Influence of pH on the sonolysis of ciprofloxacin: Biodegradability, ecotoxicity and antibiotic activity of its degradation products

Chemosphere 77 (2009) 291–295

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Technical Note

Influence of pH on the sonolysis of ciprofloxacin: Biodegradability, ecotoxicity and antibiotic activity of its degradation products Evelien De Bel a, Jo Dewulf a,*, Bavo De Witte a, Herman Van Langenhove a, Colin Janssen b a b

Research Group EnVOC, Department of Organic Chemistry, Faculty of Bioscience Engineering, Ghent University, Coupure Links 653, B-9000 Ghent, Belgium Laboratory of Environmental Toxicology and Aquatic Ecology, Faculty of Bioscience Engineering, Ghent University, J. Plateaustraat 22, B-9000 Ghent, Belgium

a r t i c l e

i n f o

Article history: Received 9 June 2009 Received in revised form 16 July 2009 Accepted 16 July 2009 Available online 14 August 2009 Keywords: Ultrasonic irradiation pH Biodegradability Antibiotic activity Pseudokirchneriella subcapitata Ciprofloxacin

a b s t r a c t The presence of antibiotics in the aquatic environment has raised concerns due to the potential risk for the emergence or persistence of antibiotic resistance. Antibiotics are often poorly degraded in conventional wastewater treatment plants. In this study, sonolysis at 520 kHz and 92 W L 1 was used for the degradation of the fluoroquinolone antibiotic ciprofloxacin. In a first experiment at pH 7, 57% of the ciprofloxacin (15 mg L 1) was degraded after 120 min of ultrasonic irradiation at 25 °C. pH proved to be an important parameter determining the degradation rate, since the pseudo first order degradation constant increased almost fourfold when comparing treatment at pH 7 (0.0058 min 1) and pH 10 (0.0069 min 1) with that at pH 3 (0.021 min 1). This effect can be attributed to the degree of protonation of the ciprofloxacin molecule. The BOD/COD ratio of the solutions, which is a measure for their biodegradability, increased from 0.06 to 0.60, 0.17, and 0.18 after 120 min of irradiation depending on the pH (3, 7, and 10, respectively). The solution treated at pH 3 can even be considered readily biodegradable (BOD/ COD > 0.4). The antibiotic activity against Escherichia coli (G ) and Bacillus coagulans (G+) of the treated solutions also reduced after sonolysis. The highest decrease was again found when irradiated at pH 3. In contrast, ecotoxicity of the solutions to the alga Pseudokirchneriella subcapitata increased 3- to 10-fold after 20 min of treatment, suggesting the formation of toxic degradation products. The toxicity slowly diminished during further treatment. Ó 2009 Elsevier Ltd. All rights reserved.

1. Introduction Yearly, thousands of tons of pharmacologically active compounds (PhAC) are used not only to treat human and animal illnesses, but also in farming and aquaculture. Depending on the nature of the PhAC, up to 95% of the administered dose can be excreted unchanged or as active metabolites (Calamari et al., 2003). Since the elimination of these compounds in conventional wastewater treatment plants (WWTP) is often incomplete, substantial quantities of pharmaceuticals are released in the environment (Ternes, 1998). More than 80 compounds, pharmaceuticals and several drug metabolites, have already been detected in municipal sewage and surface waters (Heberer, 2002). Generally, these compounds are detected in low concentrations (ng–lg L 1), but since pharmaceuticals are designed to cause a biological effect at low doses, ad-

* Corresponding author. Tel.: +32 9 264 59 49; fax: +32 9 264 62 43. E-mail address: [email protected] (J. Dewulf). 0045-6535/$ - see front matter Ó 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2009.07.033

verse effects on aquatic organisms cannot be excluded (Calamari et al., 2003). Antibiotics are of particular concern because the continuous exposure of bacteria to even small concentrations of these substances could lead to the emergence or persistence of antibacterial resistance (Kummerer and Henninger, 2003). Furthermore, antibiotic drugs can disrupt the wastewater treatment process and the microbial ecology in surface waters (Al-Ahmad et al., 1999). Huang et al. (2001) concluded that sulfonamides and fluoroquinolones are the antibiotic classes most likely to persist in the aquatic environment. It has also been found that the genotoxicity of the wastewater of a large Swiss university hospital could be largely attributed to fluoroquinolones (Hartmann et al., 1999). The most prescribed fluoroquinolone in Europe is ciprofloxacin (CIP) (Ferech et al., 2006). CIP is not biodegradable (Al-Ahmad et al., 1999). Therefore, the relatively high removal efficiencies in WWTPs are mainly attributed to sorption, rather than degradation (Lindberg et al., 2006). In final WWTP effluents, CIP has been detected in concentrations up to 5.6 lg L 1 (Batt et al., 2006). Ultrasonic irradiation and advanced oxidation processes in general have shown great potential for the degradation of toxic organic

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chemicals in wastewater (Adewuyi, 2001; Parsons and Williams, 2004). The mechanism of sonolysis is based on cavitation which is the formation, growth, and sudden collapse of bubbles in liquids (Adewuyi, 2001). The most widely accepted theory to explain the effects caused by cavitation is the ‘‘hot spot” theory (Mason and Lorimer, 2002). This theory that when the cavitation bubbles implode, temperatures as high as 5200 K and pressures higher than 100 MPa inside the collapsing cavity are generated. About 1900 K is observed at the interfacial region between the solution and the collapsing bubble (Suslick et al., 1986). Two possible degradation routes are proposed. First, the contaminant can undergo thermal degradation inside the cavity and in the interfacial region (cavity–liquid). Second, free radicals (mainly OH), formed due to thermolysis of water, can react with the contaminant in the interfacial region or in the bulk solution (Mason and Lorimer, 2002). To our knowledge, there are only a few reports on the sonolysis of pharmaceuticals (Fu et al., 2007; Suri et al., 2007; Hartmann et al., 2008; Méndez-Arriaga et al., 2008) and we have found none on the sonolysis of antibiotics. In this paper, the application of sonolysis for the degradation of CIP in water is examined. The influence of pH on the degradation rate is studied and the changes in biodegradability, ecotoxicity and antibiotic activity of the solution are evaluated as a function of pH. The identification of degradation products is outside the scope of this study. 2. Materials and methods 2.1. Chemicals and stock solutions Ciprofloxacin HCl (99.5%) was purchased from MP Biomedicals Inc. All stock and buffer solutions were prepared with deionized water. All chemicals used were of reagent grade and were used without further purification. 2.2. Experimental setup 150 mL of a 15 mg L 1 CIP solution (45.3 lM) was sonicated with a 520 kHz Undatim Ortho Reactor. The calorimetrically determined effective ultrasonic power (Mason et al., 1992) was 92 ± 2 W L 1 (n = 3). The temperature in the reactor was maintained by circulating water through the cooling jacket. The steady state reaction temperature of 25.0 ± 0.5 °C was reached in less than 20 min of sonication (Fig. 1). The pH of the solution was adjusted to the desired value by adding the appropriate phosphate buffer

(40 mM). Phosphates are known to react at relatively slow rates with hydroxyl radicals (Buxton et al., 1988). Hence, these have been selected as the most suitable pH buffering agents. 2.3. Analytical procedures CIP was measured by liquid chromatography with a photodiode array detector (Surveyer, Thermo Finnigan). A Luna C18(2) column (150 mm  3.0 mm, 3 lm, Phenomenex) was used with a mobile phase containing 88.5% water (0.1% formic acid) and 11.5% acetonitrile. Quantification of CIP took place at 278 ± 4.5 nm (De Witte et al., 2009). 2.4. Biodegradability, ecotoxicity and antibiotic activity COD was determined with commercially available test kit (Macherey–Nagel, Düren, Germany). Biodegradability tests (BOD) were based on the standard guideline OECD 301D (closed bottle test) (OECD, 1992). Effluent of the WWTP of the Maria Middelares hospital (Ghent, Belgium) was used as inoculum. Toxicity controls (sample + 2 mg L 1 sodium acetate) showed no significant toxicity of the samples towards the inoculum. Ecotoxicity of the parent compound and its degradation products was assessed using the 72 h-growth inhibition test with the green alga Pseudokirchneriella subcapitata which was conducted according to OECD guideline 201 (OECD, 2006). Antibacterial activity was examined in an agar diffusion test based on Palominos et al. (2008). Bacterial strains of Escherichia coli (G ) and Bacillus coagulans (G+) were inoculated onto nutrient agar 2% (Biokar Diagnostics, France). Agar plates were divided into four compartments: three compartments were used for injection of 20 lL sample into holes of 0.64 cm diameter and one control compartment without addition of a CIP solution. The test was executed in triplicate. Plates were incubated for 48 h at 30 °C (B. coagulans) and 37 °C (E. coli) and the inhibition zone diameter was used as a semi-quantitative indicator of residual antibacterial activity. Samples for ecotoxicity testing and COD determination were taken every 20 min. BOD and antibiotic activity were tested after 0, 60 and 120 min of sonolysis. The solutions for the ecotoxicity tests, BOD measurements and antibiotic activity testing were four times diluted and the pH was adjusted to 7 by adding an appropriate amount of HCl or NaOH. The blanks contained the same buffer as the treated solutions. 2.5. Statistical procedures The pseudo first order reaction constants k1 were determined by means of linear regression (S-PLUS 7.0). The equality of slopes of the regression lines was statistically tested using the method described by Powell and Jordan (1997). 3. Results and discussion 3.1. Degradation of ciprofloxacin

Fig. 1. Degradation of ciprofloxacin at pH 7 (C0 = 15 mg L 92 W L 1).

1

, 25 °C, 520 kHz,

The first experiment conducted at pH 7 showed that CIP is degraded by sonolysis (Fig. 1). After 120 min of ultrasonic irradiation, 57 ± 2% (n = 3) of the CIP was degraded. Statistical analysis of the data showed that CIP degradation follows first order kinetics with a reaction constant k1 of 0.0058 ± 0.0002 min 1 (R2 = 0.989, n = 3) and a half-life time of 120 min. For this regression, only the data acquired at steady state temperature (P20 min) were considered. Similar reaction kinetics has been reported for diclofenac (a nonsteroidal anti-inflammatory drug), estrogen compounds and many other pollutants (Adewuyi, 2001; Fu et al., 2007; Hartmann et al., 2008).

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3.2. Influence of pH

3.3. Biodegradability, ecotoxicity and antibiotic activity

It has been shown that the pH, which results in modification of physical properties (including charge) of molecules with ionisable functional groups, can play an important role in the sonochemical degradation of chemicals (Ince and Tezcanli-Guyer, 2003; Behnajady et al., 2008). Hence this parameter was selected as the main process variable evaluated in this study. Initially, degradation experiments were executed in triplicate at three pH values (3, 7, and 10). Degradation at pH 3 (k1 = 0.021 ± 0.0002 min 1) was almost four times faster than at pH 7. The degradation constant at pH 10 (k1 = 0.0069 ± 0.0003 min 1) was also significantly higher than at pH 7. This can be explained by the different CIP forms due to (de)protonation (Fig. 2). To elucidate the influence of pH, degradation constants were determined in the 2–11 pH range. Results are shown in Fig. 3 as well as the theoretical form in which CIP appears at each pH value. A clear correlation between the charge of CIP and the degradation rate can be observed. Degradation is clearly faster when the main part of the CIP molecules carries an overall positive charge. These positively charged molecules will accumulate at the negatively charged liquid–bubble interface (Watmough et al., 1992) where the concentration of reactive radicals and the temperature are higher. Hence, the degradation will be faster. The small increase in degradation rate at pH > 9 can be attributed to the electrophilic nature of the hydroxyl radical (Buxton et al., 1988).

Although the CIP concentration decreases, it is unlikely that the parent compound is completely mineralized. For this reason biodegradability, ecotoxicity and antibiotic activity of the treated solutions were studied after degradation at pH 3, 7, and 10. As expected, there was only a minor decrease of COD after 120 min of treatment at the three pH values, there was, however, an important increase in BOD. This resulted in a higher BOD/COD ratio, which is a good measure for biodegradability. As shown in Fig. 4, the BOD/COD ratio increases from 0.06 (1.7 mg L 1/27.4 mg L 1) to 0.60 (12.2 mg L 1/20.3 mg L 1), 0.17 (4.0 mg L 1/23.2 mg L 1), and 0.18 (4.3 mg L 1/24.4 mg L 1) after 120 min of treatment at pH 3, 7, and 10, respectively. Since the BOD/COD ratio of the solution treated at pH 3 exceeds 0.4 it may even be considered readily biodegradable (Chamarro et al., 2001). We also examined the change in ecotoxicity of the CIP solutions during treatment. Since, at environmentally relevant concentrations acute toxicity is unlikely, chronic toxicity tests are more relevant when assessing the environmental risks of pharmaceuticals (Jones et al., 2001). For this reason, the chronic toxicity of CIP towards the green alga P. subcapitata was assessed. We observed a dramatic increase of the growth inhibition after 20 min of ultrasonic irradiation at the different test conditions (Table 1) which is most likely due to the formation of highly toxic degradation products. A similar effect was also described during the ozonation of oxytetracycline (an antibiotic) (Li et al., 2008) and benzafibrate

Fig. 2. (De)protonation equilibria of ciprofloxacin (pKa’s according to Lin et al. (2004) and De Witte et al. (2007)).

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Fig. 3. Pseudo first order degradation constants for CIP between pH 2 and 11 and the theoretical form in which CIP appears at this pH.

(lipid regulator) (Dantas et al., 2007). At pH 3 the toxicity clearly decreased again when the solutions were treated longer than 20 min, which suggests the toxic degradation products are destroyed by further sonolysis. Also at pH 10 a moderate decrease of growth inhibition was observed for longer treatment times. At pH 7 no significant decrease could be detected. The slower reduc-

Table 1 Effect of irradiation time and pH values on the percentage of growth inhibition of P. subcapitata (n = 3). Irradiation time (min)

0 20 40 60 80 100 120

% Inhibition pH 3

pH 7

pH 10

6±3 60 ± 1 43 ± 1 46 ± 8 32 ± 6 38 ± 6 32 ± 4

18 ± 6 55 ± 1 53 ± 8 43 ± 7 46 ± 3 42 ± 1 49 ± 8

12 ± 3 58 ± 3 54 ± 3 54 ± 2 53 ± 1 49 ± 3 36 ± 2

Table 2 Bacterial growth inhibition zone diameter for ciprofloxacin sonolysis at pH 3, 7, and 10 at different irradiation times after 48 h of incubation (>1.5 ++, 1–1.5 cm +, <1 cm ±, no inhibition 0). Species

Irradiation time (min)

pH 3

pH 7

pH 10

B. coagulans

0 60 120 0 60 120

++ + ± + ± 0

++ + + + ± ±

++ + + + ± ±

E. coli

tion in ecotoxicity at pH 7 and 10 is in accordance with the lower degradation rate at pH values. Finally, the antibiotic activity of the treated solutions was investigated. The inhibition zone diameters (Table 2) were used as a semi-quantitative indication for the antibacterial activity. For both species, there was a clear reduction in antibacterial activity after ultrasonic irradiation. As could be expected from the degradation constants, the largest decrease was found for the solution treated at pH 3. The contradiction between the increasing toxicity towards P. subcapitata and the decreasing antibiotic activity can possibly be explained by the fact that CIP inhibits typical bacterial enzymes (DNA gyrase and DNA topoisomerase IV) (Robinson et al., 2005), which may not be present in algae. 4. Conclusions Fig. 4. COD, BOD, and BOD/COD ratio evolution during 120 min of US irradiation at (a) pH 3, (b) pH 7, and (c) pH 10.

CIP in aqueous solution is degraded when it is exposed to ultrasonic irradiation. The degradation rate is strongly dependent on

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the pH of the solution. pH determines the degree of protonation of CIP and positive charges on the CIP molecule seem to promote ultrasonic degradation, due to accumulation at the negatively charged liquid–bubble interface. Though there is only a minor decrease in COD after treatment, the biodegradability of the solution increases significantly. A decrease of antibacterial activity due to ultrasonic treatment was also observed, suggesting a reduced risk for emergence of antibacterial resistance in the environment. However, this study also demonstrated that the formed reaction products were considerably more ecotoxic than CIP. Further assessment of the degradation pathway is needed to clarify this increase in ecotoxicity. Acknowledgments This work is supported by the BOF research fund (B/07776/02) from the Ghent University, which is gratefully acknowledged. The authors would like to thank the Laboratory of Industrial Microbiology and Biocatalysis (LIMAB) at Ghent University for their cooperation for the antibacterial activity tests. References Adewuyi, Y.G., 2001. Sonochemistry: environmental science and engineering applications. Ind. Eng. Chem. Res. 40, 4681–4715. Al-Ahmad, A., Daschner, F.D., Kummerer, K., 1999. Biodegradability of cefotiam, ciprofloxacin, meropenem, penicillin G, and sulfamethoxazole and inhibition of waste water bacteria. Arch. Environ. Contam. Toxicol. 37, 158–163. Batt, A.L., Bruce, I.B., Aga, D.S., 2006. Evaluating the vulnerability of surface waters to antibiotic contamination from varying wastewater treatment plant discharges. Environ. Pollut. 142, 295–302. Behnajady, M.A., Modirshahla, N., Tabrizi, S.B., Molanee, S., 2008. Ultrasonic degradation of Rhodamine B in aqueous solution: influence of operational parameters. J. Hazard. Mater. 152, 381–386. Buxton, G.V., Greenstock, C.L., Helman, W.P., Ross, A.B., 1988. Critical review of rate constants for reactions of hydrated electrons, hydrogen atoms and hydroxyl radicals (OH/O ) in aqueous solution. J. Phys. Chem. Ref. Data 17, 513–886. Calamari, D., Zuccato, E., Castiglioni, S., Bagnati, R., Fanelli, R., 2003. Strategic survey of therapeutic drugs in the rivers Po and Lambro in northern Italy. Environ. Sci. Technol. 37, 1241–1248. Chamarro, E., Marco, A., Esplugas, S., 2001. Use of Fenton reagent to improve organic chemical biodegradability. Water Res. 35, 1047–1051. Dantas, R.F., Canterino, M., Marotta, R., Sansa, C., Esplugas, S., Andreozzi, R., 2007. Bezafibrate removal by means of ozonation: primary intermediates, kinetics, and toxicity assessment. Water Res. 41, 2525–2532. De Witte, B., Dewulf, J., Demeestere, K., De Ruyck, M., Van Langenhove, H., 2007. Critical points in the analysis of ciprofloxacin by high-performance liquid chromatography. J. Chromatogr. A 1140, 126–130. De Witte, B., Dewulf, J., Demeestere, K., Van Langenhove, H., 2009. Ozonation and advanced oxidation by the peroxone process of ciprofloxacin in water. J. Hazard. Mater. 161, 701–708. Ferech, M., Coenen, S., Malhotra-Kumar, S., Dvorakova, K., Hendrickx, E., Suetens, C., Goossens, H., 2006. European surveillance of antimicrobial consumption (ESAC): outpatient quinolone use in Europe. J. Antimicrob. Chemother. 58, 423–427. Fu, H., Suri, R.P.S., Chimchirian, R.F., Helmig, E., Constable, R., 2007. Ultrasoundinduced destruction of low levels of estrogen hormones in aqueous solutions. Environ. Sci. Technol. 41, 5869–5874.

295

Hartmann, A., Golet, E.M., Gartiser, S., Alder, A.C., Koller, T., Widmer, R.M., 1999. Primary DNA damage but not mutagenicity correlates with ciprofloxacin concentrations in German hospital wastewaters. Arch. Environ. Contam. Toxicol. 36, 115–119. Hartmann, J., Bartels, P., Mau, U., Witter, M., Tumpling, W.V., Hofmann, J., Nietzschmann, E., 2008. Degradation of the drug diclofenac in water by sonolysis in presence of catalysts. Chemosphere 70, 453–461. Heberer, T., 2002. Occurrence, fate, and removal of pharmaceutical residues in the aquatic environment: a review of recent research data. Toxicol. Lett. 131, 5–17. Huang, C., Renew, J.E., Semby, K.L., Pinkston, K., Sedlak, D.L., 2001. Assessment of potential antibiotic contaminants in water and preliminary occurrence analysis. Water Resour. Update 120, 30–40. Ince, N.H., Tezcanli-Guyer, G., 2003. Impacts of pH and molecular structure on ultrasonic degradation of azo dyes. In: Ultrasonics International 2003 Meeting, Granada, Spain, pp. 591–596. Jones, O.A.H., Voulvoulis, N., Lester, J.N., 2001. Human pharmaceuticals in the aquatic environment – a review. Environ. Technol. 22, 1383–1394. Kummerer, K., Henninger, A., 2003. Promoting resistance by the emission of antibiotics from hospitals and households into effluent. Clin. Microbiol. Infect. 9, 1203–1214. Li, K., Yediler, A., Yang, M., Schulte-Hostede, S., Wong, M.H., 2008. Ozonation of oxytetracycline and toxicological assessment of its oxidation by-products. Chemosphere 72, 473–478. Lin, C.E., Deng, Y.J., Liao, W.S., Sun, S.W., Lin, W.Y., Chen, C.C., 2004. Electrophoretic behavior and pK(a) determination of quinolones with a piperazinyl substituent by capillary zone electrophoresis. J. Chromatogr. A 1051, 283–290. Lindberg, R.H., Olofsson, U., Rendahl, P., Johansson, M.I., Tysklind, M., Andersson, B.A.V., 2006. Behavior of fluoroquinolones and trimethoprim during mechanical, chemical, and active sludge treatment of sewage water and digestion of sludge. Environ. Sci. Technol. 40, 1042–1048. Mason, T.J., Lorimer, J.P., 2002. Applied Sonochemistry: The Uses of Power Ultrasound in Chemistry and Processing. Wiley-VCH Verlag, Weinheim, Germany. Mason, T.J., Lorimer, J.P., Bates, D.M., 1992. Quantifying sonochemistry – casting some light on a black art. Ultrasonics 30, 40–42. Méndez-Arriaga, F., Torres-Palma, R.A., Pétrier, C., Esplugas, S., Gimenez, J., Pulgarin, C., 2008. Ultrasonic treatment of water contaminated with ibuprofen. Water Res. 42, 4243–4248. OECD, 1992. OECD Guidelines for Testing Chemicals, 301D Closed Bottle Test. Adopted by the council on 17th July 1992. OECD, 2006. Guidelines for Testing Chemicals, 201 Alga, Growth Inhibition Test. Adopted by the council on 23 March 2006. Palominos, R., Freer, J., Mondaca, M.A., Mansilla, H.D., 2008. Evidence for hole participation during the photocatalytic oxidation of the antibiotic flumequine. J. Photoch. Photobio. A: Chem. 193, 139–145. Parsons, S.A., Williams, M., 2004. Introduction. In: Parsons, S.A. (Ed.), Advanced Oxidation Processes for Water and Wastewater Treatment. IWA Publishing, London, UK, pp. 1–6. Powell, N., Jordan, M.A., 1997. Batch leaching data analysis: eradication of time dependency prior to statistical analysis. Miner. Eng. 10, 859–870. Robinson, A.A., Belden, J.B., Lydy, M.J., 2005. Toxicity of fluoroquinolone antibiotics to aquatic organisms. Environ. Toxicol. Chem. 24, 423–430. Suri, R.P.S., Nayak, M., Devaiah, U., Helmig, E., 2007. Ultrasound assisted destruction of estrogen hormones in aqueous solution: effect of power density, power intensity and reactor configuration. J. Hazard. Mater. 146, 472–478. Suslick, K.S., Hammerton, D.A., Cline, R.E., 1986. The sonochemical hot-spot. J. Am. Chem. Soc. 108, 5641–5642. Ternes, T.A., 1998. Occurrence of drugs in German sewage treatment plants and rivers. Water Res. 32, 3245–3260. Watmough, D.J., Shiran, M.B., Quan, K.M., Sarvazyan, A.P., Khizhnyak, E.P., Pashovkin, T.N., 1992. Evidence that ultrasonically-induced microbubbles carry a negative electrical charge. Ultrasonics 30, 325–331.