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Biological treatment of pharmaceutical wastewater from the antibiotics industry O. Lefebvre, X. Shi, C. H. Wu and H. Y. Ng
ABSTRACT Pharmaceutical wastewater generated by an antibiotics (penicillin) company was treated by aerobic membrane bioreactors (MBRs) and sequencing batch reactors (SBRs). At a low organic loading rate of 0.22 kg-COD m3 d1, both types of reactors were capable of treating the wastewater such that the treated effluent met the discharge regulation except for the total dissolved solids. However, when the loading rate was increased to 2.92 kg-COD m3 d1, foaming issues resulted in unstable performance. Overall, the MBRs achieved better solid removal but the SBRs performed better in regards to the degradation of aromatic compounds, as determined by UV absorbance (UVA). Finally, ozonation was applied on two different streams and showed promise on the strong stream – that corresponds to the formulation effluent and contains most of the biorefractory compounds. Ozonation successfully reduced the UVA, lowered the pH and increased the biochemical oxygen demand : chemical oxygen demand (BOD5 : COD) ratio of the strong stream. However, it was less efficient on the effluent having undergone pre-treatment by a biofilter due to a lack of selectivity
O. Lefebvre (corresponding author) X. Shi H. Y. Ng Department of Civil and Environmental Engineering, Centre for Water Research, National University of Singapore, 1 Engineering Dr. 2, Singapore 117576, Singapore E-mail: [email protected] C. H. Wu Department of Chemical and Materials Engineering, National Kaohsiung University of Applied Sciences, 415 Chien Kung Road, Kaohsiung 807, Taiwan, R.O.C.
towards refractory compounds. Key words
| biological treatment, membrane bioreactor, ozonation, pharmaceutical wastewater, sequencing batch reactor
INTRODUCTION Pharmaceutical wastewater is generally characterized by high toxicity and the presence of refractory compounds that limit its biodegradability, making it a potential threat to the natural environment and to wastewater treatment plants, if not handled properly (Gros et al. ). The manufacturing of pharmaceutical compounds typically involves a variety of stages including conversion of natural substances into pharmaceutical ingredients through fermentation and extraction processes and mostly chemical synthesis. These initial steps are then followed by formulation and packaging of the final product (Oktem et al. ). The amount and variety of wastes generated during the production of pharmaceuticals is significantly higher than the amount of the actual finished product and it has been reported that 200 to 30,000 kg of wastes can typically be generated for every kilogram of active ingredient produced (NRDC ). The composition of these pharmaceutical by-products varies as it depends on the type of drug manufactured, the materials used in the production and the actual operations involved. They can include biological doi: 10.2166/wst.2013.729
substances like fermentation wastes, excess extraction solvents remaining after the isolation and purification of active ingredients from natural sources, pharmacologically active agents like anti-coagulants and chemotherapeutic agents, as well as cleaning agents and disinfectants which are used to sterilize equipment. The pollution load of the wastewater stream also depends on the pharmaceutical production line. For example, the wastewater stream produced from the washing of equipment is characterized by smaller effluent flow and low pollution load (weak stream). On the other hand, the effluent generated by the formulation process is more heavily polluted and usually referred to as a strong stream. This is because the formulation effluent has low biodegradability due to the high level of active substance (Balcioglu & Otker ). These pharmaceutical byproducts from the various production lines of the pharmaceutical manufacturing facilities eventually become part of the overall pharmaceutical wastewater which can have chemical oxygen demand (COD) as high as 80,000 mg L1 (Nandy & Kaul ).
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Several categories of pharmaceuticals raise particular concerns and among them antibiotics have significant impact in the environment where they can disrupt wastewater treatment processes and adversely affect ecosystems (Schröder ; Sim et al. ). Furthermore, pharmaceutical wastewater resulting from the manufacturing of antibiotics may contain biorefractory materials that cannot be readily degraded (Schröder ). Yet, biological treatment can still be a viable choice for treatment in combination with physico-chemical processes (Zhou et al. ). Because of the elevated COD content of pharmaceutical wastewater, anaerobic treatment could be a suitable option; however, it is not always feasible in particular because the high total dissolved solids (TDS) content of such wastewater interferes with the activity of methanogenic bacteria (Lefebvre & Moletta ). In this study, wastewater was collected from a pharmaceutical company manufacturing antibiotics (penicillin). Two distinct wastewater streams are produced on site: a strong stream – corresponding to the formulation effluent and characterized by very high organic load – and a weak stream. The ratio of strong to weak stream is 1 : 2 (v/v). The current on-site treatment process is schematized in Figure 1. The two streams are first mixed in an equalization tank where the pH is adjusted to 7 by addition of phosphoric acid and urea is added as a source of easily bioavailable nitrogen. The mixed stream further undergoes aerobic treatment in the form of a biofilter followed by conventional activated sludge (CAS). In this study, two options were considered as alternative aerobic treatment processes: the membrane bioreactor (MBR) – because of its ability to consistently produce high quality effluent devoid of suspended solids – and the sequencing batch reactor (SBR), which is known as being a particularly robust system suitable for the treatment of
Figure 1
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industrial wastewater from food industries (Lefebvre et al. ) to tanneries (Lefebvre et al. ) and dairy industries (Torrijos et al. ). Furthermore, the efficacy of ozonation to increase the biodegradability of the pharmaceutical wastewater was assessed in batch tests. Again two options were considered: ozonation of the strong stream (which contains most of the biorefractory compounds) before mixing with the weak stream and ozonation of the biofilter effluent before introducing it into the CAS (Figure 1).
METHODS Wastewater collection and preparation The strong stream and weak stream of the wastewater were collected from the pharmaceutical company and stored in the dark at 4 C. Prior to feeding to the bioreactors, the strong and weak streams were mixed at a ratio of 1 : 2 (v/v) resulting in a mixed stream which pH was adjusted to 7 by addition of concentrated phosphoric acid and into which 0.28 g L1 of urea was added. The mixed stream was then fed to the bioreactors (MBR and SBR) described below. The biofilter effluent, which was used for ozonation experiments in batch tests, was also collected from the actual plant, stored in the dark at 4 C and used as is, without any further adjustment. W
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Bioreactor set-ups Two different biotechnologies were used in this study: the MBR and the SBR. All reactors were operated in duplicate. The aerobic MBRs (working volume ¼ 7 L) were equipped with flat-sheet polyolefin membranes with a nominal pore size of 0.45 μm, attached to a single module support and
Current treatment process implemented by the pharmaceutical company to treat its wastewater and options considered for ozonation in this study.
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immersed into each MBR. The effective membrane area of each module was 0.11 m2, and the contact angle of the clean membrane surface was found to be 72 (Ng & Ng ). A schematic of the MBR reactor design used in this study has also been made available by Ng & Ng () and is not reproduced here. The suction cycle consisted of 8-min suction followed by 2-min relaxation. In addition, the membrane modules were taken out from the reactors and the membrane surfaces were gently wiped with a sponge on a weekly basis. No chemical cleaning was needed during the entire course of the experiment as fouling was not severe. The aerobic SBRs (working volume ¼ 4 L) were operated with 24-h cycles divided as follows: 40 min feeding, 22.3 h reaction, 30 min settling and 30 min withdrawal. A schematic can be found elsewhere (Lefebvre et al. ). All reactors were operated at ambient temperature (25 ± 5 C) and hydraulic retention time (HRT) in the range of 7–8 d. Solids retention time (SRT) averaged 50 d for the MBRs and 20 d for the SBRs. W
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Ozonation Ozonation was performed using an Ozonia Triogen laboratory ozone generator (Suez Environment, France). Ozone was generated with bottled oxygen at a flow rate of 5 L min1. Ozonation was carried out in batch mode. For each batch, 150 mL of wastewater was treated in a 500 mL Drechsel bottle. The ozone demand was determined by subtracting the excess (i.e., unreacted) and residual ozone to the actual ozone dose. The ozone dose was determined by the iodometric titration semi-batch method of Standard Methods (APHA ) and was evaluated at 5.6 g h1 in our experimental conditions. The excess O3 leaving the reactor vessel was measured by trapping into a gas washing bottle containing potassium iodide and the residual O3 in the wastewater was determined by the indigo colorimetric method (APHA ). During ozonation, samples were taken at regular time intervals and analyzed. Analyses TDS, total suspended solids (TSS) and volatile suspended solids (VSS) were determined by centrifugation at 15,000 rpm for 15 min, following the AFNOR recommendations (AFNOR ). The supernatant was then analyzed for dissolved COD, biochemical oxygen demand (BOD5) and pH according to Standard Methods (APHA ). Aromatic compounds and color were further determined by absorbance at 254 nm (UV absorbance, UVA)
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and 580 nm (corresponding to yellow-orange, the natural color of the wastewater), respectively. Total organic carbon (TOC) and total nitrogen (TN) were analyzed by a TOC/TN analyzer (Shimadzu, Japan). Ammonium (NHþ 4 ), 3 nitrate (NO3 ), nitrite (NO2 ), phosphate (PO4 ) and sulfates (SO2 4 ) were determined by ion chromatography (Dionex, USA), after filtration using 0.45-μm glassfiber filters. Microscopic observation of Gram-stained foam bacteria was performed under bright field illumination using a ×100 magnification oil immersion objective.
RESULTS AND DISCUSSION Wastewater characterization Detailed characterization of the various streams used in this study is presented in Table 1, based on weekly analysis throughout the experimental period. The strong stream was characterized by extremely high COD and UVA, high N content but not in readily available form (little ammonium) and very low P concentration. In addition, the pH of the strong stream (11.3) required neutralization before biological treatment. On the other hand, the weak stream displayed lower pH, COD and UVA, but also lacked P and readily available N. Both N and P limitations were overcome in the mixed stream by addition of urea and phosphoric acid, and as a consequence the C : N : P ratio of the mixed stream averaged 100 : 20 : 5, indicating that both N and P were not only sufficient but in slight excess. Overall, the pH and BOD5 : COD ratio of the mixed stream were 7.3 and 0.8, respectively, making it a suitable candidate for biodegradation. Similarly, the biofilter effluent was characterized by high COD concentration, on average 47% of that of the mixed stream. However, its pH was more alkaline and its TSS (and VSS) content was higher, as a result of biomass washout from the biofilter. The additional biomass content in the biofilter effluent can explain the increased overall biodegradability (BOD5 : COD ratio of 0.9) of the biofilter effluent as compared to that of the mixed stream. Biological treatment performance Aerobic treatment of pharmaceutical wastewater in this study was performed on the mixed stream at two different organic loading rates (OLR): a low OLR (0.22 ± 0.03 kgCOD m3 d1) and a high OLR (2.92 ± 0.44 kg-COD m3 d1). The low OLR was obtained by introducing diluted wastewater in the reactors. The reactors were operated
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Table 1
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Characterization of pharmaceutical wastewater. The mixed stream was prepared in the laboratory by mixing the strong and weak streams at a ratio of 1 : 2 (v/v) adjusting the pH to 7 with phosphoric acid and adding urea at a concentration of 0.28 g L1
Strong stream
Weak stream
Mixed stream
Biofilter effluent
COD (mg L )
54,800 ± 964
6,877 ± 321
19,099 ± 3,614
11,985 ± 1,004
BOD5 (mg L1)
33,975 ± 6,390
5,957 ± 1,585
16,148 ± 1,578
10,919 ± 874 5,605 ± 646
1
1
TOC (mg L )
20,367 ± 84
4,147 ± 209
8,125 ± 1,472
UVA (cm1)
138.9 ± 37.5
4.3 ± 0.5
17.0 ± 2.9
1
32.6 ± 4.9
Color (cm )
0.25 ± 0.04
0.25 ± 0.12
0.24 ± 0.08
TN (mg L1)
3,923 ± 400
582 ± 39
1,967 ± 158
1 NHþ 4 (mg L )
360 ± 59
11 ± 2
N.D.
N.D.
1 NO 3 (mg L )
13 ± 1
5±1
N.D.
N.D. N.D.
1 NO 2 (mg L )
10 ± 1
b.d.l.
N.D.
pH
11.3 ± 0.4
6.3 ± 0.2
7.3 ± 0.1
1
1.17 ± 0.25 1,978 ± 102
8.9 ± 0.1
TDS (mg L )
66,793 ± 1,362
6,447 ± 640
25,846 ± 1,442
TSS (mg L1)
240 ± 113
153 ± 50
234 ± 55
VSS (mg L )
107 ± 80
80 ± 56
124 ± 69
928 ± 234
1 P-PO3 4 (mg L )
0.9 ± 0.1
3.7 ± 1.1
640 ± 236
430 ± 89
1 SO2 4 (mg L )
670± 127
55 ± 48
N.D.
1
28,772 ± 642 1,256 ± 294
N.D.
N.D. ¼ not determined; b.d.l. ¼ below detection limit.
Table 2
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Performance of bioreactors treating pharmaceutical wastewater as a function of the loading rate 0.22 kg-COD m3 d1
2.92 kg-COD m3 d1
MBR OLR
SBR
Treated effluent 1
COD (mg L ) 1
BOD5 (mg L )
184 ± 35 3±1
TOC (mg L1)
123 ± 37
TN (mg L1)
362 ± 55
1 P-PO3 4 (mg L )
168 ± 76
1
TDS (mg L )
5,460 ± 576
TSS (mg L1)
12 ± 8
VSS (mg L1)
RE (%)
89 ± 2 > 99
MBR
Treated effluent
135 ± 12
92 ± 1 > 99
Treated effluent
3,246 ± 977
RE (%)
87 ± 3
Treated effluent
3,152 ± 1,419
RE (%)
82 ± 10
313 ± 256
98 ± 2
933 ± 710
95 ± 4
61 ± 11
91 ± 1
628 ± 392
86 ± 13
644 ± 265
90 ± 6
0
159 ± 33
18 ± 16
67 ± 23
0
75 ± 67
0
82 ± 5
2±1
RE (%)
SBR
1,677 ± 1,522
0
687 ± 459
0
949 ± 353
0
906 ± 331
0
0
3,315 ± 78
0
38,792 ± 2,036
35,938 ± 5,410
0
95 ± 7
148 ± 49
0
105 ± 43
55 ± 24
2,792 ± 1,947
0
3±7
97 ± 9
115 ± 65
57 ± 33
52 ± 37
2,097 ± 2,117
UVA (cm1)
1.2 ± 0.3
46 ± 12
0.8 ± 0.1
60 ± 5
19.8 ± 3.4
8±9
11.1 ± 3.0
33 ± 17
Color (cm1)
0.02 ± 0.01
86 ± 9
0.01 ± 0.01
93 ± 7
0.08 ± 0.04
36 ± 23
0.11 ± 0.05
34 ± 34
0
0
RE ¼ removal efficiency.
under each OLR for a period of 3 months and the steadystate results are presented in Table 2. At low OLR, the MBRs showed better TSS and VSS removal; however the TOC, COD and BOD5 removal efficiencies, determined in the dissolved fraction, were slightly higher in the SBRs (on average 91, 92 and >99%, respectively) than
in the MBRs. Color removal was also excellent in both MBRs (86 ± 9%) and SBRs (93 ± 7%). The UVA removal was significantly higher in the SBRs than in the MBRs (on average 60% vs. 46%) and this improved performance can typically be explained by the SBR’s ability to retain sludge almost as efficiently as an MBR, combined to improved
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adsorption and trapping of the organic content in the sludge during the settling phase (Lefebvre et al. ). This is in accordance with the results of Ng et al. () who found that an SBR could remove up to 96% (HRT of 20 d) of the COD of a pharmaceutical wastewater having an initial COD of 26,500 mg L1 and on average 80% at an HRT of 6.67 d. More recently, a full-scale SBR was reported to remove up to 98% of the COD of the wastewater generated by a drug manufacturing plant located in Turkey, following pretreatment with Fenton’s oxidation (Tekin et al. ). Unsurprisingly, N and P removal efficiencies were negligible in both types of reactors because they were provided in excess. At low OLR, discharge criteria according to Singapore’s regulations for discharge in public sewers were met for all parameters (BOD5 < 400 mg L1; COD < 600 mg L1; TSS < 400 mg L1) except for TDS (TDS > 3,000 mg L1). After 42 days, the mixed liquor volatile suspended solids (MLVSS) became stable at 1,200 mg L1 in the MBRs and 600 mg L1 in the SBRs (Figure 2). The observed biomass yield was determined by summing up the biomass accumulated (slope of Figure 2) and the biomass wasted (obtained via the SRT of respectively 50 and 20 d for the MBRs and SBRs) in a period of time (expressed in g-VSS d1), and dividing the result by the amount of BOD consumed in the same period of time (expressed in g-BOD5 d1). Between day 42 and 91, corresponding to stable operation at low OLR, the observed biomass yield averaged 0.1 and 0.2 gVSS g-BOD1 in the MBR and SBR, respectively. These values are significantly lower than that obtained in typical biological treatment systems (∼0.4 g-VSS g-BOD1) and the possible explanations are: (i) higher than usual endogenous respiration due to the relatively low loading rate; and (ii) the possible presence of growth inhibitory compounds in the
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wastewater. This can be related to the nature of the pharmaceutical wastewater (antibiotics), residual antibiotics being obviously capable of inhibiting the growth of various microorganisms (Wollenberger et al. ). Another possible source of inhibition consists of the byproducts released during biological treatment. At higher OLR of 2.92 kg-COD m3 d1, a rapid increase of MLVSS was observed up to 14,450 mg L1 on average in the MBRs, slightly higher than the maximum of 10,000 mg L1 achieved in the SBRs (Figure 2). These values correspond to an observed growth yield of 0.3 g-VSS g-BOD1, roughly similar for both MBRs and SBRs. These values are still a bit lower than expected theoretically, indicating possible inhibition from residual antibiotics and byproducts; however, they are much higher than that obtained at lower OLR. This is a strong sign that the main reason for slow biomass growth at low OLR was the high rate of endogenous respiration. Biomass growth was accompanied by a rapid degradation of the performance of all reactors and even though the TOC, COD and BOD5 removal efficiencies were still up to 90, 87 and 98%, respectively, the discharge criteria were no longer met (Table 2). The most likely reason was the overloading of the bioreactors under high loading rate. Furthermore, all bioreactors became highly unstable and foaming was frequently observed, despite extensive use of a defoaming agent. The subsequent loss of biomass resulting from foaming events explains the drops of MLVSS observed in both MBRs and SBRs after day 120 (Figure 2). Foaming in aeration tanks is frequently associated with two bacteria genera, Nocardia and Microthrix, both of which were identified by microscopy in foam samples from both types of reactors (Figure 3). Their presence can be directly associated with high aeration rates (Tchobanoglous et al. ). Nocardia and Microthrix indeed have hydrophobic cell surfaces that attach to air bubbles and cause foam. In our experiment, high aeration rate was required due to the very high COD and BOD concentrations in order to maintain a dissolved oxygen concentration above 4 mg L1 at all time in the MBRs and at the end of the SBR cycles (data not shown). However, it is worth mentioning here that neither membrane filtration in the MBRs nor biomass settling in the SBRs were ever compromised at high OLR, showing the efficacy of both technologies for accumulating large amounts of biomass. Ozonation experiments
Figure 2
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Evolution of MLVSS according to the organic loading rate in the bioreactors showcased in this study.
Ozonation was tested as a means to improve the wastewater biodegradability. It was implemented in batch mode either
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Figure 3
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Microscopic observation of (a) Gram-positive Microthrix sp., characterized by thin and smooth curves (MBR foam sample), and (b) Gram-positive Nocardia sp., characterized by short, branched filaments (SBR foam sample).
on the strong stream or directly following the biofilter (cf. options in Figure 1). The results are shown in Figure 4, from which it appears that ozone was very efficient at discoloring the wastewater, as well as reducing the UVA and the pH of both the strong stream and the biofilter effluent. The impact on the UVA is particularly important because this absorbance is an excellent indicator of the aromatic content of the wastewater (Battimelli et al. ). This is particularly true in the presence of antibiotics (Homem & Santos ). The impact of the ozone dose on UVA was particularly obvious on the strong stream up to an ozone demand of 0.5 g g-COD1. A similar ozone dose was shown to be optimal for the pretreatment of biorefractory compounds (melanoidins) in wastewater from molasse fermentation industries (Battimelli et al. ). In that study, ozonation also led to a decrease in COD concentration; yet in our study, the COD was unaffected (Figure 4(a)). However, ozone proved beneficial at increasing the BOD5 of the strong stream (Figure 4(b)), which resulted in an improved BOD5 : COD ratio from 0.64 to 0.89 with an ozone dose
Figure 4
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of 0.5 g g-COD1. However, when used on the biofilter effluent, the results were not encouraging (Figure 4(c)). Therefore, it can be concluded that ozonation was better applied on the strong stream, where it was more selective towards biorefractory compounds, as reflected by the UVA. On the contrary, ozone was inefficient on the biofilter effluent due to the lower initial UVA content.
CONCLUSIONS In this study, different approaches were applied to the treatment of pharmaceutical wastewater from the antibiotics industry. Biological treatment relied on the MBR and SBR technology. The MBR achieved better suspended solid removal but the SBR competitiveness was demonstrated by its improved dissolved UVA removal, especially a low OLR. However, in both cases, high loading rate was responsible for heavy foaming and caused a degradation of the performance. In addition, the presence of aromatic
Effect of ozone demand on (a) COD, (b) BOD5, (c) BOD5 : COD ratio, (d) pH, (e) UVA, and (f) color of the strong stream (▪) and biofilter effluent (○) generated by the pharmaceutical company.
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compounds might prove detrimental to biomass growth. For this purpose, the efficacy of ozonation pre-treatment was assessed in a batch study and showed promise when targeted at the strong stream. Future work shall make use of ozonation in continuous mode to improve the performance of biological degradation of the pharmaceutical wastewater at high loading rate. TDS removal will also be considered by means of physico-chemical methods. Finally, N and P dosing should also be adjusted so as to avoid wastage of chemicals, even though they are not regulated for discharge into the sewer as per Singapore’s regulations.
REFERENCES AFNOR Qualité de l’eau, recueil des normes françaises. Agence Française de Normalisation, Paris. APHA Standard Methods for the Examination of Water and Wastewater. American Public Health Association, Washington, DC. Balcioglu, I. A. & Otker, M. Treatment of pharmaceutical wastewater containing antibiotics by O3 and O3/H2O2 processes. Chemosphere 50 (1), 85–95. Battimelli, A., Loisel, D., Garcia-Bernet, D., Carrere, H. & Delgenes, J. P. Combined ozone pretreatment and biological processes for removal of colored and biorefractory compounds in wastewater from molasses fermentation industries. J. Chem. Technol. Biotechnol. 85 (7), 968–975. Gros, M., Petrovic, M., Ginebreda, A. & Barceló, D. Removal of pharmaceuticals during wastewater treatment and environmental risk assessment using hazard indexes. Environ. Int. 36 (1), 15–26. Homem, V. & Santos, L. Degradation and removal methods of antibiotics from aqueous matrices – a review. J. Environ. Manage. 92 (10), 2304–2347. Lefebvre, O. & Moletta, R. Treatment of organic pollution in industrial saline wastewater: a literature review. Water Res. 40 (20), 3671–3682. Lefebvre, O., Habouzit, F., Bru, V., Delgenes, J. P., Godon, J. J. & Moletta, R. Treatment of hypersaline industrial wastewater by a microbial consortium in a sequencing batch reactor. Environ. Technol. 25 (5), 543–553. Lefebvre, O., Vasudevan, N., Torrijos, M., Thanasekaran, K. & Moletta, R. Halophilic biological treatment of tannery
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soak liquor in a sequencing batch reactor. Water Res. 39 (8), 1471–1480. Nandy, T. & Kaul, S. N. Anaerobic pre-treatment of herbalbased pharmaceutical wastewater using fixed-film reactor with recourse to energy recovery. Water Res. 35 (2), 351–362. Ng, T. C. A. & Ng, H. Y. Characterisation of initial fouling in aerobic submerged membrane bioreactors in relation to physico-chemical characteristics under different flux conditions. Water Res. 44 (7), 2336–2348. Ng, W. J., Yap, M. G. S. & Sivadas, M. Biological treatment of a pharmaceutical wastewater. Biol. Waste 29 (4), 299–312. NRDC Dosed without prescription: preventing pharmaceutical contamination of our nation’s drinking water. Report of the Natural Resources Defense Council, New York. Oktem, Y. A., Ince, O., Sallis, P., Donnelly, T. & Ince, B. K. Anaerobic treatment of a chemical synthesis-based pharmaceutical wastewater in a hybrid upflow anaerobic sludge blanket reactor. Bioresour. Technol. 99 (5), 1089–1096. Schröder, H. F. Substance-specific detection and pursuit of non-eliminable compounds during biological treatment of waste water from the pharmaceutical industry. Waste Manage. 19 (2), 111–123. Sim, W.-J., Lee, J.-W. & Oh, J.-E. Occurrence and fate of pharmaceuticals in wastewater treatment plants and rivers in Korea. Environ. Pollut. 158 (5), 1938–1947. Tchobanoglous, G., Burton, F. L. & Stensel, H. D. Wastewater Engineering: Treatment and Reuse. McGrawHill, New York. Tekin, H., Bilkay, O., Ataberk, S. S., Balta, T. H., Ceribasi, I. H., Sanin, F. D., Dilek, F. B. & Yetis, U. Use of Fenton oxidation to improve the biodegradability of a pharmaceutical wastewater. J. Hazard. Mater. 136 (2), 258–265. Torrijos, M., Sousbie, P., Moletta, R. & Delgenes, J. P. High COD wastewater treatment in an aerobic SBR: treatment of effluent from a small farm goat’s cheese dairy. Water Sci. Technol. 50 (10), 259–267. Wollenberger, L., Halling-Sørensen, B. & Kusk, K. O. Acute and chronic toxicity of veterinary antibiotics to Daphnia magna. Chemosphere 40 (7), 723–730. Zhou, P., Su, C. Y., Li, B. W. & Yi, Q. Treatment of highstrength pharmaceutical wastewater and removal of antibiotics in anaerobic and aerobic biological treatment processes. J Environ. Eng. ASCE 132 (1), 129–136.
First received 27 June 2013; accepted in revised form 23 October 2013. Available online 14 December 2013
© IWA Publishing 2014 Water Science & Technology
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Biological treatment of pharmaceutical wastewater from the antibiotics industry O. Lefebvre, X. Shi, C. H. Wu and H. Y. Ng
ABSTRACT Pharmaceutical wastewater generated by an antibiotics (penicillin) company was treated by aerobic membrane bioreactors (MBRs) and sequencing batch reactors (SBRs). At a low organic loading rate of 0.22 kg-COD m3 d1, both types of reactors were capable of treating the wastewater such that the treated effluent met the discharge regulation except for the total dissolved solids. However, when the loading rate was increased to 2.92 kg-COD m3 d1, foaming issues resulted in unstable performance. Overall, the MBRs achieved better solid removal but the SBRs performed better in regards to the degradation of aromatic compounds, as determined by UV absorbance (UVA). Finally, ozonation was applied on two different streams and showed promise on the strong stream – that corresponds to the formulation effluent and contains most of the biorefractory compounds. Ozonation successfully reduced the UVA, lowered the pH and increased the biochemical oxygen demand : chemical oxygen demand (BOD5 : COD) ratio of the strong stream. However, it was less efficient on the effluent having undergone pre-treatment by a biofilter due to a lack of selectivity
O. Lefebvre (corresponding author) X. Shi H. Y. Ng Department of Civil and Environmental Engineering, Centre for Water Research, National University of Singapore, 1 Engineering Dr. 2, Singapore 117576, Singapore E-mail: [email protected] C. H. Wu Department of Chemical and Materials Engineering, National Kaohsiung University of Applied Sciences, 415 Chien Kung Road, Kaohsiung 807, Taiwan, R.O.C.
towards refractory compounds. Key words
| biological treatment, membrane bioreactor, ozonation, pharmaceutical wastewater, sequencing batch reactor
INTRODUCTION Pharmaceutical wastewater is generally characterized by high toxicity and the presence of refractory compounds that limit its biodegradability, making it a potential threat to the natural environment and to wastewater treatment plants, if not handled properly (Gros et al. ). The manufacturing of pharmaceutical compounds typically involves a variety of stages including conversion of natural substances into pharmaceutical ingredients through fermentation and extraction processes and mostly chemical synthesis. These initial steps are then followed by formulation and packaging of the final product (Oktem et al. ). The amount and variety of wastes generated during the production of pharmaceuticals is significantly higher than the amount of the actual finished product and it has been reported that 200 to 30,000 kg of wastes can typically be generated for every kilogram of active ingredient produced (NRDC ). The composition of these pharmaceutical by-products varies as it depends on the type of drug manufactured, the materials used in the production and the actual operations involved. They can include biological doi: 10.2166/wst.2013.729
substances like fermentation wastes, excess extraction solvents remaining after the isolation and purification of active ingredients from natural sources, pharmacologically active agents like anti-coagulants and chemotherapeutic agents, as well as cleaning agents and disinfectants which are used to sterilize equipment. The pollution load of the wastewater stream also depends on the pharmaceutical production line. For example, the wastewater stream produced from the washing of equipment is characterized by smaller effluent flow and low pollution load (weak stream). On the other hand, the effluent generated by the formulation process is more heavily polluted and usually referred to as a strong stream. This is because the formulation effluent has low biodegradability due to the high level of active substance (Balcioglu & Otker ). These pharmaceutical byproducts from the various production lines of the pharmaceutical manufacturing facilities eventually become part of the overall pharmaceutical wastewater which can have chemical oxygen demand (COD) as high as 80,000 mg L1 (Nandy & Kaul ).
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Biological treatment of pharmaceutical wastewater from the antibiotics industry
Several categories of pharmaceuticals raise particular concerns and among them antibiotics have significant impact in the environment where they can disrupt wastewater treatment processes and adversely affect ecosystems (Schröder ; Sim et al. ). Furthermore, pharmaceutical wastewater resulting from the manufacturing of antibiotics may contain biorefractory materials that cannot be readily degraded (Schröder ). Yet, biological treatment can still be a viable choice for treatment in combination with physico-chemical processes (Zhou et al. ). Because of the elevated COD content of pharmaceutical wastewater, anaerobic treatment could be a suitable option; however, it is not always feasible in particular because the high total dissolved solids (TDS) content of such wastewater interferes with the activity of methanogenic bacteria (Lefebvre & Moletta ). In this study, wastewater was collected from a pharmaceutical company manufacturing antibiotics (penicillin). Two distinct wastewater streams are produced on site: a strong stream – corresponding to the formulation effluent and characterized by very high organic load – and a weak stream. The ratio of strong to weak stream is 1 : 2 (v/v). The current on-site treatment process is schematized in Figure 1. The two streams are first mixed in an equalization tank where the pH is adjusted to 7 by addition of phosphoric acid and urea is added as a source of easily bioavailable nitrogen. The mixed stream further undergoes aerobic treatment in the form of a biofilter followed by conventional activated sludge (CAS). In this study, two options were considered as alternative aerobic treatment processes: the membrane bioreactor (MBR) – because of its ability to consistently produce high quality effluent devoid of suspended solids – and the sequencing batch reactor (SBR), which is known as being a particularly robust system suitable for the treatment of
Figure 1
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industrial wastewater from food industries (Lefebvre et al. ) to tanneries (Lefebvre et al. ) and dairy industries (Torrijos et al. ). Furthermore, the efficacy of ozonation to increase the biodegradability of the pharmaceutical wastewater was assessed in batch tests. Again two options were considered: ozonation of the strong stream (which contains most of the biorefractory compounds) before mixing with the weak stream and ozonation of the biofilter effluent before introducing it into the CAS (Figure 1).
METHODS Wastewater collection and preparation The strong stream and weak stream of the wastewater were collected from the pharmaceutical company and stored in the dark at 4 C. Prior to feeding to the bioreactors, the strong and weak streams were mixed at a ratio of 1 : 2 (v/v) resulting in a mixed stream which pH was adjusted to 7 by addition of concentrated phosphoric acid and into which 0.28 g L1 of urea was added. The mixed stream was then fed to the bioreactors (MBR and SBR) described below. The biofilter effluent, which was used for ozonation experiments in batch tests, was also collected from the actual plant, stored in the dark at 4 C and used as is, without any further adjustment. W
W
Bioreactor set-ups Two different biotechnologies were used in this study: the MBR and the SBR. All reactors were operated in duplicate. The aerobic MBRs (working volume ¼ 7 L) were equipped with flat-sheet polyolefin membranes with a nominal pore size of 0.45 μm, attached to a single module support and
Current treatment process implemented by the pharmaceutical company to treat its wastewater and options considered for ozonation in this study.
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immersed into each MBR. The effective membrane area of each module was 0.11 m2, and the contact angle of the clean membrane surface was found to be 72 (Ng & Ng ). A schematic of the MBR reactor design used in this study has also been made available by Ng & Ng () and is not reproduced here. The suction cycle consisted of 8-min suction followed by 2-min relaxation. In addition, the membrane modules were taken out from the reactors and the membrane surfaces were gently wiped with a sponge on a weekly basis. No chemical cleaning was needed during the entire course of the experiment as fouling was not severe. The aerobic SBRs (working volume ¼ 4 L) were operated with 24-h cycles divided as follows: 40 min feeding, 22.3 h reaction, 30 min settling and 30 min withdrawal. A schematic can be found elsewhere (Lefebvre et al. ). All reactors were operated at ambient temperature (25 ± 5 C) and hydraulic retention time (HRT) in the range of 7–8 d. Solids retention time (SRT) averaged 50 d for the MBRs and 20 d for the SBRs. W
W
Ozonation Ozonation was performed using an Ozonia Triogen laboratory ozone generator (Suez Environment, France). Ozone was generated with bottled oxygen at a flow rate of 5 L min1. Ozonation was carried out in batch mode. For each batch, 150 mL of wastewater was treated in a 500 mL Drechsel bottle. The ozone demand was determined by subtracting the excess (i.e., unreacted) and residual ozone to the actual ozone dose. The ozone dose was determined by the iodometric titration semi-batch method of Standard Methods (APHA ) and was evaluated at 5.6 g h1 in our experimental conditions. The excess O3 leaving the reactor vessel was measured by trapping into a gas washing bottle containing potassium iodide and the residual O3 in the wastewater was determined by the indigo colorimetric method (APHA ). During ozonation, samples were taken at regular time intervals and analyzed. Analyses TDS, total suspended solids (TSS) and volatile suspended solids (VSS) were determined by centrifugation at 15,000 rpm for 15 min, following the AFNOR recommendations (AFNOR ). The supernatant was then analyzed for dissolved COD, biochemical oxygen demand (BOD5) and pH according to Standard Methods (APHA ). Aromatic compounds and color were further determined by absorbance at 254 nm (UV absorbance, UVA)
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and 580 nm (corresponding to yellow-orange, the natural color of the wastewater), respectively. Total organic carbon (TOC) and total nitrogen (TN) were analyzed by a TOC/TN analyzer (Shimadzu, Japan). Ammonium (NHþ 4 ), 3 nitrate (NO3 ), nitrite (NO2 ), phosphate (PO4 ) and sulfates (SO2 4 ) were determined by ion chromatography (Dionex, USA), after filtration using 0.45-μm glassfiber filters. Microscopic observation of Gram-stained foam bacteria was performed under bright field illumination using a ×100 magnification oil immersion objective.
RESULTS AND DISCUSSION Wastewater characterization Detailed characterization of the various streams used in this study is presented in Table 1, based on weekly analysis throughout the experimental period. The strong stream was characterized by extremely high COD and UVA, high N content but not in readily available form (little ammonium) and very low P concentration. In addition, the pH of the strong stream (11.3) required neutralization before biological treatment. On the other hand, the weak stream displayed lower pH, COD and UVA, but also lacked P and readily available N. Both N and P limitations were overcome in the mixed stream by addition of urea and phosphoric acid, and as a consequence the C : N : P ratio of the mixed stream averaged 100 : 20 : 5, indicating that both N and P were not only sufficient but in slight excess. Overall, the pH and BOD5 : COD ratio of the mixed stream were 7.3 and 0.8, respectively, making it a suitable candidate for biodegradation. Similarly, the biofilter effluent was characterized by high COD concentration, on average 47% of that of the mixed stream. However, its pH was more alkaline and its TSS (and VSS) content was higher, as a result of biomass washout from the biofilter. The additional biomass content in the biofilter effluent can explain the increased overall biodegradability (BOD5 : COD ratio of 0.9) of the biofilter effluent as compared to that of the mixed stream. Biological treatment performance Aerobic treatment of pharmaceutical wastewater in this study was performed on the mixed stream at two different organic loading rates (OLR): a low OLR (0.22 ± 0.03 kgCOD m3 d1) and a high OLR (2.92 ± 0.44 kg-COD m3 d1). The low OLR was obtained by introducing diluted wastewater in the reactors. The reactors were operated
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Table 1
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Characterization of pharmaceutical wastewater. The mixed stream was prepared in the laboratory by mixing the strong and weak streams at a ratio of 1 : 2 (v/v) adjusting the pH to 7 with phosphoric acid and adding urea at a concentration of 0.28 g L1
Strong stream
Weak stream
Mixed stream
Biofilter effluent
COD (mg L )
54,800 ± 964
6,877 ± 321
19,099 ± 3,614
11,985 ± 1,004
BOD5 (mg L1)
33,975 ± 6,390
5,957 ± 1,585
16,148 ± 1,578
10,919 ± 874 5,605 ± 646
1
1
TOC (mg L )
20,367 ± 84
4,147 ± 209
8,125 ± 1,472
UVA (cm1)
138.9 ± 37.5
4.3 ± 0.5
17.0 ± 2.9
1
32.6 ± 4.9
Color (cm )
0.25 ± 0.04
0.25 ± 0.12
0.24 ± 0.08
TN (mg L1)
3,923 ± 400
582 ± 39
1,967 ± 158
1 NHþ 4 (mg L )
360 ± 59
11 ± 2
N.D.
N.D.
1 NO 3 (mg L )
13 ± 1
5±1
N.D.
N.D. N.D.
1 NO 2 (mg L )
10 ± 1
b.d.l.
N.D.
pH
11.3 ± 0.4
6.3 ± 0.2
7.3 ± 0.1
1
1.17 ± 0.25 1,978 ± 102
8.9 ± 0.1
TDS (mg L )
66,793 ± 1,362
6,447 ± 640
25,846 ± 1,442
TSS (mg L1)
240 ± 113
153 ± 50
234 ± 55
VSS (mg L )
107 ± 80
80 ± 56
124 ± 69
928 ± 234
1 P-PO3 4 (mg L )
0.9 ± 0.1
3.7 ± 1.1
640 ± 236
430 ± 89
1 SO2 4 (mg L )
670± 127
55 ± 48
N.D.
1
28,772 ± 642 1,256 ± 294
N.D.
N.D. ¼ not determined; b.d.l. ¼ below detection limit.
Table 2
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Performance of bioreactors treating pharmaceutical wastewater as a function of the loading rate 0.22 kg-COD m3 d1
2.92 kg-COD m3 d1
MBR OLR
SBR
Treated effluent 1
COD (mg L ) 1
BOD5 (mg L )
184 ± 35 3±1
TOC (mg L1)
123 ± 37
TN (mg L1)
362 ± 55
1 P-PO3 4 (mg L )
168 ± 76
1
TDS (mg L )
5,460 ± 576
TSS (mg L1)
12 ± 8
VSS (mg L1)
RE (%)
89 ± 2 > 99
MBR
Treated effluent
135 ± 12
92 ± 1 > 99
Treated effluent
3,246 ± 977
RE (%)
87 ± 3
Treated effluent
3,152 ± 1,419
RE (%)
82 ± 10
313 ± 256
98 ± 2
933 ± 710
95 ± 4
61 ± 11
91 ± 1
628 ± 392
86 ± 13
644 ± 265
90 ± 6
0
159 ± 33
18 ± 16
67 ± 23
0
75 ± 67
0
82 ± 5
2±1
RE (%)
SBR
1,677 ± 1,522
0
687 ± 459
0
949 ± 353
0
906 ± 331
0
0
3,315 ± 78
0
38,792 ± 2,036
35,938 ± 5,410
0
95 ± 7
148 ± 49
0
105 ± 43
55 ± 24
2,792 ± 1,947
0
3±7
97 ± 9
115 ± 65
57 ± 33
52 ± 37
2,097 ± 2,117
UVA (cm1)
1.2 ± 0.3
46 ± 12
0.8 ± 0.1
60 ± 5
19.8 ± 3.4
8±9
11.1 ± 3.0
33 ± 17
Color (cm1)
0.02 ± 0.01
86 ± 9
0.01 ± 0.01
93 ± 7
0.08 ± 0.04
36 ± 23
0.11 ± 0.05
34 ± 34
0
0
RE ¼ removal efficiency.
under each OLR for a period of 3 months and the steadystate results are presented in Table 2. At low OLR, the MBRs showed better TSS and VSS removal; however the TOC, COD and BOD5 removal efficiencies, determined in the dissolved fraction, were slightly higher in the SBRs (on average 91, 92 and >99%, respectively) than
in the MBRs. Color removal was also excellent in both MBRs (86 ± 9%) and SBRs (93 ± 7%). The UVA removal was significantly higher in the SBRs than in the MBRs (on average 60% vs. 46%) and this improved performance can typically be explained by the SBR’s ability to retain sludge almost as efficiently as an MBR, combined to improved
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adsorption and trapping of the organic content in the sludge during the settling phase (Lefebvre et al. ). This is in accordance with the results of Ng et al. () who found that an SBR could remove up to 96% (HRT of 20 d) of the COD of a pharmaceutical wastewater having an initial COD of 26,500 mg L1 and on average 80% at an HRT of 6.67 d. More recently, a full-scale SBR was reported to remove up to 98% of the COD of the wastewater generated by a drug manufacturing plant located in Turkey, following pretreatment with Fenton’s oxidation (Tekin et al. ). Unsurprisingly, N and P removal efficiencies were negligible in both types of reactors because they were provided in excess. At low OLR, discharge criteria according to Singapore’s regulations for discharge in public sewers were met for all parameters (BOD5 < 400 mg L1; COD < 600 mg L1; TSS < 400 mg L1) except for TDS (TDS > 3,000 mg L1). After 42 days, the mixed liquor volatile suspended solids (MLVSS) became stable at 1,200 mg L1 in the MBRs and 600 mg L1 in the SBRs (Figure 2). The observed biomass yield was determined by summing up the biomass accumulated (slope of Figure 2) and the biomass wasted (obtained via the SRT of respectively 50 and 20 d for the MBRs and SBRs) in a period of time (expressed in g-VSS d1), and dividing the result by the amount of BOD consumed in the same period of time (expressed in g-BOD5 d1). Between day 42 and 91, corresponding to stable operation at low OLR, the observed biomass yield averaged 0.1 and 0.2 gVSS g-BOD1 in the MBR and SBR, respectively. These values are significantly lower than that obtained in typical biological treatment systems (∼0.4 g-VSS g-BOD1) and the possible explanations are: (i) higher than usual endogenous respiration due to the relatively low loading rate; and (ii) the possible presence of growth inhibitory compounds in the
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wastewater. This can be related to the nature of the pharmaceutical wastewater (antibiotics), residual antibiotics being obviously capable of inhibiting the growth of various microorganisms (Wollenberger et al. ). Another possible source of inhibition consists of the byproducts released during biological treatment. At higher OLR of 2.92 kg-COD m3 d1, a rapid increase of MLVSS was observed up to 14,450 mg L1 on average in the MBRs, slightly higher than the maximum of 10,000 mg L1 achieved in the SBRs (Figure 2). These values correspond to an observed growth yield of 0.3 g-VSS g-BOD1, roughly similar for both MBRs and SBRs. These values are still a bit lower than expected theoretically, indicating possible inhibition from residual antibiotics and byproducts; however, they are much higher than that obtained at lower OLR. This is a strong sign that the main reason for slow biomass growth at low OLR was the high rate of endogenous respiration. Biomass growth was accompanied by a rapid degradation of the performance of all reactors and even though the TOC, COD and BOD5 removal efficiencies were still up to 90, 87 and 98%, respectively, the discharge criteria were no longer met (Table 2). The most likely reason was the overloading of the bioreactors under high loading rate. Furthermore, all bioreactors became highly unstable and foaming was frequently observed, despite extensive use of a defoaming agent. The subsequent loss of biomass resulting from foaming events explains the drops of MLVSS observed in both MBRs and SBRs after day 120 (Figure 2). Foaming in aeration tanks is frequently associated with two bacteria genera, Nocardia and Microthrix, both of which were identified by microscopy in foam samples from both types of reactors (Figure 3). Their presence can be directly associated with high aeration rates (Tchobanoglous et al. ). Nocardia and Microthrix indeed have hydrophobic cell surfaces that attach to air bubbles and cause foam. In our experiment, high aeration rate was required due to the very high COD and BOD concentrations in order to maintain a dissolved oxygen concentration above 4 mg L1 at all time in the MBRs and at the end of the SBR cycles (data not shown). However, it is worth mentioning here that neither membrane filtration in the MBRs nor biomass settling in the SBRs were ever compromised at high OLR, showing the efficacy of both technologies for accumulating large amounts of biomass. Ozonation experiments
Figure 2
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Evolution of MLVSS according to the organic loading rate in the bioreactors showcased in this study.
Ozonation was tested as a means to improve the wastewater biodegradability. It was implemented in batch mode either
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Figure 3
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Biological treatment of pharmaceutical wastewater from the antibiotics industry
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Microscopic observation of (a) Gram-positive Microthrix sp., characterized by thin and smooth curves (MBR foam sample), and (b) Gram-positive Nocardia sp., characterized by short, branched filaments (SBR foam sample).
on the strong stream or directly following the biofilter (cf. options in Figure 1). The results are shown in Figure 4, from which it appears that ozone was very efficient at discoloring the wastewater, as well as reducing the UVA and the pH of both the strong stream and the biofilter effluent. The impact on the UVA is particularly important because this absorbance is an excellent indicator of the aromatic content of the wastewater (Battimelli et al. ). This is particularly true in the presence of antibiotics (Homem & Santos ). The impact of the ozone dose on UVA was particularly obvious on the strong stream up to an ozone demand of 0.5 g g-COD1. A similar ozone dose was shown to be optimal for the pretreatment of biorefractory compounds (melanoidins) in wastewater from molasse fermentation industries (Battimelli et al. ). In that study, ozonation also led to a decrease in COD concentration; yet in our study, the COD was unaffected (Figure 4(a)). However, ozone proved beneficial at increasing the BOD5 of the strong stream (Figure 4(b)), which resulted in an improved BOD5 : COD ratio from 0.64 to 0.89 with an ozone dose
Figure 4
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of 0.5 g g-COD1. However, when used on the biofilter effluent, the results were not encouraging (Figure 4(c)). Therefore, it can be concluded that ozonation was better applied on the strong stream, where it was more selective towards biorefractory compounds, as reflected by the UVA. On the contrary, ozone was inefficient on the biofilter effluent due to the lower initial UVA content.
CONCLUSIONS In this study, different approaches were applied to the treatment of pharmaceutical wastewater from the antibiotics industry. Biological treatment relied on the MBR and SBR technology. The MBR achieved better suspended solid removal but the SBR competitiveness was demonstrated by its improved dissolved UVA removal, especially a low OLR. However, in both cases, high loading rate was responsible for heavy foaming and caused a degradation of the performance. In addition, the presence of aromatic
Effect of ozone demand on (a) COD, (b) BOD5, (c) BOD5 : COD ratio, (d) pH, (e) UVA, and (f) color of the strong stream (▪) and biofilter effluent (○) generated by the pharmaceutical company.
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compounds might prove detrimental to biomass growth. For this purpose, the efficacy of ozonation pre-treatment was assessed in a batch study and showed promise when targeted at the strong stream. Future work shall make use of ozonation in continuous mode to improve the performance of biological degradation of the pharmaceutical wastewater at high loading rate. TDS removal will also be considered by means of physico-chemical methods. Finally, N and P dosing should also be adjusted so as to avoid wastage of chemicals, even though they are not regulated for discharge into the sewer as per Singapore’s regulations.
REFERENCES AFNOR Qualité de l’eau, recueil des normes françaises. Agence Française de Normalisation, Paris. APHA Standard Methods for the Examination of Water and Wastewater. American Public Health Association, Washington, DC. Balcioglu, I. A. & Otker, M. Treatment of pharmaceutical wastewater containing antibiotics by O3 and O3/H2O2 processes. Chemosphere 50 (1), 85–95. Battimelli, A., Loisel, D., Garcia-Bernet, D., Carrere, H. & Delgenes, J. P. Combined ozone pretreatment and biological processes for removal of colored and biorefractory compounds in wastewater from molasses fermentation industries. J. Chem. Technol. Biotechnol. 85 (7), 968–975. Gros, M., Petrovic, M., Ginebreda, A. & Barceló, D. Removal of pharmaceuticals during wastewater treatment and environmental risk assessment using hazard indexes. Environ. Int. 36 (1), 15–26. Homem, V. & Santos, L. Degradation and removal methods of antibiotics from aqueous matrices – a review. J. Environ. Manage. 92 (10), 2304–2347. Lefebvre, O. & Moletta, R. Treatment of organic pollution in industrial saline wastewater: a literature review. Water Res. 40 (20), 3671–3682. Lefebvre, O., Habouzit, F., Bru, V., Delgenes, J. P., Godon, J. J. & Moletta, R. Treatment of hypersaline industrial wastewater by a microbial consortium in a sequencing batch reactor. Environ. Technol. 25 (5), 543–553. Lefebvre, O., Vasudevan, N., Torrijos, M., Thanasekaran, K. & Moletta, R. Halophilic biological treatment of tannery
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soak liquor in a sequencing batch reactor. Water Res. 39 (8), 1471–1480. Nandy, T. & Kaul, S. N. Anaerobic pre-treatment of herbalbased pharmaceutical wastewater using fixed-film reactor with recourse to energy recovery. Water Res. 35 (2), 351–362. Ng, T. C. A. & Ng, H. Y. Characterisation of initial fouling in aerobic submerged membrane bioreactors in relation to physico-chemical characteristics under different flux conditions. Water Res. 44 (7), 2336–2348. Ng, W. J., Yap, M. G. S. & Sivadas, M. Biological treatment of a pharmaceutical wastewater. Biol. Waste 29 (4), 299–312. NRDC Dosed without prescription: preventing pharmaceutical contamination of our nation’s drinking water. Report of the Natural Resources Defense Council, New York. Oktem, Y. A., Ince, O., Sallis, P., Donnelly, T. & Ince, B. K. Anaerobic treatment of a chemical synthesis-based pharmaceutical wastewater in a hybrid upflow anaerobic sludge blanket reactor. Bioresour. Technol. 99 (5), 1089–1096. Schröder, H. F. Substance-specific detection and pursuit of non-eliminable compounds during biological treatment of waste water from the pharmaceutical industry. Waste Manage. 19 (2), 111–123. Sim, W.-J., Lee, J.-W. & Oh, J.-E. Occurrence and fate of pharmaceuticals in wastewater treatment plants and rivers in Korea. Environ. Pollut. 158 (5), 1938–1947. Tchobanoglous, G., Burton, F. L. & Stensel, H. D. Wastewater Engineering: Treatment and Reuse. McGrawHill, New York. Tekin, H., Bilkay, O., Ataberk, S. S., Balta, T. H., Ceribasi, I. H., Sanin, F. D., Dilek, F. B. & Yetis, U. Use of Fenton oxidation to improve the biodegradability of a pharmaceutical wastewater. J. Hazard. Mater. 136 (2), 258–265. Torrijos, M., Sousbie, P., Moletta, R. & Delgenes, J. P. High COD wastewater treatment in an aerobic SBR: treatment of effluent from a small farm goat’s cheese dairy. Water Sci. Technol. 50 (10), 259–267. Wollenberger, L., Halling-Sørensen, B. & Kusk, K. O. Acute and chronic toxicity of veterinary antibiotics to Daphnia magna. Chemosphere 40 (7), 723–730. Zhou, P., Su, C. Y., Li, B. W. & Yi, Q. Treatment of highstrength pharmaceutical wastewater and removal of antibiotics in anaerobic and aerobic biological treatment processes. J Environ. Eng. ASCE 132 (1), 129–136.
First received 27 June 2013; accepted in revised form 23 October 2013. Available online 14 December 2013
