introduction The presence of endocrine disrupting compounds (EDCs) in the water cycle (wastewaters-aquatic systems-drinking water) is considered as major environmental issue. EDCs refer to various ‘‘exogenous substances that cause adverse effects in an intact organism, or its progeny, consequent to changes in endocrine function” and include diverse groups of heterogeneous contaminants (alkylphenols, polychlorinated biphenyls, selected pesticides, steroid sex hormones, phthalates etc.) (EC, 1997; Metzler and Pfeiffer, 2001). These compounds often occur in domestic effluents, industrial wastewaters, landfill effluents and livestock wastes that are considered as main sources of EDCs to the aquatic environment (Ahel et al., 1994; Staples et al., 1998; Ying et al., 2002a,b; Voutsa et al., 2006). The occurrence of selected EDCs (alkylphenols and their ethoxylate oligomers, bisphenol A and steroids) in the major coastal area of Thessaloniki, inland waters and marine waters, has been recently investigated (Arditsoglou and Voutsa, 2008a,b). Alkylphenols (APs) and alkylphenol ethoxylates (APEOs) are used in industrial, agricultural and household applications as detergents, emulsifiers, wetting agents, dispersants or solubilizers (Ying * Corresponding author. Tel.: +30 2310 997858; fax: +30 2310 997747. E-mail address: [email protected] (D. Voutsa). 0045-6535/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2008.09.037

et al., 2002a). The potential adverse effects of APEOs and APs to humans and other organisms resulted in reduction of their use in several countries. Since 2005, there is a restriction on the sale and use of products that contain more than 0.1% of 4-nonylphenol ethoxylates or 4-nonylphenols (EC, 2003). Nonylphenols have been identified as priority hazardous substances whereas octylphenols are subject for inclusion in this category and environmental quality criteria for surface waters have been set (EC, 2001; EC, 2006). Moreover, in the 3rd draft of the working paper for sludge, a limit concentration for nonylphenol and nonylphenolethoxylates with 1 or 2 ethoxy groups has been proposed for the application of sewage sludge onto agricultural land (EC, 2000). Bisphenol A is used in the production of polycarbonate and epoxy resins, as stabilizing agent in plastics, as antioxidant in tire production and as basic chemical in the production of certain flame retardants (Staples et al., 1998). Triclosan is an important bactericide used in various personal care and consumer products (Ying and Kookana, 2007). Natural or synthetic steroids are excreted by mammals and eventually, they occur in domestic effluents and in livestock waste (Ying et al., 2002b). In the study area, Northern Greece, there are many mediumscale wastewater treatment plants with specific operational characteristics that service small municipalities and have never been thoroughly investigated. Moreover, although there are many small to medium sized industrial factories such as textiles and tanneries in the major area of Thessaloniki there is no information about the

1717

P. Pothitou, D. Voutsa / Chemosphere 73 (2008) 1716–1723

fate of EDCs in these type of wastewaters. The aim of this study was to investigate the occurrence and fate of EDCs in three wastewater treatment plants: one treating municipal wastewater and the other two treat industrial wastewater from textile and tannery factories. For this purpose a method was developed for the simultaneous determination of EDCs (4-nonylphenol, 4-octylphenol, mono- and di-ethoxylate of nonylphenol, mono- and di-ethoxylate of octylphenol, bisphenol A, triclosan, estrone, 17a-estradiol, 17bestradiol and estriol) by employing gas chromatography coupled to ion trap mass analyzer. Target compounds were determined in dissolved fraction of wastewater, in suspended particles and sludge. The distribution of EDCs between dissolved phase and particulate matter in various steps of treatment and the removal efficiency of studied plants are discussed. 2. Materials and methods 2.1. Sampling The studied sewage treatment plant (STP) is a small unit representative of those usually covering the treatment needs of small to medium municipalities. This plant receives 1000 m3 d1 domestic sewage wastewater from the major area of Kallikratia and a small input from local industries. The hydraulic load reach up to 2500 m3 d1 due to increase of population during summer period. About 87% of the domestic wastes are from the sewage distribution system of three small municipalities, the rest are wastewaters from house septic tanks. The STP includes screening, receiving/storage tank (500 m3) with 12 h residence time, conventional activated sludge treatment (aeration tank 2650 m3, 64 h, secondary sedimentation tank 530 m3, 13 h) and disinfection of effluent using NaOCl (200 m3, 5 h). The effluents drain through a 1 km pipe to a small stream and are discharged after 2.5 km to the sea. About 180 kg d1 sludge is produced, dewatered and finally deposited in drying beds, close to the plant. Five samplings campaigns were conducted in the STP from April to June 2007. Grab samples were collected between 9:00 and 11:00 p.m. from all stages of the treatment and post-treatment effluent: influent-raw wastewater (IN), aeration (AE), secondary sedimentation (SE), final effluent (EF) and sewage sludge (SS). The samples were collected in brown glass vessels precleaned with acetone and n-hexane. The samples kept cool after sampling and during transport to the laboratory and stored at <4 °C prior to extraction. Samples from influent (IN) and effluent (EF) of two wastewater treatment plants (WWTPs) that receive wastewaters from textile (TEX) and tannery (TAN) factories, respectively, were also collected. The TEX treatment plant aims mainly at the discoloration of the wastewaters through coagulation and sedimentation processes. The TAN treatment plant receives three different wastewater lines, sulfur-containing wastes, chromium-containing wastes and wastes from others activities in tanneries. Treatment involves

chromium precipitation as chromium trioxide followed by conventional activated sludge process. The physicochemical parameters (temperature, pH and conductivity) in the wastewaters were measured during sampling and are shown in Table 1. 2.2. Sample preparation The sample processing and analysis is described in details elsewhere (Arditsoglou and Voutsa, 2008a,b). Briefly, wastewater samples were filtered through 0.7 lm glass fiber filters (GF/F, Whatman) to separate total suspended solids (TSS). EDCs from filtrates were recovered after solid phase extraction through OASIS HLB cartridges (200 mg, 6 mL, Waters). The cartridges were placed on a vacuum manifold and conditioned sequentially with 5 mL of acetone, 5 mL of methanol and 3  5 mL of ultrapure water at a flow rate of 1 mL min1. Then the filtrates (500–1000 mL) were percolated through the cartridges at a flow rate of 8 mL min1. The cartridges were dried under vacuum and EDCs were eluted with 2  5 mL of acetone. The eluates were reduced to 0.5 mL with rotary evaporation and until dryness under a gentle stream of nitrogen. The dried samples were submitted to the derivatization procedure. EDCs from loaded filters (dried at 40 °C) and freeze-dried sludge samples (0.5 g) were recovered after sonication with 20 mL acetone-methanol (1:1) for 20 min, in triplicate. A clean up step through florisil cartridges (500 mg, 4 mL, Alltech) was necessary for sludge samples. The eluates were reduced to 0.5 mL with rotary evaporation and until dryness under a gentle stream of nitrogen. The dried samples were submitted to the derivatization procedure. The derivatization reaction was performed by addition of 100 lL BSTFA at 70 °C for 30 min. The silylated derivatives were allowed to cool to room temperature, transferred to autosampler vials and analysed. Surrogate internal standards (BPA-d16 and bT2-d2) were added in all samples before extraction whereas instrumental performance internal standards (nNP and nBP) were added to the samples extracts before derivatization. Calibration standard solutions contained also the same amount of internal standards (100 ng of each compound) were added in environmental samples. 2.3. Analytical determination The analysis of silylated derivatives was performed using a gas chromatograph (Trace GC ultra, Thermo Finnigan Electron Corporation) coupled with an ion trap mass spectrometer (Polaris Q, Thermo Finnigan) and an autosampler (AI 3000, Thermo Finnigan Electron Corporation). Compounds were separated on a Rtx–5MS Crossbond 5% diphenyl-95% dimethyl polysiloxane capillary column (30 m length, 0.25 mm i.d., 0.25 lm film thickness) with a capillary precolumn Rtx–5MS (7 m length, 0.32 mm i.d), from Thames Restek UK Ltd. Samples were injected (1 lL) in a

Table 1 Physicochemical parameters of wastewaters during sampling period (mean values) Parameters

T (°C) EC (mS cm1) pH DOC (mg L1) ABS254 (cm1) TSS (mg L1) POC (mg g1) TOCa (mg g1) a

Textile wastewaters (TEX)

Tannery wastewaters (TAN)

IN

Sewage wastewaters (STP) AE

SE

EF

IN

EF

IN

EF

22.3 1.98 7.41 50 0.94 205 46

22.7 1.63 7.37 15 0.22 1676 37

22.7 1.62 7.46 14 0.20 15 26

22.1 1.64 7.76 16 0.20 4.2 24 20

28.1 7.35 9.23 81 0.99 92 28

23.9 4.50 8.12 13 0.23 84 23

24.0 18.1 7.41 765 3.70 1566 20

25.0 19.5 7.72 430 3.30 1082 18

Organic content of sludge.

1718

P. Pothitou, D. Voutsa / Chemosphere 73 (2008) 1716–1723

PTV-injector at splitless mode. The silylated compounds were separated using the following oven program: the column temperature was initially set at 60 °C for 1.5 min, then increased at a rate of 20 °C min1 up to 180 °C, 5 °C min1 up to 250 °C, 20 °C min1 up to final temperature of 300 °C, which was maintained for 6 min. Helium carrier gas (99.99% purity) was maintained at a constant rate of 1.5 mL min1. The temperature at the injector was 280 °C. The ion source and transfer line temperature was set at 200 °C and at 300 °C, respectively. Mass spectra were obtained using electron impact ionization (70 eV). After 7.5 min of solvent delay, the ion trap was operated using MS-SIM scan up to 20 min for determination of alkylphenols followed by multiple reaction monitoring MS/MS (20–30 min) for steroids. Under these conditions the following compounds were determined: 4-n-butylphenol (nBP), 4-toctylphenol (tOP), 4-nonylphenol (NP), 4-n-octylphenol (nOP), 4-n-nonylphenol (nNP), 4-octylphenol monoethoxylate (OP1EO), 4-nonylphenol monoethoxylate (NP1EO), triclosan (TCS), bisphenol A-d16 (BPA-d16), bisphenol A (BPA), 4-octylphenol diethoxylate (OP2EO), 4-nonylphenol diethoxylate (NP2EO), estrone (E1), 17aestradiol (aE2), 17b-estradiol (bE2), 17b-estradiol-d2 (bE2-d2) and estriol (E3) (Table 2). 2.4. Identification–quantification The identification of silylated derivatives of the analytes was based on relative retention time, the presence of target ions and their relative abundance. Two to three ions were monitored in MS-SIM mode and one precursor and two/three daughter ions were monitored in MS/MS mode (Table 2). The criteria concerning the performance of analytical methods reported in Commission Decision 2002/657/EC were followed in this study (EC, 2002). The relative retention time of the analyte to that of the internal standard corresponded to retention time of calibration standard solutions at a tolerance of ±0.5%. The maximum permitted tolerances for relative ion intensities using EI-GC–MS ranged from ±10% for ions with relative intensity >50% of the base peak up to ±50% for

ions with relative intensity <10%. For EI-GC–MS/MS the criteria ranged from ±20% for peaks with relative intensity >50% up to ±50% for peaks with relative intensity 610%. The quantification of OP, OP1EO, OP2EO, BPA and TCS was carried out by calculating the relative response factors based on the area of the surrogate standard BPA-d16. NP, NP1EO and NP2EO were quantified comparing the integrated peak area of the summed selected ions within a retention time window (8 peaks) with the peak area of BPA-d16. Steroids were quantified by calculating the relative response factors based on the area of the surrogate standard bE2-d2. 2.5. Quality assurance The linearity of the method was tested with calibration standards at seven concentration levels. A linear fit with high correlation coefficient was obtained for the studied compounds (R2 > 0.998). The performance internal standards were used to monitor the instrument response for each sample. Any sample or calibration solution which had responses outside one standard deviation of the average was rejected. Recovery tests were conducted on wastewater and particulate matter samples spiked with target compounds at three concentration levels (70, 100 and 150 ng). The recoveries for all EDCs ranged from 72% to 115%. E3 and bE2 showed lower recoveries (53%), similarly to the results reported by other investigators (Shareef et al., 2006; Arditsoglou and Voutsa, 2008a). Procedural blanks and field blanks were analyzed regularly. 2.6. Determination of organic carbon Dissolved organic carbon (DOC) in filtrates was measured by TOC–VCSH/SCN Total Carbon Analyzer (Shimadzu). Organic carbon in suspended solids (POC) and in sludge (TOC) was measured by TOC–VSERIES SSM–5000A Solid Sample Module for Total Carbon Analyzer (Shimadzu). UV absorbance of filtrates was measured at

Table 2 Identification and quantification of the silylated derivatives of the target compounds Molecular mass

Retention time (min)

Target ions (Abundance)b

4-n-Butylphenol 4-t-Octylphenol 4-Nonylphenol

nBP tOP NP

a

150.22 206.33 220.35

7.62 9.37 10.09–10.86

4-n-Octylphenol 4-n-Nonylphenol

nOP nNPa

206.33 220.35

11.20 12.43

4-Octylphenol monoethoxylate 4-Nonylphenol monoethoxylate Triclosan

OP1EO NP1EO TCS

250.38 264.41 289.54

13.06 14.32–15.02 15.57

Bisphenol A-d16 Bisphenol A 4-Octylphenol diethoxylate 4-Nonylphenol diethoxylate

BPA-d16a BPA OP2EO NP2EO

244.38 228.29 294.44 308.47

16.60 16.73 17.20 18.53–19.29

179.1 207.2 179.1 207.1 207.0 179.2 292.2 251.1 251.0 200.0 362.0 197.2 357.2 295.1 294.9

Phenolic compounds

Steroid compounds

179.1 207.2 Sum of 8 peaks 278.1 179.2

252.1 265.1 360.0

(1.7) (60.2) (37.8)

251.1 Sum of 8 peaks 200.0

368.3 358.2 296.2 308.9

(100) (33.3) (1.4) (52.2)

368.3 357.2 295.1 Sum of 8 peaks

Precursor

Confirmation ions (abundance)c 285.2 326.2 244.2 342.1 287.2 328.2 285.2 326.2 311.2 386.2

272.39

22.73

416.2

Estrone

E1

270.37

22.78

342.2

17b-Estradiol-d2

bE2-d2a

274.39

23.08

418.0

17b-Estradiol

bE2

272.39

23.14

416.2

Estriol

E3

288.39

24.58

414.2

c

(41.4) (18.7) (48.2) (47.8) (100) (17.7)

Retention time (min)

aE2

a

222.1 208.2 193.1 221.1 278.1 180.2

Molecular mass

17a-Estradiol

b

(100) (100) (41.6) (100) (7.7) (100) (33.6) (100) (100) (100) (39.4) (3.4) (100) (100) (100)

Quantitation ion

Internal standards. GC–MS. GC–MS/MS, Quantitation ions are in bold.

(86.5) (53.4) (22.1) (100) (72.7) (42.4) (85.9) (55.6) (81.5) (100)

298.2 416.2 257.2

(17.9) (100) (51.4)

300.0 418.0 298.2 416.2 324.2 414.2

(17.4) (100) (18.8) (100) (87.6) (86.8)

1719

P. Pothitou, D. Voutsa / Chemosphere 73 (2008) 1716–1723

254 nm by UV–Vis meter, U-2001 (Hitachi). The organic content of samples and UV absorbance are shown in Table 1.

3. Results and discussion 3.1. Sewage treatment plant 3.1.1. Dissolved concentrations Dissolved (ng L1) concentrations of EDCs along the STP are shown in Table 3. Alkylphenols (NP and tOP), their ethoxylate oligomers (NP1EO, NP2EO, OP1EO and OP2EO), bisphenol A (BPA) and triclosan (TCS) were determined along water treatment from the raw wastewater up to final effluent, in effluents from all stages of the treatment. The concentrations of EDCs decreased significantly in secondary effluent (AE) relatively to influent (IN), whereas there were not differences between secondary effluent and effluent from clarifier (SE). These findings reveal that bioreactor is a crucial step in removing these compounds, whereas clarifier, although has an important role in settling out the sludge, does not have significant contribution to the removal of target compounds (Tan et al., 2007). Nonylphenol and its ethoxylates were the most abundant compounds. Their concentrations in influent ranged from 545 to 3022 ng L1 for NP, 466–4025 ng L1 for NP1EO and 490– 2670 ng L1 for NP2EO. Lower concentrations were determined in effluents ranged from 126 to 1965 ng L1 for NP, 13–573 ng L1 for NP1EO and 26–216 ng L1 for NP2EO. The high concentrations of NP occasionally observed in effluents could be attributed to inefficiency of the plant due to the high loads of nonylphenolic compounds those days, as well as to the formation of NP from the degradation of long chain ethoxylated compounds (Ahel et al., 1994). Octylphenols (tOP, OP1EO and OP2EO) were found at lower concentrations in raw wastewaters and eventually in effluents

(10–82 ng L1 for tOP, <5–40 ng L1 for OP1EO and <5–69 ng L1 for OP2EO) due to their lower use in the commercial products (10% for OPEOs vs 90% for NPEOs). Wide variation in concentrations of nonylphenols and octylphenols in effluents from various treatment plants have been reported by other investigators (57– 3210 ng L1 for NP, 5–470 ng L1 for OP, 69–1800 ng L1 for NP1EO and 42–830 ng L1 for NP2EO) (Lee et al., 2005; Johnson et al., 2005; Clara et al., 2007; Tan et al., 2007). The differences among treatment plants could be attributed to the differences in input loads, treatment technologies and environmental conditions (Johnson et al., 2005; Tan et al., 2007; Cirja et al., 2008). Environmental quality criteria for nonylphenol and octylphenol in surface waters (annual averages: 0.3 lg L1 and 0.1 lg L1, respectively) have been recently proposed (EC, 2006). In order to minimize the risk of emissions of EDCs to the receiving waters according to the proposed environmental quality criteria, a 3–5-fold dilution of the effluent is necessary. The concentrations of BPA ranged from 468 to 857 ng L1 in raw wastewaters and 20–48 ng L1 in effluents. These concentrations laid to the lower values reported in literature (<1–2847 ng L1 in influent and 50–450 ng L1 in effluents) (Lee et al., 2005; Tan et al., 2007). The concentrations of TCS ranged from 304 to 760 ng L1 in raw wastewaters and 15–290 ng L1 in effluents. Similar concentrations have been reported by other investigators. Lee et al. (2005) reported TCS concentrations ranged from 870 to 1830 ng L1 in influents and 50–360 ng L1 in effluents from STPs in Ontario, Canada. Concentrations ranged from 23 to 434 ng L1 were determined in effluents from WWTPs in Australia (Ying and Kookana, 2007). Spearman correlation analysis was employed in the dataset of dissolved concentrations of phenolic EDCs. Significant correlation coefficients were observed among the compounds NP-NP1EONP2EO-TCS-BPA (0.711–0.890, p = 0.01). Also significant correlation (0.492–0.583, p = 0.05) was observed between DOC and

Table 3 Dissolved and particulate concentrations of EDCs at various treatment stages of sewage treatment plant (mean values ± sd, n = 5) Treatment stages

IN (ng L1)

AE (ng L1)

SE (ng L1)

EF (ng L1)

Dissolved tOP NP nOP OP1EO NP1EO TCS BPA OP2EO NP2EO aE2 T1 bE2 E3

94 ± 157 1574 ± 1063 BDL 33.9 ± 21.0 2224 ± 1772 445 ± 181 676 ± 151 402 ± 355 1479 ± 1093 BDL BDL BDL BDL

25 ± 13 352 ± 292 BDL 14.5 ± 13.1 83 ± 67 28.7 ± 27.8 42 ± 25 29 ± 21 111 ± 85 BDL BDL BDL BDL

21 ± 9 314 ± 182 BDL 11.2 ± 21.2 59 ± 40 35.3 ± 27.2 69 ± 51 32 ± 26 103 ± 68 BDL BDL BDL BDL

40 ± 31 786 ± 794 BDL 9.4 ± 17.0 154 ± 242 76 ± 121 33 ± 11 28 ± 27 87 ± 78 BDL BDL BDL BDL

Particulate tOP NP nOP OP1EO NP1EO TCS BPA OP2EO NP2EO aE2 T1 bE2 E3

IN (ng mg1)

AE (ng mg1)

SE (ng mg1)

EF (ng mg1)

Sludge (ng g1)

1.20 ± 0.61 160 ± 108 0.45 ± 0.32 2.51 ± 1.53 152 ± 94 26.8 ± 26.3 0.48 ± 0.71 0.70 ± 0.54 92 ± 65 BDL BDL BDL BDL

13.2 ± 16.0 3.8 ± 4.1 0.018 ± 0.022 0.015 ± 0.012 1.6 ± 1.8 0.472 ± 0.454 0.082 ± 0.097 0.013 ± 0.013 BDL BDL BDL BDL BDL

0.110 ± 0.047 BDL BDL 0.036 ± 0.025 BDL 0.067 ± 0.059 0.015 ± 0.010 0.010 ± 0.009 BDL BDL BDL BDL BDL

1.78 ± 3.26 9.5 ± 14 BDL 0.145 ± 0.146 0.84 ± 1.25 0.112 ± 0.056 0.40 ± 0.62 0.41 ± 0.44 BDL BDL BDL BDL BDL

179 ± 95 1089 ± 215 BDL 8.1 ± 5.9 1080 ± 371 461 ± 178 29.9 ± 10.6 16.1 ± 9.4 BDL BDL BDL BDL BDL

BDL: below detection limit. In dissolved phase: nOP < 7 ng L1, E1and E3 < 3 ng L1, aE2, aE2 < 2 ng L1. In particulate phase: NP, NP1EO, NP2EO < 13 ng g1, nOP < 10 ng g1, E1and E3 < 5 ng g1 and aE2, aE2 < 3 ng g1.

1720

P. Pothitou, D. Voutsa / Chemosphere 73 (2008) 1716–1723

NP-NP1EO-NP2EO-TCS showing that the presence of DOC could enhance the solubility of the less polar compounds. Low concentrations of steroid estrogens were found in the studied STP (E1, E3 < 3 ng L1 and aE2, bE2 < 2 ng L1). These compounds are usually determined in wastewaters at not detectable concentrations or in very low ng L1 range, although peak values have been reported (Ternes et al., 1999a; Braga et al., 2005; Johnson et al., 2005; Tan et al., 2007). In German WWTPs a median value of 9 ng L1 for E1 and <3 ng L1 for bE2 has been reported by Ternes et al. (1999a). Johnson et al. (2005) in a study of 16 STPs employing secondary treatment throughout Europe reported a median value of <3 ng L1 for E1, whereas E2 was only detected in the effluents of six plants at concentrations 0.7– 5.7 ng L1. Tan et al. (2007) found low concentrations of steroids in effluents from WWTPs (<1–42 ng L1 for E1 and <1–2 ng L1 for bE2), whereas Braga et al. (2005) reported concentrations <0.1 ng L1 for both E1 and E2. The low concentrations of steroids in our study could be attributed to their degradation by organisms present in the sewerage system during wastewater transport from the various domestic sources to STP as well as to the long residence time of wastewater to the reception/storage tank of STP (12 h) and during treatment processes (82 h). Suzuki and Maruyama (2006) reported that natural estrogens adsorb onto activated sludge and therefore are easily biodegraded. Ternes et al. (1999b) found that both bE2 and E1 are vanished in a control activated sludge batch within 5 h. Braga et al. (2005) also reported that biological treatment rapidly removes steroids whereas reverse osmosis and chlorination fully eliminate these compounds from the effluents. 3.1.2. Particulate and sludge concentrations The concentrations of EDCs in suspended solids and in sludge are shown in Table 3. Most of the target compounds were determined in suspended solids and sludge. The solids from influent exhibited the highest concentrations of EDCs. Nonylphenolic compounds are the dominant analytes in solids, similarly to the dissolved phase. The particulate concentrations of NP, NP1EO and NP2EO in influent were 160 ng mg1, 152 ng mg1 and 92 ng mg1, respectively. Low concentrations of BPA were found in solids (0.48 ng mg1 in influent and 0.40 ng mg1 in effluent) probably due to the hydrophilic nature of this compound. TCS was found at concentrations ranged from 26.8 ng mg1 in influent to 0.112 ng mg1 in effluent. There are few data about the concentrations of EDCs in suspended solids. Tan et al. (2007) reported particulate concentrations ranged from 839 to 9830 ng mg1 for NP and <1–271 ng mg1 for BPA in raw wastewaters. Isobe et al. (2001) found 59.0 ng mg1 of NP and 4.36 ng mg1 of OP in particulate phase of primary effluents and 9.52 ng mg1 of NP and 1.11 ng mg1 of OP in secondary effluents. NP and NP1EO exhibited high concentrations in sludge (1089 ng g1 and 1080 ng g1 respectively), followed by TCS (461 ng g1), tOP (179 ng g1) and BPA (30 ng g1). Wide variation regarding the concentrations of EDCs in sludge has been reported in literature. For NP, concentrations of 20.5–429 ng g1 found in Australian WWTPs and 137–470 lg g1 in STPs from Toronto (Lee and Peart, 1995; Tan et al., 2007). For BPA, concentrations up to 1750 ng g1 determined in Greek WWTPs and much lower 3.1–3.8 ng g1 in Australian WWTPs (Tan et al., 2007; Stasinakis et al., 2008). The concentrations of TCS in sludge ranged from 0.09 to 16.79 lg g1 in various WWTPs (Ying and Kookana, 2007; Stasinakis et al., 2008). In order to eliminate the possible risk from the application of sewage sludge onto agricultural land, the concentration of nonylphenol and nonylphenolethoxylates with 1 or 2 ethoxy groups must not exceed the limit value of 50 mg kg1. In the present study, the sum of NP, NP1EO and NP2EO in sludge was well below this criterion.

Steroids were not found at detectable concentrations in suspended solids or in sludge (E1, E3 < 5 ng g1 and aE2, bE2 < 3 ng g1). Tan et al. (2007) also reported concentrations below detection limits for steroids in particulate solids. Other investigators reported low concentrations of E1 and E2, typically <10 ng g1, in sludge, although peak values (up to 37 ng g1 and 49 ng g1, respectively) were also determined (Ternes et al., 2002; Andersen et al., 2003). The low concentrations of steroids could be attributed to the biodegradation process that predominately affects their fate in STPs, whereas sorption is not considered as important process. For a typical STP the removal of steroid estrogens with excess sludge was estimated to be only 1.5–1.8% (Andersen et al., 2005). Spearman correlation analysis was employed in the dataset of particulate (suspended solids and sludge) concentrations of EDCs. Significant correlation coefficients were found among the target compounds (0.555–0.832, p = 0.01). These compounds (except BPA, OP1EO, OP2EO) exhibited also significant correlation (0.430–0.668, p = 0.05) with organic content of samples. 3.1.3. Behavior of EDCs along STP The concentrations of EDCs in dissolved and particulate fractions (in ng L1) were considered for the estimation of distribution of target compounds between the two phases. The distribution pattern in different treatment steps of STP is illustrated in Fig. 1. Almost all the target compounds were predominately (>75%) associated with suspended solids in influent wastewater. However, BPA showed a significant dissolved fraction (85%) that could be attributed to the higher solubility of this compound and to the more polar character compared to the other compounds. As the wastewater being treated, the percentage of the target compounds in dissolved phase is increased and finally the dissolved fraction dominated the treated effluent. Similarly, Isobe et al. (2001) reported that the particulate phase of NP and OP in primary effluents represents 80% and 60%, respectively whereas in secondary effluents is reduced to 30% and 10%, respectively. The removal efficiencies of wastewater contaminants during the treatment process were calculated from the total concentration (sum of dissolved and particulate in ng L1) in influent (CIN) and effluent (CEF) in each treatment stage as follows:

Rð%Þ ¼ ðC IN  C EF Þ=C IN The removal rates of the target compounds are shown in Fig. 2. The removal rates ranged from 86% for tOP up to 99% for nOP, NP1EO and NP2EO. Substantial removal for all compounds observed in bioreactor (effluent AE), revealing the critical role of biological treatment for the removal of EDCs. The observed removal rates were similar to the higher values reported in literature. Removal ranges of 85–>99% for NP, 38–>99% for BPA and 72–93% for TCS have been reported by other investigators (Tan et al., 2007; Ying and Kookana, 2007). The removal of EDCs during wastewater treatment can be influenced by several factors such as the state of microorganisms, the acclimatization of biomass, the influent load, the hydraulic residence time and the environmental conditions (Tan et al., 2007; Ying and Kookana, 2007; Cirja et al., 2008). The main processes that affect the fate of a compound during treatment are sorption and biodegradation (Fauser et al., 2001). In order to assess the contribution of these processes during treatment, the mass load of EDCs that was lost due to sum of all transformation processes WLOST (g d1) was calculated using the following equation (Heidler and Halden, 2007):

W LOST ¼ ðQ IN  C IN Þ  ðQ EF  C EF Þ  W SLUDGE where QIN and QEF is the flow rate of influent and treated effluent (L d1) respectively, CIN and CEF is the total concentration in unfiltered influent and treated effluent (g L1), respectively and WSLUDGE is the mass output in sludge (g d1). CIN and CEF were calculated as

1721

P. Pothitou, D. Voutsa / Chemosphere 73 (2008) 1716–1723

IN 100

PAR

(%)

75 DIS 50 25 0 tOP

NP

nOP

OP1EO NP1EO

TCS

BPA

OP2EO NP2EO

AE

(%)

100 75

PAR

50

DIS

25 0 tOP

NP

nOP

OP1EO NP1EO

TCS

BPA

OP2EO NP2EO

SE 100 PAR

(%)

75 DIS

50 25 0 tOP

NP

nOP

OP1EO NP1EO

TCS

BPA

OP2EO NP2EO

EF 100

PAR

(%)

75 DIS

50 25 0 tOP

NP

nOP

OP1EO NP1EO

TCS

BPA

OP2EO NP2EO

Fig. 1. Distribution of EDCs between dissolved and particulate phase in various treatment stages of STP.

100

Removal ( %)

75 AER 50

TOT

25

0 tOP

NP

nOP

OP1EO NP1EO

TCS

BPA

OP2EO NP2EO

Fig. 2. Removal (%) of EDCs in sewage treatment plant (AER: removal in aeration tank, TOT: overall removal).

the sum of concentrations in dissolved and particulate phase of influent and effluent, respectively. The fate of EDCs in STP is shown in Fig. 3. Most of the compounds are removed in bioreactor (78– >99%) and only a small amount is sorbed in sludge (0.5–9%). Clara et al. (2007) also reported high biotransformation rates in WWTPs, with more than 85% of overall removal rates of NP, NP1EO and NP2EO. Biological degradation is also considered as predominant removal mechanism for TCS, although adsorption onto sludge plays a significant role (Stasinakis et al., 2007; Ying and Kookana, 2007).

However, in real full scale WWTPs almost equal contribution of transformation process and sorption to sludge (45%) during treatment has been reported and was attributed to the particle settling in primary clarifiers (Heidler and Halden, 2007; Stasinakis et al., 2008). In this study, high degradation rates can be attributed to the absence of primary treatment in STP. In addition, high temperatures during sampling period (22–30 °C), high residence times in STP (94 h) and relative low loads enhanced the transformation/ degradation of EDCs (Ahel et al., 1994; Clara et al., 2007).

1722

P. Pothitou, D. Voutsa / Chemosphere 73 (2008) 1716–1723

Mass balance EDCs

100 WLOST

WSLUDGE

90

WEF 80

70 10 0 tOP

NP

nOP

OP1EO NP1EO

TCS

BPA

OP2EO NP2EO

Fig. 3. Fate of EDCs in sewage treatment plant.

3.2. Industrial wastewater treatment plants 3.2.1. Dissolved and particulate concentrations The occurrence of EDCs was also examined in two wastewater treatment plants located in the industrial area of Thessaloniki; plant TEX treats wastewaters from a textile industry and plant TAN treats tannery wastewaters. In textile/tannery activities NPEOs are applied in auxiliaries formulations used for pretreatment operations or in additives as detergents or wetting agents in wool scouring, tanning, hydrogen peroxide bleaching and dyeing process (Loos et al., 2007). The dissolved and particulate concentrations of EDCs in influents and effluents are shown in Table 4. Nonylphenol and its ethoxylates were the most abundant compounds in influents as well as in effluents. The concentrations of dissolved nonylphenolic compounds in textile industry effluents were 0.54 lg L1 for NP, 0.71 lg L1 for NP1EO and 0.61 lg L1 for NP2EO. Loos et al. (2007) reported similar concentrations in effluent samples from textile WWTPs and receiving waters (0.25–2.5 lg L1 for NP, 0.022–2.1 lg L1 for NP2EO). Wastewater from tannery industries exhibited high concentrations of dissolved nonylphenolic compounds (433 lg L1 for NP, 167 lg L1 for NP1EO and 52 lg L1 for NP2EO in effluents). These compounds exhibited also high concentrations in particulate phase (6.2 lg mg1 for NP, 0.91 lg mg1 for NP1EO and 0.26 lg mg1 for NP2EO in effluents). According to our knowledge there are no available literature data for comparison. The distribution of EDCs between dissolved and particulate phase, showed that EDCs in textile wastewaters occurred mainly in dissolved phase (>80% in effluents). On the contrary, particulate

Table 4 Dissolved and particulate concentrations of EDCs in industrial wastewaters (n = 2) Compound

tOP NP nOP OP1EO NP1EO TCS BPA OP2EO NP2EO

Textile (TEX)

Tannery (TAN)

Dissolved (lg L1)

Particulate (ng mg1)

Dissolved (lg L1)

IN

EF

IN

EF

IN

EF

IN

EF

0.073 1.16 BDL 0.013 7.9 0.085 0.128 BDL 10.5

0.051 0.54 BDL 0.075 0.71 0.082 0.035 0.022 0.61

0.167 5.2 BDL BDL 15 0.360 0.047 BDL 56

0.37 4.1 BDL BDL 18 0.54 0.005 0.029 BDL

0.004 65 BDL 0.035 15.4 0.188 0.83 0.068 12.5

0.004 433 BDL BDL 167 0.025 1.00 0.20 52

0.71 7915 BDL BDL 335 0.124 0.069 0.039 162

0.72 6218 BDL BDL 907 0.104 0.154 0.063 263

BDL: below detection limits.

Particulate (ng mg1)

EDCs represent significant contribution in tannery wastewater probably due to high concentrations of TSS (Table 1). The particulate fraction of nonyl phenolic compounds dominated (>90%) the tannery effluents. Lower particulate fractions were found for TCS (40%) and BPA (20%). 3.2.2. Removal efficiency The removal efficiency of the textile WWTP is considered satisfactory. The removal rates of nonylphenolic compounds and BPA ranged from 66% to 97%. On the contrary, in tannery WWTP most of the studied compounds exhibited higher concentrations in effluent than in influent. This plant operated in full scale during last year and there are still significant operational problems. This type of wastewaters (high loads of organic content, of nonylphenolic compounds, fats and oils) require longer period for acclimatization of microorganisms (Cirja et al., 2008). During treatment processes long chain NPEOs degrade in compounds with lower number of ethoxy-groups (NP1EO, NP2EO) and nonylphenols (Ahel et al., 1994). Thus, besides the operational problems, the high concentrations of NP, NP1EO, NP2EO in effluents suggest that their formation rate exceeds the removal rate. The concentrations of nonylphenols in tannery effluents are many-folds above the proposed environmental quality criteria (EC, 2000; EC, 2006). Eventually, these wastewaters could pose a significant environmental risk if they are discharged directly to the aquatic and terrestrial environment. The formation of more estrogenic compounds NP, NP1EO, NP2EO from the long chain NPEOs during treatment, resulted in higher estrogenic activity of effluents. The estrogen equivalent concentrations (EEQ) were calculated from the equation: EEQ = Ci  EEFi where, Ci is the concentration of i compound and EEFi is the estrogenic equivalency factor (Tan et al., 2007). The EEQs showed an increase in estrogenic activity from 41 ng L1 in influent to 270 ng L1 in effluent wastewaters. These effluents are lead for further treatment to the centralized treatment plant of industrial area and it is possible to affect the treatment efficiency of the central plant. Thus, nonylphenolic compounds should be among the parameters regularly determined in this type of wastewaters. 4. Conclusions The occurrence and fate of phenolic and steroid EDCs was studied in a medium scale sewage treatment plant, as well as in industrial wastewater from textile and tannery treatment plants. The target compounds were determined in dissolved phase, suspended solids and sludge by employing gas chromatography coupled with ion trap mass spectrometry.

P. Pothitou, D. Voutsa / Chemosphere 73 (2008) 1716–1723

Phenolic compounds were determined in dissolved and particulate phase of wastewaters after various treatment steps along STP. The removal rates of the studied compounds ranged from 86% to 99%, with bioreactor being the critical treatment step. In order to minimize the risk of the emissions of EDCs to the receiving waters, a 3–5-fold dilution of the effluent is necessary. Sewage sludge fulfils the criterion for nonylphenols, thus possible application onto agricultural land do not pose significant risk. Nonylphenolic compounds were the most abundant compounds in textile and tannery wastewaters. However, tannery effluents exhibited very high concentrations and could be considered as significant source of these compounds to the aquatic and terrestrial environment. The occurrence of nonylphenol in both urban and industrial wastewaters results in an increase risk to the receiving surface waters thus this compound should be among the parameters regularly determined in wastewaters. Acknowledgement The instrumentation used in this study was obtained through the Project AKMON 5 co-financed by E.U. – European Regional Development Fund (70%) and the Greek Ministry of Development – GSRT (30%). References Ahel, M., Giger, W., Koch, M., 1994. Behaviour of alkylphenol polyethoxylate surfactants in the aquatic environment: I. Occurrence and transformation in sewage treatment. Water Res. 28, 1131–1142. Andersen, H.R., Siegrist, H., Halling-Sorensen, B., Ternes, T.A., 2003. Fate of estrogens in a municipal sewage treatment plant. Environ. Sci. Technol. 37, 4021–4026. Andersen, H.R., Hansen, M., Kjolholt, J., Stuer-Lauridsen, F., Ternes, T., HallingSorensen, B., 2005. Assessment of the importance of sorption for steroid estrogens removal during activated sludge treatment. Chemosphere 61, 139– 146. Arditsoglou, A., Voutsa, D., 2008a. Determination of phenolic and steroid endocrine disrupting compounds in environmental matrices. Environ. Sci. Pollut. R 15, 228–236. Arditsoglou, A., Voutsa, D., 2008b. Passive sampling of selected endocrine disrupting compounds using polar organic chemical integrative samplers. Environ. Pollut.. doi:10.1016/j.-envpol.2008.02.007. Braga, O., Smythe, G.A., Schafer, A.I., Feitz, A.J., 2005. Fate of steroids in Australian inland and coastal wastewater treatment plants. Environ. Sci. Technol. 39, 3351–3358. Clara, M., Scharf, S., Scheffknecht, C., Gans, O., 2007. Occurrence of selected surfactants in untreated and treated sewage. Water Res. 41, 4339–4348. Cirja, M., Ivashechkin, P., Schäffer, A., Corvini, P.F.X., 2008. Factors affecting the removal of organic micropollutants from wastewater in conventional treatment plants (CTP)and membrane bioreactors (MBR), review article. Rev. Environ. Sci. Biotechnol. 7, 61–78. EC, European Commission, 1997. European Workshop on the Impact of Endocrine Disrupters in Human Health and the Environment. Environment and Climate Research Programme, DG XII, Report EUR 17549, 1997. EC, European Commission, 2000. Working document on sludge, 3rd draft, Brussels 27 April 2000, ENV.E.3/LM. EC, European Commission, 2001. Decision No 2455/2001/EC of the European Parliament and of the Council of 20 November 2001 establishing the list of priority substances in the field of water policy and amending Directive 2000/60/ EC. EC, European Commission, 2002. Commission Decision 2002/657//EC implementing Council Directive 96/23/EC concerning the performance of analytical methods and the interpretation of results. EC, European Commission, 2003. Directive 2003/53/EC of the European Parliament and of the council of 18 June 2003 amending for the 26th time Council Directive 76/769/EEC relating to restrictions on the marketing and use of certain

1723

dangerous substances and preparations (nonylphenol, nonylphenol ethoxylate and cement). EC, European Commission, 2006. Proposal for a Directive of the European Parliament and of the Council on environmental quality standards in the field of waters policy and amending Directive 2000/60/EC, 2006/0129, Brussels. Fauser, P., Sörensen, P.B., Carlsen, L., Vikelsöe, J., 2001. Phthalates and nonylphenols and LAS in Rosklide wastewater treatment plant. Fate and modelling based on measured concentrations in wastewater and sludge, NERI Technical Report No 354. Heidler, J., Halden, R.U., 2007. Mass balance assessment of triclosan removal during conventional sewage treatment. Chemosphere 66, 362–369. Isobe, T., Nishiyama, H., Nakashima, A., Takada, H., 2001. Distribution and behavior of nonylphenol, octylphenol, and nonylphenol monoethoxylate in Tokyo Metropolitan area: thir association with aquatic particles and sedimentary distributions. Environ. Sci. Technol. 35, 1041–1049. Johnson, A.C., Aerni, H.R., Gerritsen, A., Gibert, M., Giger, W., Hylland, K., Jürgens, M., Nakari, T., Pickering, A., Suter, M.J., Svenson, A., Wettstein, F.E., 2005. Comparing steroid estrogen, and nonylphenol content across a range of European sewage plants with different treatment and management practices. Water Res. 39, 47– 58. Lee, H.-B., Peart, T.E., 1995. Determination of 4-nonylphenol in effluent and sludge from sewage treatment plants. Anal. Chem. 67, 1976–1980. Lee, H.-B., Peart, T.E., Svoboda, M.L., 2005. Determination of endocrine-disrupting phenols, acidic pharmaceuticals, and personal-care products in sewage by solid-phase extraction and gas chromatography-mass spectrometry. J. Chromatogr. A 1094, 122–129. Loos, R., Hanke, G., Umlauf, G., Eisenreich, S.J., 2007. LC/MS/MS analysis and occurrence of octyl- and nonylphenol, their ethoxylates and their carboxylates in Belgian and Italian textile industry, waste water treatment plant effluents and surface waters. Chemosphere 66, 690–699. Metzler, M., Pfeiffer, E., 2001. Chemistry of natural and anthropogenic endocrine active compounds. In: Metzler, M. (Ed.), Endocrine Disruptors, Part I. The Handbook of Environmental Chemistry. Springer-Verlag, Berlin, Heidelberg, pp. 63–80. Shareef, A., Angove, M.J., Ewlss, J.D., 2006. Optimization of silylation using N -methyl-N-(trimethylsilyl)-trifluoroacetamide,N,O-bis(trimethylsilyl)-trifluoacetamide and N-(tert-butyl-dimethylsilyl)-N-methyltrifluoroacetamide for the determination of the estrogens estrone and 17ª-ethynylestradiol by gas chromatography-mass spectrometry. J. Chromatogr. A 1108, 121–128. Staples, A.C., Dorn, B.P., Klecka, M.G., O’Block, T.S., Harris, R.L., 1998. A review of the environmental fate, effects, and exposures of bisphenol A. Chemosphere 36, 2149–2173. Stasinakis, A.S., Gatidou, G., Mamais, D., Thomaidis, N.S., Lekkas, T.D., 2008. Occurrence and fate of endocrine disrupters in Greek sewage treatment plants. Water Res. 42, 1796–1804. Stasinakis, A.S., Petalas, A., Mamais, D., Thomaidis, N.S., Gatidou, G., Lekkas, T.D., 2007. Investigation of triclosan fate and toxicity in continuous-flow activated sludge systems. Chemosphere 68, 375–381. Suzuki, Y., Maruyama, T., 2006. Fate of natural estrogens in batch mixing experiments using municipal sewage and activated sludge. Water Res. 40, 1061–1069. Tan, B.L.L., Hawker, D.V., Müller, J.F., Leusch, F.D.L., Tremblay, L.A., Chapman, H.F., 2007. Comprehensive study of endocrine disrupting compounds using grab and passive sampling at selected wastewater treatment plants in South East Queensland. Australia Environ. Int. 33, 654–669. Ternes, T.A., Stumpf, M., Mueller, J., Haberer, K., Wilken, R.-D., Servos, M., 1999a. Behavior and occurrence of estrogens in municipal sewage treatment plants-I. Investigations in Germany, Canada and Brazil. Sci. Total Environ. 225, 81–90. Ternes, T.A., Kreckel, P., Mueller, J., 1999b. Behavior and occurrence of estrogens in municipal sewage treatment plants-II. Aerobic batch experiments with activated sludge. Sci. Total Environ. 225, 91–99. Ternes, T.A., Andersen, H., Gilberg, D., Bonerz, M., 2002. Determination of estrogens in sludge and sediments by liquid extraction and GC/MS/MS. Anal. Chem. 74, 3498–3504. Voutsa, D., Hartmann, P., Schaffer, C., Giger, W., 2006. Benzotriazols, alkylphenols and bisphenol A in municipal wastewaters and in the Glatt River, Switzerland. Environ. Sci. Pollut. R 13, 333–341. Ying, G.G., Williams, B., Kookana, R.S., 2002a. Environmental fate of alkylphenols and alkylphenol ethoxylates – a review. Environ. Int. 28, 215–226. Ying, G.G., Kookana, S.R., Ru, Y.J., 2002b. Occurrence and fate of hormone steroids in the environment. Environ. Int. 28, 545–551. Ying, G.G., Kookana, R.S., 2007. Triclosan in wastewaters and biosolids from Australian wastewater treatment plants. Environ. Int. 33, 199–205.