Reduction of toxicity of antimicrobial compounds by degradation processes using activated sludge, gamma radiation, and UV

Chemosphere 93 (2013) 2480–2487

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Reduction of toxicity of antimicrobial compounds by degradation processes using activated sludge, gamma radiation, and UV Hyun Young Kim a,c, Junho Jeon b, Seungho Yu c, Myunjoo Lee c, Tae-Hun Kim a,c, Sang D. Kim a,⇑ a

School of Environmental Science and Engineering, Gwangju Institute of Science and Technology (GIST), 261 Cheomdan-gwagiro, Buk-gu, Gwangju 500-712, Republic of Korea Eawag, Swiss Federal Institute of Aquatic Science and Technology, 8600 Dübendorf, Switzerland c Advanced Radiation Technology Institute, Korea Atomic Energy Research Institute, Jeongeup, Jeonbuk 580-185, Republic of Korea b

h i g h l i g h t s  Degradation of antimicrobials was performed by activated sludge,

c-radiation and UV.

 Toxicity of TCN, LMC and SMZ to green algae was evaluated before and after treatment.  The algal toxicity after the degradation processes was significantly decreased. 

c-irradiation showed the highest removal efficiencies for all of the target compounds.

 Toxicity result showed the existence of transformed byproducts in the treated sample.

a r t i c l e

i n f o

Article history: Received 4 March 2013 Received in revised form 12 August 2013 Accepted 22 August 2013 Available online 29 September 2013 Keywords: Antimicrobials c-irradiation UV Activated sludge Pseudokirchneriella subcapitata Toxicity change

a b s t r a c t The occurrence and persistence of pharmacologically active compounds in the environment has been an increasingly important issue. The objectives of this study were to investigate the decomposition of aqueous antimicrobial compounds using activated sludge, c-irradiation, and UV treatment, and to evaluate the toxicity towards green algae, Pseudokirchneriella subcapitata, before and after treatment. Tetracycline (TCN), lincomycin (LMC) and sulfamethazine (SMZ) were used as target compounds. Gamma (c)-irradiation showed the highest removal efficiency for all target compounds, while UV and activated sludge treatment showed compound-dependent removal efficiencies. TCN and SMZ were well degraded by all three treatment methods. However, LMC showed extremely low removal efficiency for UV and activated sludge treatments. Overall, the algal toxicity after degradation processes was significantly decreased, and was closely correlated to removal efficiency. However, in the case of c-irradiated TCN, UV and activated sludge treated LMC as well as sludge treated SMZ, the observed toxicity was higher than expected, which indicates the substantial generation of byproducts or transformed compounds of a greater toxicity in the treated sample. Consequently, c-radiation treatment could be an effective method for removal of recalcitrant compounds such as antibiotics. Ó 2013 Elsevier Ltd. All rights reserved.

1. Introduction The occurrence and persistence of pharmacologically active compounds in the environment is an increasingly important issue. Among the associated antimicrobials are natural or semi-synthetic compounds, which inhibit microbial growth, such as bacteria, fungi and protozoa. For the last five decades, antimicrobials have been extensively used to prevent or treat a variety of infections in humans and animals, and consequently, are ubiquitous contaminants in environments (Sarmah et al., 2006). Antimicrobials are excreted in parent and metabolized form and released into water environment through wastewater treatment plant. The antimicrobials for

⇑ Corresponding author. Tel.: +82 62 970 2445; fax: +82 62 970 2434. E-mail address: [email protected] (S.D. Kim). 0045-6535/$ - see front matter Ó 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.chemosphere.2013.08.091

animal use can also be excreted into terrestrial and water environment through manure and contaminated soil (Heberer, 2002). Tetracycline (TCN), lincomycin (LMC) and sulfamethazine (SMZ) are representative antimicrobials which are widely detected in natural aquatic environments: Tetracycline has been determined as an influent (48 lg L1) as well as an effluent (at 3.6 lg L1) of wastewater treatment plants (Karthikeyan and Meyer, 2006) and in surface water (0.11 lg L1) (Kolpin et al., 2002) in USA, and has also been determined in ground water adjacent to a swine farm in Germany (at 25 lg L1) (Hamscher et al., 2002). Lincomycin and sulfamethazine have been detected in surface water at concentrations of up to 0.73 and 0.22 lg L1, respectively (Kolpin et al., 2002). These compounds were regarded as high priority antimicrobials due to their potential to enter the environment, their usage, and the hazards which they pose to terrestrial and aquatic organisms (Boxall et al., 2003).

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One of the environmental concerns surrounding such antimicrobials is the relatively low removal efficiency of removal and degradation which is observed through application to conventional treatment methods such as physical separation, chemical oxidation and adsorption, biological process, and advanced oxidation processes. For example, only 10–20% of antimicrobials are degraded through direct and indirect photolysis and hydrolysis (Andreozzi et al., 2004). Anaerobic–aerobic combined biological treatment processes have shown less than 10% removal efficiencies for ampicillin and aureomycin (Zhou et al., 2006). Antimicrobials released into watersheds via wastewater treatment plants or other sources are of potential risk to aquatic organisms and human health. Although antimicrobials are designed to work against target-sites, unexpected effects, such as cytotoxicity and genotoxicity on non-target organisms (Hartmann et al., 1998), and multidrug resistance to sublethal level of antibiotics such as norfloxacin, ampicillin, kanamycin, tetracycline, and chloramphenicol in Escherichia coli and Staphylococcus aureus have been reported (Kohanski et al., 2010). Therefore, due to their increased usage, there is an increasing necessity for suitable treatment methods. To enhance the removal efficiency of antimicrobials, advanced treatment methods, such as ozonation, photolysis using UV and UV/H2O2 (Andreozzi et al., 2006), and c-irradiation (Yu et al., 2008) have been investigated. The concentrations of lincosamides were decreased by more than 80% using the UV/H2O2 combined process in Italy (Andreozzi et al., 2006), and radiolysis process in drinking and wastewater treatment plant in Spain has shown a 70% removal rate for nitroimidazole antimicrobials (Sánchez-Polo et al., 2009). In addition, more than 90% removal rates for antimicrobials were reported for treatment with microfiltration and reverse osmosis (MF/RO) in wastewater treatment plant in Australia (Watkinson et al., 2007). Although a number of studies investigated the removal efficiency and degradation mechanism of various treatment techniques, limited information is available on the alteration in toxicity which occurs through treatment processes. Recently, achieving a reduction in toxicity has been promoted as a key factor in evaluating treatment technologies for the degradation of environmental pollutants (Boxall et al., 2004), as toxicity assessments have been conventionally carried out through determining the removal efficiency of a single compound. Hence, for an integrated assessment of a treatment technology, it is necessary to evaluate the overall change in toxicity, so as to provide for a realistic assessment of the effects on the natural microbial environment. In this study, three representative antimicrobials (TCN, LMC, and SMZ) in water were applied to three different treatment methods: c-irradiation, UV and activated sludge. The methods were then comparatively evaluated in terms of removal efficiencies and toxicity reductions. This information could be used for determining of c-irradiation as an alternative technique to treat recalcitrant compounds such as antibiotics.

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2.2. Decomposition experiments with different treatment methods 2.2.1. c-radiation treatment The degradation experiments, subsequent quantification analysis and toxicity test for each target compound was separately conducted. The test solutions containing 30 mg L1 of each target compound were treated with c-radiation, with absorbed doses of 0, 0.25, 0.5, 0.75, 1, 2, 3 and 4 kGy. c-irradiation was achieved using a high-level 60Co source (Nordion Inc., Canada) at the Korean Atomic Energy Research Institute, Jeongeup, Korea. The dose rates of the source ranged from 6300 to 14 300 Gy h1, which was dependent on the distance of the sample from the source. The initial radioactivity was about 1.47  1017 Bq (=397 949 Ci). The absorbed dose was measured using the alanine-EPR dosimetry system (ISO/ASTM 51607:2003). Before c-irradiation was applied, glass bottles containing the antimicrobial sample solutions were allowed to reach equilibrium at room temperature (22 ± 2 °C). The efficiency of radiolysis was expressed by the radiation-chemical yields of target compounds (G-value), which is defined as the number of molecules of a product formed, or the change in a reactant per 100 eV of absorbed dose. The efficiency of radiolysis was calculated using the following equation (Sánchez-Polo et al., 2009):

G¼

Rð6:023  1023 Þ Dð6:24  1016 Þ

where R is the change in concentration of the solute (mol L1), D the absorbed dose in Gy, 6.023  1023 Avogadro’s number and 6.24  1016 the conversion factor from Gy to 100 eV L1 (Yu et al., 2008).

2.2.2. Ultraviolet (UV) photolysis UV photolysis for degrading target compounds in water sample was conducted using a UV-C lamps (Philips TUV G6T5, 6W), which predominantly emit at 254 nm, in a quartz tube, and installed in a 1 L batch reactor. The electrical power was 6W, which arc length and the diameter of UV was 210 mm and 15 mm, respectively. The UV lamp was placed at the center of the batch reactor. During the experiment, the exterior of the batch reactor was completely wrapped in aluminum foil so as to prevent light penetration. UV irradiation was conducted over a range of UV intensity and penetration distance. The intensity of UVC lamp was determined by H2O2 actinometric measurement (Glaze et al., 1995; Kim et al., 2012). In detail, after illumination of 10 mM H2O2, 4 mL sample was added to 4 mL of solution, containing 2 mL of 0.1 M potassium biphthalate and 2 mL iodide reagent solution (0.4 M potassium iodide, 0.06 M sodium hydroxide and 0.1 mM ammonium molybdate). The UV intensity was measured using spectrophotometer (Cary 50 spectrophotometer, Varian, USA) and calculated by the Lambert–Beer’s law.



d½H2 O2  ¼ /H2 O2 F H2 O2 I0 ð1 - e2:303eH2 O2 b½H2 O2  Þ dt

2. Experimental 2.1. Chemicals and reagents TCN, LMC and SMZ hydrochloride were purchased from Sigma (St. Louis, MO, USA). The chemical structures and physicochemical properties of the target compounds are shown in Table 1. HPLC grade acetonitrile was purchased from Burdick & Jackson (Muskegon, MI, USA), and HPLC grade methanol and water were purchased from Fisher Scientific (Pittsburgh, PN, USA). All other reagents used in this study were of analytical grade.

where /H2 O2 is the quantum yield of H2O2 for direct photolysis, F H2 O2 is the fraction of UV irradiation absorbed by a light absorbing species. I0 is the initial UV intensity, eH2 O2 is the molar extinction coefficient of H2O2 at 254 nm, b is the light path of the photo-reactor. As a result, UV irradiation intensity was 1.91  106 Einstein L1 s1 and samples were exposed to UV light for specific time intervals (e.g., 0, 5, 15, 30, 60 and 90 min), with an initial target compound concentration of 30 mg L1. All experiments were carried out at atmospheric pressure and room temperature (22 ± 2 °C). The elimination of target compounds by UV followed first order reaction kinetics.

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Table 1 The structures of the target compounds. Compounds

Tetracycline

Lincomycin

Sulfamethazine

Tetracyclines C22H24N2O8 444.44 3.30, 7.68, 9.69 1.19 Tolls (2001)

Macrolide C18H34N2O6S 406.54 7.6 0.56 Lissemore et al. (2006)

Sulfonamides C12H14N4O2S 278.32 2.65, 7.4 0.89 Lissemore et al. (2006)

Structure

Group Formula MW (g mol1) pKa Log Kow Reference

2.2.3. Activated sludge A degradation test using activated sludge was conducted according to the OECD method 301E (OECD, 1993). The activated sludge used as an inoculum was collected from the Gwangju municipal wastewater treatment plant in Gwangju, Korea, with hydraulic retention times (HRT) of 7 h and sludge retention times (SRT) of 10 d. The inoculum was pre-conditioned to the experimental conditions by aerating for 7 d at room temperature, which gave 7290 ± 176 mg L1 of mixed liquid suspended solid (MLSS) at pH 5.90. The inoculum concentration of the activated sludge was 3000 mg L1 of MLSS. For profiling the degradation of the target chemicals according to the decomposition mechanism (e.g. biodegradation and adsorption to activated sludge), three different types of test were conducted to identify biodegradation, adsorption and hydrolysis of target chemicals. Inoculated samples for the overall degradation by activated sludge were prepared by mixing the activated sludge (3000 mg L1 of MLSS) and target compound. To distinguish biodegradation from the other mechanisms of decomposition, the adsorption test was conducted by sterilizing the activated sludge (Autoclave at 121 °C for 20 min). The test of target chemicals in sterilized Milli-Q water under the same condition was conducted for hydrolysis. The rate of biodegradation can be identified by subtracting the portions of absorption and hydrolysis from the overall degradation. Each sample was shake-incubated at 120 rpm and 22 °C in the dark to prevent photodegradation of the target compounds. The total test duration was 14 d, with sampling for chemical analysis and bioassay conducted at 0, 1 h, 5 h, 10 h, 1 d, 2 d, 5 d, 10 d, and 14 d. The samples were filtered immediately using a 0.2 lm (pore size) cellulose acetate filter (Advantec, Toyo Roshi International Inc., Japan), and were stored at 4 °C prior to additional treatments. For the degradation kinetics, a first-order decay equation was applied to fit the biodegradation data. Using this equation, half-life (t1/2) was calculated as (ln2/kbio).

2.3. Bioassay experiment Pseudokirchneriella subcapitata was used as the bioassay species to evaluate the toxicity of samples before and after treatments. P. subcapitata, a bioindicator species to access the toxicity in freshwater environment, is known to be sensitive, well documented and standardized (USEPA, 1994). The growth inhibition test with P. subcapitata was carried out according to the procedure described in the EPA manual (USEPA, 1994). Briefly, 50 mL of the test solution containing algae and each treated antimicrobial sample was prepared in a 100 mL glass flask. The treated samples were diluted to make five different concentrations and control, following the

whole effluent toxicity test (WET) (USEPA, 1994). The initial cell density was approximately 104 cells mL1. The solutions were incubated in a mechanical shaker at 25 ± 1 °C and 120 rpm for 96 h. The cell density was then estimated by measuring the optical density at 683 nm using a UV–Vis spectrophotometer (Shimadzu, Kyoto, Japan) (Gonçalves et al., 2013). Each test was performed in triplicate within a static non-renewal system. To express the toxicity towards green algae, the toxic unit (TU) was used, which was calculated as follows (USEPA, 1991):

Toxic unit ðTUÞ ¼

100 IC50 ð%Þ

where IC50 (%) is the percentage median cell growth inhibition of the treated sample. 2.4. Analytical method The chemical analysis for target compounds after c-irradiation was performed using a high-performance liquid chromatography (HPLC, Agilent 1200 Series, Agilent technologies, Santa Clara, CA, USA) system equipped with a UV detector. The system consisted of a pair of LC pumps, a degasser, an autosampler and a detector. Fifty microliters of the sample was loaded onto the column (XTerra™ RP18, 4.6  250 mm) at 30 °C. The mobile phase was water, containing 0.1% formic acid (A) and acetonitrile (B). The flow rate was 1.5 lL min1. The mobile phase was programmed as follows; A:B = 90:10 from 0 to 1 min, which was then raised linearly to 10:90 from 1 to 15 min, remaining stable from 15 to 18 min, and then linearly returning to 90:10 from 18 to 20 min. The detections of TCN, LMC and SMZ were performed at 350, 210 and 254 nm, respectively (Haagsma and Van De Water, 1985; Augugliaro et al., 2005; Zhou et al., 2009). In addition, liquid chromatography mass spectrometry (LCMS2010EV, Shimadzu, Kyoto, Japan) was used for the chemical analysis after UV and activated sludge treatment. All analytes were separated using 150  2.1 mm ACE 3 C18 with 3 lm particle size. Mobile phase solvents were 1% HCOOH-water (A) and methanol (B) at a flow rate of 0.2 mL min1. The linear gradient program was as follows: 10% B held for 10 min, increased linearly to 90% for 20 min, remaining stable for 5 min, and finally, a 5 min equilibration step at 10% was used, where the total run time per sample was 40 min. An injection volume of 10 lL was used for analysis. Analytes were detected in positive electrospray ionization (ESI) mode and were analyzed in the selected ion mode (SIM), with m/ z 445, 427 and 410 for TCN, m/z 407, 359 and 126 for LMC and m/z 279 and 156 for SMZ. The temperature conditions of the curved desolvation line (CDL) and heat block were 230 °C and

H.Y. Kim et al. / Chemosphere 93 (2013) 2480–2487

200 °C, respectively, with nitrogen as a nebulizer gas at a flow rate of 1.5 L min-1. The voltage of the detector was operated at 1.8 kV. 3. Results and discussion Results from control experiments showed that all target compounds were stable and no significant losses were caused by hydrolysis during the test period (max. 14 d). Therefore, the hydrolysis reaction was neglected for the decomposed behavior of target compounds in this study. 3.1. c- irradiation Results of the gamma treatment showed that the concentration of the parent compound was exponentially decreased with increasing absorbed dose (Fig. 1). More than 99% of the TCN and LMC were radiolytically decomposed with 0.5 kGy, while about 60% of the SMZ was decomposed. The elimination phase of the target compound by c-irradiation was fitted with the first order decay equation. Table 2 summarizes the rate constants, half-doses and correction coefficients (r2). The values of kr for TCN, LMC, and SMZ were 0.742 (±0.074), 4.637 (±0.023), and 1.656 (±0.012) kGy1, respectively (r2 > 0.99). A number of studies have examined c-radiation mediated degradation for various environmental pollutants, including pharmaceutical compounds, which have shown relatively high removal efficiencies. For example, 30 mg L1 of cefaclor (cephalosporin antimicrobials) showed a removal efficiency of more than 80% with an absorbed dose of 0.4 kGy (Yu et al., 2008). Fluoroquinolones antimicrobials have been reported to require approximately 1.7 kGy to be decomposed to 50% of their initial concentrations (Santoke et al., 2009). Nitroimidazoles showed a 70% degradation efficiency with 0.7 kGy (Sanchez-Polo et al., 2009). These diverse removal rates are partially attributable to the differences in the initial concentration which was used (Yu et al., 2008; Santoke et al., 2009), pH of the sample solution (Santoke et al., 2009), as well as the molecular structure of the target analyte (Sánchez-Polo et al., 2009). Interestingly, different nitroimidazoles have identical backbone structures (i.e., azole moiety) and show highly similar degradation rate constants and removal efficiencies, due to the similar affinities of the compounds to hydroxyl radicals (Sánchez-Polo et al., 2009), which is the primary species produced by water radiolysis (Mezyk et al., 2007). It has been reported that water radiolysis leads to the production of reactive oxidizing agents, such as hydroxyl radicals (OH), hydrated electrons ðe eq Þ, hydrogen atoms

1.2

C/C0

0.8 0.6

400

300

200

0.4

Toxic Unit

TU-TCN TU-LMC TU-SMZ TCN LMC SMZ

1.0

0.2 100 0.0 0 0

1

2

3

4

Absorbed Dose (kGy) Fig. 1. Decomposition of TCN, LMC and SMZ by c-irradiation and algal toxicity change to P. subcapitata. The initial concentration was 30 mg L1 and Toxic unit (TU) = 100/IC50 (%).

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(H+), hydrogen peroxide (H2O2), superoxide radical ðO 2 Þ anions and perhydroxyl radicals (HO2) (Mezyk et al., 2007). Furthermore, the addition of hydroxyl radicals to a sulfanilic acid ring structure is known as the predominant degradation mechanism of sulfa drugs such as SMZ (Mezyk et al., 2007). The treatment efficiency of irradiation applied to contaminants should be evaluated based on the radiation-chemical yield (G-value) as well as the removal efficiency. In the present study, the G-value of each target compound decreased with increasing absorbed dose; from 0.751, 0.920 and 0.892 to 0.131, 0.106, and 0.198 for TCN, SMZ, and LMC, respectively, which are consistent with results from other studies (Yu et al., 2008; Sánchez-Polo et al., 2009). The decrease in the G-value can be caused by competition between the remaining parent compounds and the newly produced products of degradation, or by radical–radical recombination, resulting in a reduced availability of radicals (Basfar et al., 2005). In comparison with other treatment methods, c-radiation showed higher removal efficiencies and a non-specific degradative action (i.e., independent of chemical class) for all target compounds, indicating its suitability in spite of high cost as an alternative to conventional methods for the treatment of wastewater containing high concentration of persistent compounds such as antimicrobials. 3.2. Photoassisted degradation by UV The removal efficiencies of UV treatment for aqueous solutions containing 30 mg L1 of the target antimicrobials are shown as a function of the irradiation time in Fig. 2. About 80% of the TCN and 90% of the SMZ were decomposed in 60 and 90 min, respectively, whereas only about 30% of the LMC was photodegraded in 90 min. The elimination of the target compounds by UV also followed first order reaction kinetics. The degradation rate constants (kUV) for TCN, LMC and SMZ were estimated to be 0.019 (±0.003), 0.004 (±0.0004) and 0.019 (±0.006) min1, and the half-lifes were estimated to be 36.87, 182.41 and 37.27 min, respectively, revealing that LMC is the least degradable compound under UV treatment. Addamo et al. (2005) reported that 70% of TCN was removed at 5 h irradiation; whereas only a 20% reduction in the LMC content was achieved, without the addition of any photocatalysts. Antimicrobials can be degraded directly or indirectly by reactive species, such as hydroxyl radicals, which are generated by oxidation of water or dissolved oxygen in solution (Boreen et al., 2004; Abellán et al., 2007). Photolysis occurs when a chemical absorbs light, inducing a photochemical reaction. When the absorption spectra of substance and emission spectra of light are overlapped, the degradation rate is higher than that of substance which is not overlapped (Addamo et al., 2005). Similarly, the degradation of sulfamethoxazole was reduced to half after removing the overlapping area with optical filter (Abellán et al., 2007). In the present study, the absorption spectra of TCN and SMZ in aqueous solution and in the UV-C emission range (240–260 nm) were partially overlapped, whereas the absorbance of LMC was relatively lower than that of other target compounds (Fig. 3). Therefore, the various degradation rates of target compounds can be explained by differences in their absorption spectra. 3.3. Decomposition by activated sludge The degradation profiles of the target compounds by activated sludge are shown in Fig. 4 and Table 2. The total degradations of target compounds in inoculum samples were approximately 95.7%, 0% and 99.9% for TCN, LMC and SMZ in 14 d, respectively. However, considering the contributions of adsorption and hydrolysis, the removal rates caused by biodegradation were approximately 6.57%, 0%, and 99.9% for TCN, LMC and SMZ, respectively,

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Table 2 Values of the kinetic constants and the removal efficiencies of c-irradiation, UV and activated sludge treatment. Compound

a b c

c-irradiation

UV irradiation

Removal efficiency (% at 1 kGy)

Dose constanta (kc, kGy1)

Half dose (d1/2, kGy)

r

TCN

99.3

0.146

LMC

100

SMZ

80.8

4.742 (±0.074) 4.637 (±0.247) 1.656 (±0.012)

2

Activated sludge

Removal efficiency (% at 60 min)

Rate constanta (kUV, min1)

0.99

78.4

0.149

0.99

22.1

0.419

0.99

80.7

0.019 (±0.003) 0.004 (±0.0004) 0.019 (±0.006)

Half-life (t1/2, min)

r

2

b

Removal efficiency (% at 14 d)

Rate constanta (kbio, d1)

Halflife (t1/ 2, d)

r2

36.87

0.96

95.7

3.433 (±0.998)

0.253

0.94

182.41

0.96

0.0

–

–

–

37.27

0.87

99.9

0.377c (±0.206)

4.63c

0.97c

Constant (±standard error). Rate constant for activated sludge was calculated with inoculum samples (biodegradation + adsorption to activated sludge). Calculated except for the induction period.

500

1.0

C/C0

0.8

400

300

0.6 200

Toxic Unit

TU-TCN TU-LMC TU-SMZ TCN LMC SMZ

1.2

0.4 100

0.2 0.0 0

20

40

60

0 100

80

Time (min) Fig. 2. Decomposition of TCN, LMC and SMZ by UV and algal toxicity change to P. subcapitata. The initial concentration was 30 mg L1.

2.5 TCN LMC SMZ

Absorbance

2.0

1.5

1.0

0.5

0.0 200

250

300

350

400

450

500

Wavelength (nm) Fig. 3. Absorption spectra of TCN, LMC, and SMZ compared with the emission wavelength range of the UV-C lamp (area plot). The initial concentration of target compounds was 30 mg L1.

which showed highly antimicrobial class-dependent degradation. Adsorption rather than biodegradation was a significant pathway for TCN, whereas SMZ was mostly degraded by biodegradation, and was not adsorbed to activated sludge. LMC was not eliminated with either of the removal mechanisms. Activated sludge showed a compound-specific degradation tendency. This result was consistent with the study of Pérez et al. (2005), which revealed that antimicrobials with an identical backbone structure showed similar degradation trends.

As shown in Fig. 4(a), there was no change in SMZ concentration in sorption controls (sterilized sludge). This can be explained by its low hydrophobicity and ionizability (Table 1). Sulfonamide class antimicrobials have two ionizable functional groups, indicating two different pKa values: pKa1 (2.65) describes the protonation of an ammonia group and pKa2 (7.4) describes the deprotonation of the sulfonamide (SO2NH) moiety (Table 1). Theoretically, under the test conditions (pH 7.39 ± 0.03), half of the SMZ should be present in the ionized form, which acquires enhanced mobility in the aqueous phase. This hydrophilic species of SMZ could be responsible for the lower sorption to biomass (Boxall et al., 2002). Thus, it is unlikely that sorption is an effective mechanism for SMZ removal in activated sludge treatment. Conversely, SMZ was efficiently removed by microbial activity. The degradation profile of SMZ showed an initial lag phase of 3 d at 22 °C. According to (Ingerslev and Halling-Sørensen, 2000), the observed lag period was 7–10 d in the degradation test of sulfonamide class antimicrobials. This is explicable by the lag phase induced by the required time period for the growth of specific sulfonamide-degrading microorganisms (Zhang et al., 2012), or that for the activation of appropriate enzymes (Park and Choung, 2007) in the activated sludge. This was demonstrated by re-addition of the same target compounds immediately after completion of the first degradation, which did not show the same induction period (Ingerslev and Halling-Sørensen, 2000). The removal of TCN in solution was mainly achieved by adsorption rather than by biodegradation (Fig. 4(b)). Approximately 89% of the TCN was significantly and rapidly eliminated in the adsorption test for 14 d, while only 6.6% biodegradation was observed. The non-biodegradability of TCN observed in this study was in agreement with the results of previous investigations (Kim et al., 2005; Li and Zhang, 2010), indicating that adsorption was the major mechanism for TCN degradation, rather than biodegradation, despite the low Kow value and high water solubility (Kim et al., 2005; Li and Zhang, 2010). Figueroa et al. (2004) explored the sorption mechanism of tetracycline antimicrobials to clays, speculating the ionic interaction and surface complexation of zwitterions of TCN. TCN has three functional groups which determine its speciation according to pH (pKa1 = 3.30, pKa2 = 7.68, and pKa3 = 9.69, Table 1). Under the neutral pH conditions in the present study, monoanion and zwitterionic TCN are the dominant species, which have the higher binding capacity with the multivalent ions in the medium. The metal-TCN complexes can sorb strongly to organic matter by cation bridging between negatively charged TCN and organic matter (MacKay and Canterbury, 2005). For example, degradation studies with 3600 and 1000 mg L1 of a biomass inoculum required equilibrium times of only 1 and 10 h, respectively, to degrade 250 lg L1 of TCN (Kim et al., 2005). Similarly, the half-life of tetracycline was determined to be about 6 h with 3000 mg L1 of inoculum biomass in this study. However, even though the removal caused by adsorption was relatively high, the potential

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1.6

35 TU-SMZ Activated sludge Activated sludge (sterilized)

1.4 1.2

(a)

0.6

15

0.4

Toxic Unit

C/C0

20

0.8

10

0.2 5 0.0 0 2

4

6

8

10

12

14

Time (d) 1.2

160 TU-TCN Activated sludge Activated sludge (Sterilized)

C/C0

100 80 0.4

60

0.2

40

0.0

20 6

8

10

12

14

Time (d) 1.6 TU-LMC Activated sludge Activated sludge (sterilized)

1.4 1.2

C/C0

1.0

200

150

0.8 100

0.6

(b)

0.4

0.6

0.8

1.0

1.2

TCN LMC SMZ

1.0

0.8

0.6

0.4

0.2

0.0

Toxic Unit

(c)

0.2

Observed relative algal cell growth

0 4

0.2

0.0

0.6

2

0.4

120

0.8

0

0.6

0.0

Expected relative algal cell growth

1.0

0.8

140

Toxic Unit

(b)

TCN LMC SMZ

1.0

25 1.0

0

1.2

30

Expected relative algal cell growth

(a)

0.0

0.2

0.4

0.6

0.8

1.0

Observed relative algal cell growth

0.4 50

0.2

(c) TCN LMC SMZ

1.0

0.0 2

4

6

8

10

12

14

Time (d) Fig. 4. Degradation rates and algal toxicity changes of the test compounds with an initial concentration of 30 mg L1 for SMZ (a), TCN (b) and LMC (c), and the contributions of the adsorption (sterilized activated sludge treated) and adsorption and biodegradation together (activated sludge treated).

ecological effect of non-biodegraded TCN should be considered (Chander et al., 2005). In contrast, the removal by activated sludge, including adsorption to sludge and biodegradation by microbes, could not be achieved in the case of LMC (Fig. 4(c)). LMC was eliminated by neither biodegradation nor adsorption to the sludge. In the biodegradation test, LMC was also confirmed as the least degradable compound (García-Galán et al., 2009). This may be attributed to the suppression of LMC-degrading microbial species during the initial period of the test, which was likely due to toxicity. The inoculated biomass could be insufficient, or the amount of LMC in the activated sludge could be too much for the LMC-degrading microbes, as observed elsewhere (Pérez et al., 2005).

Expected relative algal cell growth

0 0

0.8

0.6

0.4

0.2

0.0 0.0

0.2

0.4

0.6

0.8

1.0

Observed relative algal cell growth Fig. 5. Comparison of the expected and observed cell growths after c-radiation (a), UV treatment (b), and degradation by activated sludge (c). The solid line indicates the equality of the observed and expected results.

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3.4. Comparison of toxicity change The toxicity change of target compounds during each treatment was evaluated by an algal growth test using P. subcapitata. According to the results from a previous study (Kim et al., 2009), P. subcapitata exhibited higher sensitivity toward the toxicity of antimicrobials than Daphnia magna or Vibrio fischeri, and the median cell growth inhibition concentrations (IC50) of TCN, LMC and SMZ were 0.167 ± 0.107, 0.439 ± 0.052 and 1.246 ± 0.852 mg L1, respectively. The change in the algal toxicity before and after each treatment is depicted in Figs. 1, 2 and 4. Overall, total algal toxicity was decreased with increasing duration and intensity of each treatment. c-irradiation demonstrated the highest reduction in total algal toxicity. With an adsorbed dose of 1.0 kGy, which is one of the normal doses applied in the treatment system, the toxicities were reduced by 94.9%, 99.1% and 69.9% for TCN, LMC and SMZ, respectively. In the case of UV treatment with a 1 h irradiation time, the toxicities were reduced by 73.8%, 68.2% and 79.6%, respectively. For the degradation by activated sludge, 95.8%, 38.8% and 95.8% reductions of the algal toxicities of TCN, LMC and SMZ occurred after 14 d, respectively (Fig. 4). The reduced rates of toxicity were in the order: c-irradiation > activated sludge > UV, for TCN; c-irradiation > UV > activated sludge for LMC; and activated sludge > UV > c-irradiation for SMZ. The expected and observed cell growths are compared in Fig. 5. To estimate the expected toxicity of the target compounds, the remaining parent compound after each treatment was quantified, and the expected cell growth was determined using dose-response curves for each compound (Fig. S1, y = 0.0188 + 1.1207e6.4531x for TCN, y = 0.0618 + 1.1688e2.3251x for LMC and y = 0.0809 + 0.9522e0.6932x for SMZ, where x and y indicate the concentration of target compounds and the relative cell growth, respectively). The solid line indicates the equality of the observed and expected results (1:1 correspondence). If observed cell growth was lower than expected, it indicated that the algal toxicity value was an underestimation, being increased by unknown compounds (i.e., by-products) after treatment. The significance of the difference between observed and expected algal toxicity was analyzed using the student’s t-test (Table 3). As a result, c-irradiated LMC and SMZ, and UV treated SMZ showed no differences between expected and observed cell growths (p > 0.05). Identical expected and observed cell growth indicated that there were no components other than the parent compound inducing toxicity after treatment. Therefore, the reduced toxicity towards P. subcapitata was adequately represented by the concentrations of the target parent compounds remaining. However, all compounds treated with activated sludge, UV treated TCN and LMC, and c-irradiated TCN were significantly different (p < 0.05). In particular, lower cell growth was observed for c-irradiated TCN, UV and active sludge treated LMC, and activated sludge treated SMZ than expected (Table 3). These results indicate the formation of byproducts or transformed products during the treatment process, which were of a comparable or greater toxicity than their parent compounds. Toxicity enhancements have previously been observed after various treatments, such as UV, UV/ H2O2 degradation (Yuan et al., 2011), and c-radiation (Sánchez-Polo et al., 2009). The toxicity assessed by V. fischeri was initially increased, and then decreased with increasing intensity of UV and c-radiation. In addition, some products of phototransformation exerted a greater toxicity on Ceriodaphnia dubia than the parent compound. The increased toxicity after treatment was explained by the changed substituent group of the parent compound (Yuan et al., 2011). The changed substituent group might reduce steric resistance due to a reduction in size, resulting in membrane greater ease of cell penetration than the parent compound (Lu et al., 2002), and consequently leading to increased toxicity.

Table 3 p Values of significant tests of differences in the observed and expected algal cell growth after each treatment of target compounds. The values in parentheses represent the t-statistics, and negative values indicate where observed cell growth was lower than expected cell growth, while positive values indicate where observed cell growth was higher than expected. Compounds

TCN LMC SMZ

p-Value

c-irradiation

UV irradiation

Activated sludge

0.023 (2.42) 0.354 0.777

0.046 (2.09) 0.000 (7.17) 0.907

0.000 (5.17) 0.000 (3.85) 0.038 (2.19)

Conversely, for UV and activated sludge treated TCN, the observed cell growth rates were higher than expected. One explanation for increased algal growth was that degradation products, such as nitro compounds, can be used as algal nutrients (Wei et al., 1998). The toxicity change after the treatment is induced not only by the reduced concentration of parent compound, but also by the transformed products. Though the transformation mechanisms for target compounds in each treatment process are unclear, the deviation of the observed algal toxicity from that which was expected strongly implies the presence of the products which altered the overall algal toxicity. 4. Conclusions In the present study, c-irradiation, UV treatment and activated sludge were comparatively evaluated in terms of the degradability and toxicity change of antimicrobials. Of these treatment methods, c-irradiation was shown to have the highest removal efficiencies for all of the target compounds. Conversely, UV and activated sludge treatment showed compound-dependent removal efficiencies. The concentration of target compounds was dramatically reduced by all the treatments, except for LMC, particularly under treatment with UV and activated sludge. In the case of activated sludge treatment, biodegradation and adsorption were the main degradation mechanisms, which were also degraded in a compound-dependent manner. The total algal toxicity after the degradation process was significantly decreased in comparison with the untreated control. However, in some cases, the observed toxicity was higher than the expected one, which indicates the existence of byproducts or transformed compounds in the treated sample. The results of this study indicate the applicability of gamma irradiation as an alternative technique for degrading recalcitrant compounds including antibiotics. Acknowledgements This work was supported by a Nuclear Research & Development Program and Mid-career Researcher Program through NRF grant funded by the MEST (2013R1A2A2A03014187). Appendix A. Supplementary material Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.chemosphere. 2013.08.091. References Abellán, M.N., Bayarri, B., Giménez, J., Costa, J., 2007. Photocatalytic degradation of sulfamethoxazole in aqueous suspension of TiO2. Appl. Catal. B – Environ. 74, 233–241. Addamo, M., Augugliaro, V., Di Paola, A., García-López, E., Loddo, V., Marcí, G., Palmisano, L., 2005. Removal of drugs in aqueous systems by photoassisted degradation. J. Appl. Electrochem. 35, 765–774.

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