Environ Sci Pollut Res (2014) 21:4297–4308 DOI 10.1007/s11356-013-2301-x
RESEARCH ARTICLE
Sonochemical degradation of a pharmaceutical waste, atenolol, in aqueous medium K. K. Nejumal & P. R. Manoj & Usha K. Aravind & C. T. Aravindakumar
Received: 10 July 2013 / Accepted: 28 October 2013 / Published online: 5 December 2013 # Springer-Verlag Berlin Heidelberg 2013
Abstract Atenolol is a β-blocker drug and an identified emerging pollutant. Advanced oxidation processes (AOPs) utilise the reaction of a highly oxidising species (hydroxyl radicals, •OH) for the mineralisation of emerging pollutants since conventional treatment methodologies generally fail to degrade these compounds. In the present work, degradation of atenolol was carried out using ultrasound with frequencies ranging from 200 kHz to 1 MHz as a source of hydroxyl radical. The degradation was monitored by HPLC, total organic carbon (TOC) and chemical oxygen demand (COD) reduction and ion chromatography (IC). Nearly 90 % of degradation of atenolol was observed with ultrasound having 350 kHz. Both frequency and power of ultrasound affect the efficiency of degradation. Nearly 100 % degradation was obtained at a pH of 4. Presence of various additives such as sodium dodecyl sulphate, chloride, sulphate, nitrate, phosphate and bicarbonate was found to reduce the efficiency of degradation. Although nearly 100 % degradation of atenolol was observed under various experimental conditions, only
Responsible editor: Hongwen Sun Electronic supplementary material The online version of this article (doi:10.1007/s11356-013-2301-x) contains supplementary material, which is available to authorized users. K. K. Nejumal : P. R. Manoj : C. T. Aravindakumar (*) School of Environmental Sciences, Mahatma Gandhi University, Kottayam 686560, Kerala, India e-mail: [email protected] U. K. Aravind Advanced Centre of Environmental Studies and Sustainable Development, Mahatma Gandhi University, Kottayam 686560, Kerala, India C. T. Aravindakumar Inter University Instrumentation Centre, Mahatma Gandhi University, Kottayam 686560, Kerala, India
about 62 % mineralisation (from TOC and COD measurements) was obtained. Nearly eight intermediate products were identified using high-resolution mass spectrometry (LC-QTOF). These products were understood as the results of hydroxyl radical addition to atenolol. The degradation studies were also carried out in river water which also showed a similar degradation profile. A mechanism of degradation and mineralisation is presented. Keywords Emerging pollutants . Sonolysis . Atenolol . Advanced oxidation processes . Hydroxyl radical . Mineralisation
Introduction Pharmaceuticals and personal care products (PPCPs) under the common banner, emerging pollutants, have gained lot of attention in recent years. Emerging pollutants are a class of compounds in the environment such as endocrine disruptors resulting from the degradation of organic compounds or introduction of medicine in the natural environment. The United States Environmental Protection Agency (USEPA) defines “emerging pollutants as new chemicals without regulatory status and which impact on environment and human health are poorly understood” (Kümmerer 2009a, b; Meyer et al. 2011). These compounds are a source of concern because they are used and released in large quantities and results in the widespread distribution into the environment mainly due to their physical and chemical properties. Even the trace level of PPCPs can cause chronic toxicity, endocrine disruption and the development of pathogen resistance. Wastewater treatment plant effluents and secondarily terrestrial run-offs (roofs, pavement, roads and agricultural land) including atmospheric deposition are the major sources of environmentally relevant emerging contaminants. It is not necessary that these
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chemicals need to be persistent in the environment to cause harmful effects, but it creates an environmental catastrophe due to its continuous introduction into the environment. Many of these emerging contaminants, risk assessment and ecotoxicological data are not available at the moment and hence it is difficult to predict which health effects they may cause on humans, terrestrial and aquatic organisms and the ecosystem (De Andrés et al. 2009; Fernández et al. 2010) The general issue associated with the emerging pollutants is that many of these compounds are resistant to degradation under natural conditions. Even if they undergo some kind of transformation, their metabolites also may not be easily degradable. Conventional methodologies for their removal such as flocculation, coagulation, sedimentation, filtration and disinfection are generally employed in drinking water treatment facilities. These techniques generally remove some of the contaminants, but they are inefficient to degrade these chemicals. Microbial degradation is also inefficient in this case as these compounds primarily resist microbes. In this context, oxidation technologies (generally termed as advanced oxidation technology (AOT)) are widely experimented for the destruction of these chemicals. Advanced oxidation processes (AOPs) are still under experimental stage, particularly in the case of emerging pollutants. AOPs are based on the production of a highly reactive oxidant such as hydroxyl radical in solution state and utilise its unselective oxidation of organic chemicals to carbon dioxide and water (Eq. 1). It is a process of mineralisation.
been reported to exist in surface water (Camacho-Muñoz et al. 2010; Meyer et al. 2011; Winter et al. 2008). The same case study highlighted the various aspects of the existence of atenolol and its possible impact in the environment and health. Some of the β-blockers are reported to undergo photodegradation under simulated sunlight and indirect photodegradation in the presence of humic substances and fulvic acid (Chen et al. 2007; Chen et al. 2009; Liu and Williams 2006). Among various AOPs, ultrasound (frequencies between 20 kHz and 1 MHz)-initiated degradation of organic contaminants is an active research area. Sonolysis is considered as a versatile technique which is used to investigate radicalmediated waste water treatment. Perhaps the major advantage of sonolysis is its safety, cleanliness and energy conservation (Chowdhury and Viraraghavan 2009; Eren 2012). One can come across several reports based on the studies of sonolytic degradation of chlorinated organic compounds, phenolic compounds, organic dyes, pesticides, endocrine disrupting compounds, POPs, perfluorinated chemicals, pharmaceuticals and microcystins (Adewuyi 2001; Chowdhury and Viraraghavan 2009; Song et al. 2006). In the present study, an investigation of the sonochemical degradation of atenolol has been carried out in aqueous medium. The effect of ultrasound frequency, applied power and added salts on the degradation efficiency is investigated. The extent of mineralisation in all these cases is also probed.
Hydroxyl radical ð• OHÞ þ PPCPs→ðintermediatesÞ→CO2 þ H2 O
Materials and methods
ð1Þ The production of the hydroxyl radical is the most important part of this technology. Photolysis of hydrogen peroxide, ozone and titanium dioxide, chemical method such as Fenton reaction and sono- and electrochemical methods are some of the techniques come under this category (Brillas et al. 2009; Sunil Paul et al. 2013; Wang and Xu 2011). The crucial step in this kind of processes is to control the intermediate products of oxidation and finally get all those into carbon dioxide and water. Atenolol falls under the category of beta blockers, a class of drugs mainly used in cardiovascular diseases (Fernández et al. 2010). It is a selective receptor antagonist drug and is widely used in many countries. Due to its extensive use as a drug, beta blockers are recently identified as an emerging pollutant and are found to exist in the environment, particularly in sewage effluents and surface waters. Since beta adrenergic receptors were detected in fish and other aquatic animals, there is a high chance that many physiological processes regulated by these receptors in wild animals can be influenced by the presence of beta blockers. Among various beta blockers, atenolol have
Reagents The atenolol of pure grade was purchased from Sigma Aldrich. Aqueous solution was prepared by using CascadaTM Lab Water systems (18.2 MΩ cm−1). Ferrous ammonium sulphate, hydrogen peroxide and other salts used for this study was purchased from Merck India. Sonolysis The reactions were carried out in a sonoreactor having four variable frequencies (202, 351, 623 and 1,003 kHz). The reactor was carried out in a glass reactor; ultrasound was produced using an L3 ELAC Nautik ultrasound generator powered by an Allied Signal R/F generator (T & C power conversion, Model AG 1006). The temperature was maintained at 25±1 °C by using a water circulator. HPLC analysis The degradation of atenolol (ATL) was monitored by highperformance liquid chromatography (Shimadzu prominence
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UFLC, LC 20 AD) connected with a diode array detector (SPD-M20 A). The eluent consists of 0.1 % formic acid in water/methanol (90:10).
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Chemical oxygen demand (COD) of atenolol solutions before and after the treatment was analysed using standard dichromate method. Most type of organic matters is oxidised by a boiling mixture of chromic acid and sulphuric acid. A sample is refluxed in strongly acid solution with a known excess of potassium dichromate (K2Cr2O7). After digestion, the remaining unreduced K2Cr2O7 is titrated with ferrous ammonium sulphate to determine the amount of K2Cr2O7
degradation of atenolol with respect to time of sonication is shown in Fig. 1. There is a fast disappearance initially with time of irradiation. The same trend is followed at all the selected frequencies. The first-order nature of the degradation pattern is also shown in Fig. 1. These calculated values of the first-order rate constant are 0.01649 min−1, 0.04476 min−1, 0.03717 min−1 and 0.0372 min−1 for 200, 350, 620 kHz and 1 MHz, respectively (these are summarised in open resource ESM_1). It can be seen from these results that when the frequency is increased from 200 to 350 kHz, there is a sharp increase in the rate of degradation as well as efficiency of degradation. However, nearly similar degradation efficiency is observed thereafter for 620 kHz and 1 MHz. The degradation efficiency observed at various frequencies can be explained in terms of the number of bubbles, size of the bubbles and the availability of hydroxyl radicals in the interface and bulk. Resonance size is the maximum size of the bubble reached by the cavity before collapse. Resonance size of the cavitating bubble is found to be maximum at lower frequency. The resonant size and frequency are related by the equation (KyuichiYasui 2011).
Analysis of primary intermediate product
Rr ¼ 3KP0 ρωr
For determining the intermediate product, a H class UPLC system with a C18 column, 50×2.1 mm, 1.7 μm connected to a Waters Xevo G2 Q TOF by using electrospray ionisation (ESI) source. The column temperature was set at 35 °C. Water containing formic acid (A) and methanol (B) were used as the eluent. A gradient elution was performed at a flow rate of 0.2 ml min−1. The ionisation was carried out with cone voltage of 20 V.
Where, R r is the resonant radius, K polytropic gas coefficient, P 0 hydrostatic pressure, ρ density of the liquid and ω r is the radiant resonant frequency. The number of bubbles formed and hence the cavitation intensity is expected to be rather low at lower frequency (200 kHz). These bubbles will have large size together with low acoustic cycle leading to higher recombination possibility at interface. This will reduce the chances of having enough • OH in the bulk for larger degree of ATL degradation. When the frequency is increased to 350 kHz, the bubble formation
TOC analysis Total organic carbon (TOC) was determined with a total organic carbon analyser, which utilises a non-dispersive IR (NDIR) detector. The instrument used was a HiPer TOC programmed by ThEuS software. COD analysis
Ion chromatography analysis
2
2
ð2Þ
The ions (both anions and cations) released during the sonolysis was analysed by Dionex ICS-1100 ion chromatography using conductivity detection. For the detection of anions, a mixture of sodium carbonate (2.7×10−3 mol dm−3) and sodium bicarbonate (3×10−4 mol dm−3) in water was used as the mobile phase (flow rate 1 ml min−1). Column used for the detection of anion was Ion Pac AS12A (4 mm×200 mm). For cations, a solution of methane sulfonic acid (20 × 10−3 mol dm−3) in water (1 ml min−1) against an Ion Pac CS12A column was used. Injection volume was 25 μ L.
Results and discussion Effect of frequency on degradation An aqueous solution of ATL was exposed to ultrasound of four varying frequencies at a fixed power (50 W) for 1 h. The
Fig. 1 Effect of frequency on the sonochemical degradation, power= 50 W, [ATL] =10−5 mol dm−3. Inset: first-order kinetics of degradation
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will increase with a size reduction. This will effectively increase the surface-to-volume ratio of bubbles (Navarro et al. 2011; Petrier et al. 1992; Sunartio et al. 2007; Xie et al. 2011). The chances of finding free radicals at interface and in bulk region are higher. More and more ATL will come across hydroxyl radicals enhancing the degradation to a larger extent. Further increase in frequencies to 620 and 1,000 kHz does not result in much variation in degradation since larger frequencies result in fragmentary transient cavitation (Ashokkumar 2011). Effective energy will be lower so as to improve further degradation.
mainly due to the increase of number of active cavitation bubbles and their implosive energy. The increased vibration with power moves the ATL more towards the interface resulting in higher observed degradation. It is already clear from Fig. 1 that at a power of 50 W, 350 kHz resulted in higher degradation. But when the power was increased to 80 W, 620 kHz showed better degradation efficiency (Fig. 2). The number of acoustic cycles increases with frequency whereas the resonance size increases with power. The combined effect of these two processes may be responsible for the observed shift of degradation to 620 kHz (Riez et al. 1985).
Effect of applied power
Effect of solute concentration
The degradation studies were also carried out at higher and lower applied powers (80 and 20 W) for frequencies 350, 620 and 1,000 kHz. The results are presented in Fig. 2. It is found that degradation rate increases with power at all selected frequencies. The increase in degradation along with power is as expected and is in agreement with the results discussed in literature (Xie et al. 2011). As the power increases, energy of collapse increases along with the resonance size. The power lowers the threshold limit of cavitation and thereby induces the formation of bubbles. This will increase the number of cavitating bubbles (Xie et al. 2011). The increase in degradation with respect to power is
The degradation experiments were carried out using varying ATL concentrations in the range 1–50×10−6 mol dm−3 (Fig. 3) by keeping a fixed frequency at 350 kHz and power at 50 W. It is observed that the degradation is more efficient at lower concentration. But when the amount degraded is considered, even the higher solute concentration also gives rise to efficient degradation. This is mainly because of the increase in the probability of reaction of the solute molecule with•OH. This mainly depends on the chemistry of •OH in the interface region. When the concentration of the solute is increased, the collapse rate of bubble is decreased. In addition to this, the competition
Fig. 2 Effect of power on the degradation under the frequency (a) 350 kHz, (b) 620 and (c) 1 MHz. Inset: ln C/C0 vs time plot for each frequency
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Fig. 3 Effect of initial solute concentration on the degradation efficiency; frequency=350 kHz and power 50 W, inset: first-order kinetic plots of degradation with concentrations (black circles) 10−6 mol dm−3, (white circles) 5×10−6 mol dm−3, (black triangles) 10−5 mol dm−3 and (white triangles) 5×10−5 mol dm−3
reaction for the OH radical with the solute will increase, and therefore the degradation will reduce (Isariebel et al. 2009). Effect of pH pH is an important parameter in the water treatment process. Sonolytic degradation was carried at varying pH of the solution (Fig. 4). The effect of pH on degradation was more pronounced at frequency 350 kHz and power 50 W. As the figure displays, highest rate of degradation is observed at acidic pH (k =0.0634 min−1) The pK a value of ATL is reported to be in the range 8.6–9.7 (Sandra Babic et al 2007) and this is attributed for the dissociation of the amine functional group. For most of the aliphatic
Fig. 4 Degradation of atenolol at varying pH; [ATL]=10−5 mol dm−3; Freq=350 kHz, power=50 W. Inset: effect of pH on the first-order kinetics
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amines, the hydrogen bonding between the lone pair of electrons in the nitrogen atom and the proton in the water molecule is the main reason for their solubility in aqueous solution (Zaitseva et al. 2012). But, under acidic pH, the nitrogen atom is protonated and therefore the strength of hydrogen bonding and its solubility got reduced. Thus, the ATL molecule can budge into the highly reactive interface region by decreasing the pH. It may enhance the degradation. The kinetic plot (ln C0/C vs Time) gave a straight line which revealed that the degradation pattern obeyed a pseudo first-order kinetics. This is a demonstration of the involvement of a single reactive species. This means that the degradation mainly occurs through the hydroxyl radical attack. It is known that only about 10 % of the total•OH formed in the bubble interface diffuse into the bulk solution state (Goel et al. 2004). Since the recombination reaction at the interface (which is the main reason for loss of •OH) becomes less efficient under acidic pH (Merouani et al. 2010), it is expected that more •OH is available in the liquid region for the reaction. Effect of Fe2+ Fe2+ is an important additive in sonolysis as it can initiate Fenton reaction with H2O2 which is produced as a result of sonolysis (Ghodbane and Hamdaoui 2009). Figure 5 represents the degradation of ATL by varying the Fe2+ concentration at pH 6. No significant change was observable in the degradation compared to that without the addition of Fe2+. However, when the rate constants were compared, a slight reduction of rate constant was observed in the presence of Fe2+. Initially, the rate constant was 0.0449 min−1, but it decreased to k =0.03636 min−1 and 0.0345 min−1 with the
Fig. 5 The degradation of ATL in the presence of (black circles ) 0 mol dm−3 (white circles) 10−6 mol dm−3 and (black triangles) 5× 10−6 mol dm−3 Fe2+; [ATL] =10−5 mol dm−3; frequency 350 kHz; power 50 W and pH 6. Inset: pseudo first-order kinetics
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addition of 10−6 mol dm−3 and 5×10−6 mol dm−3, respectively. This slight decrease of rate constant with the addition of Fe2+ is likely due to scavenging of •OH by Fe2+ as given in the Eq. 3 (Ghodbane and Hamdaoui 2009). This may reduce the concentration of •OH in the solution state.
Effect of inorganic ions (chloride, sulphate, nitrate and phosphate)
Sodium dodecyl sulphate (SDS) is selected as a model surfactant as it is often found as a pollutant in water bodies. The effect of SDS on the degradation of ATL is also studied. The degradation was found to decrease with the addition of SDS as shown in Fig. 6. The surface active reagents have a tendency to accumulate in the liquid–gas interface of the cavitating bubble (Sunartio et al. 2007). Rate of bubble coalescence is greatly affected by the nature of the surface active solute. SDS belongs to the class of anionic surfactant. Therefore, their accumulation in the interface results in the formation of negatively charged bubbles. This will cause the electrostatic repulsion of bubbles, instead of coalescence and hence an increase in the number of bubbles (Sunartio et al. 2007). Therefore, bubble expansion takes place only through rectified diffusion (Ashokkumar 2011). Consequently, the bubble will not achieve the critical size. In addition, the accumulation of SDS in the interface will result in the reaction of •OH with the SDS instead of ATL, as •OH has a high reactivity with many surfactant (Laughrey et al. 2001). These two factors affect the degradation efficiency as can be seen from the figure.
Sonochemical degradation is highly dependent on the surface tension and the bubble potential. Surface tension decreases the bubble formation rate and cavitation. The increase in bubble potential increases the number of bubbles by reducing the bubble coalescence. Inorganic ions present in a solution can generally alter these parameters and hence their presence is critical in sonochemical degradation. Therefore, the effect of five inorganic anions such as chloride, sulphate, nitrate and phosphate has been investigated. The degradation profile of ATL in the presence of 25 ppm of various inorganic ions at 350 kHz and 50 W is given in Fig. 7. The results show that only sulphate ions inhibit the degradation efficiency and all the other ions give similar degradation pattern. There are reports on the positive and negative effects of anions on the degradation of pollutants (Minero et al. 2008). This behaviour is mainly explained by the interference of secondary radicals in the liquid region as well as the difference in the ionic strength. An enhancement of degradation in the presence of various anions was reported for chlorobenzene, pethyl phenol and phenol-type compounds which are mainly attributed to the salting-out mechanism (Seymour and Gupta 1997). An improvement in the degradation efficiency in the case of Acid blue 40 was realised as the reaction of secondary radicals in the liquid (Minero et al. 2008). The decrease in degradation efficiency in the present case can be explained by considering three factors; they are salting out effect of ions and surface tension. The inorganic ions are generally classified according to their salting out effect (Hofmeister series) (Chen et al. 2007). The salting out effect is generally in the order of SO42− >PO43− >NO3− >Cl−. The
Fig. 6 Change of degradation with the addition (black circles ) 0 mol dm−3, (white circles ) 10−6 mol dm−3, (black triangles ) 5× 10−6 mol dm−3 and (white triangles ) 10−5 mol dm−3 of SDS, at 350 kHz and 50 W. Inset: effect of SDS on the first-order kinetics
Fig. 7 Degradation of ATL in the presence of 25 ppm of various inorganic ions, [ATL] =10−5 mol dm−3; Inset: ln C/C0 vs time plot
2þ
●
Fe þ OH→Fe þ OH− 3þ
ð3Þ
Effect of sodium dodecyl sulphate
Environ Sci Pollut Res (2014) 21:4297–4308 Table 1 Intermediate compounds formed during the sonolysis of ATL
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change of surface tension of the solution by the addition of salt is in the same order. The efficiency of degradation in the present case when 25 ppm salts was added, is in the order no salt>PO43− ∼Cl− ≥NO3− ≥SO42−. According to Hofmeister series, we expect a maximum degradation in the presence of sulphate. But the order is seen reversed in the present case
Scheme 1 Degradation pathway of ATL by sonolysis
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(Chen et al. 2007). Change in surface tension with the addition of any salt is prominent only at higher concentration (Jones and Ray 1941). Thus, the effect of surface tension in this case is having less importance. Therefore, the reduction in the degradation can be explained by taking two possible mechanisms. One is change of surface potential and the other is
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radical scavenging effect (reaction 4) (Minero et al. 2008). The bubble potential will decrease with the addition of sulphate ions. Thus, it will cause more efficient bubble coalescence and hence there will be a reduction in the number of active bubbles. This leads to a reduction in the degradation of ATL in the presence of sulphate (Cheng et al. 2009; Jarvis and Scheiman 1968). Based on the reactions (4), it is obvious that part of the •OH will be scavenged by various ions. The resulting radicals may also react with ATL, but with much lower rate compared to • OH. From this data, it is evident that the present methodology is valid even in the water system containing inorganic matrix. ●
2−
OH þ SO4 →H2 O þ SO4
●−
ð4Þ Fig. 8 TOC reduction during sonolysis at (black circles) pH=5 and (white circles) pH=3.6 at 350 kHz and 80 W. [ATL]=10−5 mol dm−3
Transformation products of ATL during sonolysis One of the important considerations in the case of any advanced oxidation processes is the information of possible intermediate or by-products of oxidation. Such products may greatly influence the total degradation of the compound. In the present case, the intermediate products during sonolysis were analysed from a solution from which nearly 80 % of the parent compound was degraded, using a LC-Q-TOF. Nearly eight products were identified and are presented in Table 1. The identity of these products was confirmed using their MS/MS pattern in the multimode ionisation. The Table 1 shows that two masses are obtained corresponding to the m /z value 283.1659 indicating two monohydroxylated products (related to the addition of 16 mass unit as OH) are formed (I and III). This implies that like other •OH reactions, two possible degradation pathways are present in sonolysis also (Song et al. 2008; Tay et al. 2011; Yang et al. 2010). They are hydroxylation of aromatic ring (path 1, Scheme 1) and hydroxylation at 2-hydroxy3(isopropylamino)propoxy group (path 2, Scheme 1). The more polar compound (retention time 1.61) may be the aromatic hydroxylated compound whereas the compound with retention time 5.16 may be the hydroxylation at 2-hydroxy-3 (isopropylamino)propoxy group. The product at compound (II) with [M+H]+ =299.1596 are obtained which represents the dihydroxylation at the aromatic ring. In addition to II, VI and VII are formed as dihydroxilated product. The product VI may assume to be formed from the hydroxylation of either I or III. The product IV having [M+H]+ =152.0684 denote a radical attack at the ipso position of the ATL. An amino diol (V) ([M+H]+ =134.1189) is assumed to be formed from the cleavage of side chain by •OH. The total ion chromatogram and MS/MS pattern for each product is given as Online Resource ESM_2, ESM_3. The complete degradation pattern is given in the Scheme 1.
Mineralisation during sonolysis The TOC reduction during sonolysis under two different experimental conditions is shown in Fig. 8. Even though sonolysis gives a 100 % transformation, only 62 % TOC reduction was observed during 60 min of irradiation. The product analysis, as explained in the preceding section, showed the formation of various stable intermediate compounds during irradiation rather than complete mineralisation. Under acidic pH, the TOC reduction was found to be very fast. This is mainly due to the accumulation of intermediate compound in the interface region which favours radical attack. Generally, all the beta blockers are found to release ammonium ions during the oxidation reaction (Yang et al. 2010). It is noted that in the present case, NH4+ is continuously released during the sonication time (Fig. 9). It shows that the bound
Fig. 9 The evolution of NH4+ during sonolysis at 350 kHz and 80 W at (black circles) pH 5 (white circles) pH 3.6, [ATL]=10−5 mol dm−3
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Fig. 10 Release of NO3− during sonolysis at (black circles) pH=5 and (white circles) pH=3.6 at 350 kHz and 80 W, [ATL]=10−5 mol dm−3
Fig. 11 Degradation of ATL in pure water (black squares) and in river water (black circles)
nitrogen in the compound is converted into ammonium ions. In addition to NH4+, a constant release of nitrate was also observed (Fig. 10). From the degradation pattern, the NO3− release is likely to be the result of oxidation of V, which is having a free –CONH2 group. The •OH may attack at NH2 group like in the case of amide, and this reaction will release nitrite ion and the nitrate ions (Pelizzetti et al. 2004). The entire degradation mechanism based on these product studies and mineralisation studies is presented in Scheme 1. It is also noted that some amount of organic compounds (intermediate products) are still resistant to complete mineralisation.
The degradation of ATL (10−5 mol dm−3) in river water was monitored after varying time of sonication (350 kHz and 50 W). The decrease in the concentration of ATL is presented in Fig. 11. The degradation of ATL in pure water is also presented in the same figure as monitored by HPLC. At a time scale of 60 min, the degradation in river water was nearly 83 %. This is less than that of pure water (in pure water, it was 93 %). The lower degradation of ATL in river water is likely due to the scavenging of OH radical by the inorganic or organic matrix. In order to look at the mineralization, the corresponding COD reduction was also determined in river water and in pure water as given in Fig. 12. It is seen that the COD value got decreased from 380 to 100 mg/l in river water whereas it went down from 50 to 10 mg/l in pure water. The COD of river water was found to be 350 mg O2 l−1. Thus, it is clear that the
Degradation studies in river water In order to study the efficiency of this methodology in real system, we have carried out degradation studies in river water taken from one of the local rivers in Kerala i.e. Periyar. The various parameters of the water were measured initially. The parameters are listed in the Table 2.
Table 2 Various water quality parameter of river water
pH Conductivity TDS Salinity Turbidity COD Na+ K+ Mg2+ Ca2+
6.28 252 125 mg/l 0.12 psu 1.96 NTU 350 mgO2/l 10.9221 mg/l 0.3171 mg/l 2.8498 mg/l 0.3168 mg/l
Fig. 12 COD reduction of ATL in river water (black circles) and in pure water (black squares) during sonolysis
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river water already contains many organic compounds other than ATL. At the same time, the reduction of COD is well comparable in both the cases. This is an encouraging result. The reasonably comparable COD reduction in river water may be due to several factors related to complex radical reactions. Some of the hydrophobic organic compounds can get into the gas–liquid interface which may result an efficient degradation.
Conclusions AOPs are among the important technologies for the mineralisation of organic water pollutants. This technique is widely experimented in recent years particularly in the degradation of emerging pollutants such as pharmaceutical wastes. Atenolol is a β-blocker drug which is identified as an emerging pollutant. Degradation using ultrasound mainly proceeds through radical reaction as well as pyrolysis. In the present study, radical reaction is identified as the major source of degradation which is clear from the product analysis. The degradation depends on the frequency and power of the ultrasound used. The studies in the presence of various additives which are generally present in contaminated water gave a clear idea about the mechanism of degradation in river water. The observed complete degradation of atenolol and nearly 62 % mineralisation under the present conditions demonstrate the potential utility of the sonochemical method in the removal of atenolol from contaminated water. The information of on the production of various intermediate products would open up further challenges to completely mineralise these compounds using other experimental protocols. One could also think of combining more than one AOPs for its complete mineralisation such as sonophotocatalysis. The present report is a direction towards this aspect. To the best of our knowledge, this is the first report on the sonochemical degradation of atenolol in aqueous medium. Acknowledgments Financial support from KSCSTE, Thiruvananthapuram is gratefully acknowledged. Part of the financial support is also from DST (under FIST and PURSE Prgramme), New Delhi.
References Adewuyi YG (2001) Sonochemistry: environmental science and engineering applications. Ind Eng Chem Res 40:4681–4715 Ashokkumar M (2011) The characterization of acoustic cavitation bubbles—an overview. Ultrason Sonochem 18:864–872 Brillas E, Sirés I, Oturan MA (2009) Electro-Fenton process and related electrochemical technologies based on Fenton’s reaction chemistry. Chem Rev 109:6570–6631 Camacho-Muñoz MD, Santos JL, Aparicio I, Alonso E (2010) Presence of pharmaceutically active compounds in Doñana Park (Spain) main watersheds. J Hazard Mater 177:1159–1162
4307 Chen X, Yang T, Kataoka S, Cremer PS (2007) Specific ion effects on interfacial water structure near macromolecules. J Am Chem Soc 129:12272–12279 Chen Y, Hu C, Hu X, Qu J (2009) Indirect photodegradation of amine drugs in aqueous solution under simulated sunlight. Environ Sci Technol 43:2760–2765 Cheng J, Vecitis CD, Park H, Mader BT, Hoffmann MR (2009) Sonochemical degradation of perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) in groundwater: kinetic effects of matrix inorganics. Environ Sci Technol 44:445–450 Chowdhury P, Viraraghavan T (2009) Sonochemical degradation of chlorinated organic compounds, phenolic compounds and organic dyes—a review. Sci Total Environ 407:2474–2492 De Andrés F, Castañeda G, Ríos Á (2009) Use of toxicity assays for enantiomeric discrimination of pharmaceutical substances. Chirality 21:751–759 Eren Z (2012) Ultrasound as a basic and auxiliary process for dye remediation: a review. J Environ Manag 104:127–141 Fernández C, González-Doncel M, Pro J, Carbonell G, Tarazona JV (2010) Occurrence of pharmaceutically active compounds in surface waters of the Henares-Jarama-Tajo river system (Madrid, Spain) and a potential risk characterization. Sci Total Environ 408:543– 551 Goel M, Hongqiang H, Mujumdar AS, Ray MB (2004) Sonochemical decomposition of volatile and non-volatile organic compounds—a comparative study. Water Res 38:4247–4261 Ghodbane H, Hamdaoui O (2009) Degradation of Acid Blue 25 in aqueous media using 1700 kHz ultrasonic irradiation: ultrasound/ Fe(II) and ultrasound/H2O2 combinations. Ultrason Sonochem 16: 593–598 Isariebel Q-P, Carine J-L, Ulises-Javier J-H, Anne-Marie W, Henri D (2009) Sonolysis of levodopa and paracetamol in aqueous solutions. Ultrason Sonochem 16:610–616 Jarvis NL, Scheiman MA (1968) Surface potentials of aqueous electrolyte solutions. J Phys Chem 72:74–78 Jones G, Ray WA (1941) The surface tension of solutions of electrolytes as a function of the concentration II. J Am Chem Soc 63:288– 294 Kümmerer K (2009a) Antibiotics in the aquatic environment—a review—Part I. Chemosphere 75:417–434 Kümmerer K (2009b) Antibiotics in the aquatic environment—a review—Part II. Chemosphere 75:435–441 KyuichiYasui 2011 (Editor): Fundamentals of acoustic cavitation and sonochemistry theoretical and experimental sonochemistry involving inorganic systems Springer, 1–25 pp Laughrey Z, Bear E, Jones R, Tarr MA (2001) Aqueous sonolytic decomposition of polycyclic aromatic hydrocarbons in the presence of additional dissolved species. Ultrason Sonochem 8:353–357 Liu Q-T, Williams HE (2006) Kinetics and degradation products for direct photolysis of β-blockers in water. Environ Sci Technol 41: 803–810 Merouani S, Hamdaoui O, Saoudi F, Chiha M (2010) Influence of experimental parameters on sonochemistry dosimetries: KI oxidation, Fricke reaction and H2O2 production. J Hazard Mater 178: 1007–1014 Meyer B, Pailler J-Y, Guignard C, Hoffmann L, Krein A (2011) Concentrations of dissolved herbicides and pharmaceuticals in a small river in Luxembourg. Environ Monit Assess 180:127– 146 Minero C, Pellizzari P, Maurino V, Pelizzetti E, Vione D (2008) Enhancement of dye sonochemical degradation by some inorganic anions present in natural waters. Appl Catal B Environ 77:308– 316 Navarro NM, Chave T, Pochon P, Bisel I, Nikitenko SI (2011) Effect of ultrasonic frequency on the mechanism of formic acid sonolysis. J Phys Chem B 115:2024–2029
4308 Pelizzetti E, Calza P, Mariella G, Maurino V, Minero C, Hidaka H (2004) Different photocatalytic fate of amido nitrogen in formamide and urea. Chemical Communications, 1504–1505 Petrier C, Jeunet A, Luche JL, Reverdy G (1992) Unexpected frequency effects on the rate of oxidative processes induced by ultrasound. J Am Chem Soc 114:3148–3150 Riez P, Berdahl D, Christman CL (1985) Free radical generation by ultra sound in aqueous solution and non aqueous solutions. Environ Health Perspect 64:233–252 Sandra B, Horvat AJM, Kastelan-Macan M (2007) Determination of pKa values of active pharmaceutical ingredients. Trends Anal Chem 26: 1045–1061 Seymour JD, Gupta RB (1997) Oxidation of aqueous pollutants using ultrasound: salt-induced enhancement. Ind Eng Chem Res 36:3453– 3457 Song W, de la Cruz AA, Rein K, O’Shea KE (2006) Ultrasonically induced degradation of microcystin-LR and -RR: identification of products, effect of pH, formation and destruction of peroxides. Environ Sci Technol 40:3941–3946 Song W, Cooper WJ, Mezyk SP, Greaves J, Peake BM (2008) Free radical destruction of β-blockers in aqueous solution. Environ Sci Technol 42:1256–1261 Sunartio D, Ashokkumar M, Grieser F (2007) Study of the coalescence of acoustic bubbles as a function of frequency, power, and watersoluble additives. J Am Chem Soc 129:6031–6036
Environ Sci Pollut Res (2014) 21:4297–4308 Sunil Paul MM, Aravind UK, Pramod G, Aravindakumar CT (2013) Oxidative degradation of fensulfothion by hydroxyl radical in aqueous medium. Chemosphere 91:295–301 Tay KS, Rahman NA, Abas MRB (2011) Characterization of atenolol transformation products in ozonation by using rapid resolution highperformance liquid chromatography/quadrupole-time-of-flight mass spectrometry. Microchem J 99:312–326 Wang JL, Xu LJ (2011) Advanced oxidation processes for wastewater treatment: formation of hydroxyl radical and application. Crit Rev Environ Sci Technol 42:251–325 Winter MJ, Lillicrap AD, Caunter JE, Schaffner C, Alder AC, Ramil M, Ternes TA, Giltrow E, Sumpter JP, Hutchinson TH (2008) Defining the chronic impacts of atenolol on embryo-larval development and reproduction in the fathead minnow (Pimephales promelas). Aquat Toxicol 86:361–369 Xie W, Qin Y, Liang D, Song D, He D (2011) Degradation of m-xylene solution using ultrasonic irradiation. Ultrason Sonochem 18:1077– 1081 Yang H, An T, Li G, Song W, Cooper WJ, Luo H, Guo X (2010) Photocatalytic degradation kinetics and mechanism of environmental pharmaceuticals in aqueous suspension of TiO2: a case of βblockers. J Hazard Mater 179:834–839 Zaitseva KV, Varfolomeev MA, Solomonov BN (2012) Hydrogen bonding of aliphatic and aromatic amines in aqueous solution: thermochemistry of solvation. Russ J Gen Chem 8:669–1674
RESEARCH ARTICLE
Sonochemical degradation of a pharmaceutical waste, atenolol, in aqueous medium K. K. Nejumal & P. R. Manoj & Usha K. Aravind & C. T. Aravindakumar
Received: 10 July 2013 / Accepted: 28 October 2013 / Published online: 5 December 2013 # Springer-Verlag Berlin Heidelberg 2013
Abstract Atenolol is a β-blocker drug and an identified emerging pollutant. Advanced oxidation processes (AOPs) utilise the reaction of a highly oxidising species (hydroxyl radicals, •OH) for the mineralisation of emerging pollutants since conventional treatment methodologies generally fail to degrade these compounds. In the present work, degradation of atenolol was carried out using ultrasound with frequencies ranging from 200 kHz to 1 MHz as a source of hydroxyl radical. The degradation was monitored by HPLC, total organic carbon (TOC) and chemical oxygen demand (COD) reduction and ion chromatography (IC). Nearly 90 % of degradation of atenolol was observed with ultrasound having 350 kHz. Both frequency and power of ultrasound affect the efficiency of degradation. Nearly 100 % degradation was obtained at a pH of 4. Presence of various additives such as sodium dodecyl sulphate, chloride, sulphate, nitrate, phosphate and bicarbonate was found to reduce the efficiency of degradation. Although nearly 100 % degradation of atenolol was observed under various experimental conditions, only
Responsible editor: Hongwen Sun Electronic supplementary material The online version of this article (doi:10.1007/s11356-013-2301-x) contains supplementary material, which is available to authorized users. K. K. Nejumal : P. R. Manoj : C. T. Aravindakumar (*) School of Environmental Sciences, Mahatma Gandhi University, Kottayam 686560, Kerala, India e-mail: [email protected] U. K. Aravind Advanced Centre of Environmental Studies and Sustainable Development, Mahatma Gandhi University, Kottayam 686560, Kerala, India C. T. Aravindakumar Inter University Instrumentation Centre, Mahatma Gandhi University, Kottayam 686560, Kerala, India
about 62 % mineralisation (from TOC and COD measurements) was obtained. Nearly eight intermediate products were identified using high-resolution mass spectrometry (LC-QTOF). These products were understood as the results of hydroxyl radical addition to atenolol. The degradation studies were also carried out in river water which also showed a similar degradation profile. A mechanism of degradation and mineralisation is presented. Keywords Emerging pollutants . Sonolysis . Atenolol . Advanced oxidation processes . Hydroxyl radical . Mineralisation
Introduction Pharmaceuticals and personal care products (PPCPs) under the common banner, emerging pollutants, have gained lot of attention in recent years. Emerging pollutants are a class of compounds in the environment such as endocrine disruptors resulting from the degradation of organic compounds or introduction of medicine in the natural environment. The United States Environmental Protection Agency (USEPA) defines “emerging pollutants as new chemicals without regulatory status and which impact on environment and human health are poorly understood” (Kümmerer 2009a, b; Meyer et al. 2011). These compounds are a source of concern because they are used and released in large quantities and results in the widespread distribution into the environment mainly due to their physical and chemical properties. Even the trace level of PPCPs can cause chronic toxicity, endocrine disruption and the development of pathogen resistance. Wastewater treatment plant effluents and secondarily terrestrial run-offs (roofs, pavement, roads and agricultural land) including atmospheric deposition are the major sources of environmentally relevant emerging contaminants. It is not necessary that these
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chemicals need to be persistent in the environment to cause harmful effects, but it creates an environmental catastrophe due to its continuous introduction into the environment. Many of these emerging contaminants, risk assessment and ecotoxicological data are not available at the moment and hence it is difficult to predict which health effects they may cause on humans, terrestrial and aquatic organisms and the ecosystem (De Andrés et al. 2009; Fernández et al. 2010) The general issue associated with the emerging pollutants is that many of these compounds are resistant to degradation under natural conditions. Even if they undergo some kind of transformation, their metabolites also may not be easily degradable. Conventional methodologies for their removal such as flocculation, coagulation, sedimentation, filtration and disinfection are generally employed in drinking water treatment facilities. These techniques generally remove some of the contaminants, but they are inefficient to degrade these chemicals. Microbial degradation is also inefficient in this case as these compounds primarily resist microbes. In this context, oxidation technologies (generally termed as advanced oxidation technology (AOT)) are widely experimented for the destruction of these chemicals. Advanced oxidation processes (AOPs) are still under experimental stage, particularly in the case of emerging pollutants. AOPs are based on the production of a highly reactive oxidant such as hydroxyl radical in solution state and utilise its unselective oxidation of organic chemicals to carbon dioxide and water (Eq. 1). It is a process of mineralisation.
been reported to exist in surface water (Camacho-Muñoz et al. 2010; Meyer et al. 2011; Winter et al. 2008). The same case study highlighted the various aspects of the existence of atenolol and its possible impact in the environment and health. Some of the β-blockers are reported to undergo photodegradation under simulated sunlight and indirect photodegradation in the presence of humic substances and fulvic acid (Chen et al. 2007; Chen et al. 2009; Liu and Williams 2006). Among various AOPs, ultrasound (frequencies between 20 kHz and 1 MHz)-initiated degradation of organic contaminants is an active research area. Sonolysis is considered as a versatile technique which is used to investigate radicalmediated waste water treatment. Perhaps the major advantage of sonolysis is its safety, cleanliness and energy conservation (Chowdhury and Viraraghavan 2009; Eren 2012). One can come across several reports based on the studies of sonolytic degradation of chlorinated organic compounds, phenolic compounds, organic dyes, pesticides, endocrine disrupting compounds, POPs, perfluorinated chemicals, pharmaceuticals and microcystins (Adewuyi 2001; Chowdhury and Viraraghavan 2009; Song et al. 2006). In the present study, an investigation of the sonochemical degradation of atenolol has been carried out in aqueous medium. The effect of ultrasound frequency, applied power and added salts on the degradation efficiency is investigated. The extent of mineralisation in all these cases is also probed.
Hydroxyl radical ð• OHÞ þ PPCPs→ðintermediatesÞ→CO2 þ H2 O
Materials and methods
ð1Þ The production of the hydroxyl radical is the most important part of this technology. Photolysis of hydrogen peroxide, ozone and titanium dioxide, chemical method such as Fenton reaction and sono- and electrochemical methods are some of the techniques come under this category (Brillas et al. 2009; Sunil Paul et al. 2013; Wang and Xu 2011). The crucial step in this kind of processes is to control the intermediate products of oxidation and finally get all those into carbon dioxide and water. Atenolol falls under the category of beta blockers, a class of drugs mainly used in cardiovascular diseases (Fernández et al. 2010). It is a selective receptor antagonist drug and is widely used in many countries. Due to its extensive use as a drug, beta blockers are recently identified as an emerging pollutant and are found to exist in the environment, particularly in sewage effluents and surface waters. Since beta adrenergic receptors were detected in fish and other aquatic animals, there is a high chance that many physiological processes regulated by these receptors in wild animals can be influenced by the presence of beta blockers. Among various beta blockers, atenolol have
Reagents The atenolol of pure grade was purchased from Sigma Aldrich. Aqueous solution was prepared by using CascadaTM Lab Water systems (18.2 MΩ cm−1). Ferrous ammonium sulphate, hydrogen peroxide and other salts used for this study was purchased from Merck India. Sonolysis The reactions were carried out in a sonoreactor having four variable frequencies (202, 351, 623 and 1,003 kHz). The reactor was carried out in a glass reactor; ultrasound was produced using an L3 ELAC Nautik ultrasound generator powered by an Allied Signal R/F generator (T & C power conversion, Model AG 1006). The temperature was maintained at 25±1 °C by using a water circulator. HPLC analysis The degradation of atenolol (ATL) was monitored by highperformance liquid chromatography (Shimadzu prominence
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UFLC, LC 20 AD) connected with a diode array detector (SPD-M20 A). The eluent consists of 0.1 % formic acid in water/methanol (90:10).
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Chemical oxygen demand (COD) of atenolol solutions before and after the treatment was analysed using standard dichromate method. Most type of organic matters is oxidised by a boiling mixture of chromic acid and sulphuric acid. A sample is refluxed in strongly acid solution with a known excess of potassium dichromate (K2Cr2O7). After digestion, the remaining unreduced K2Cr2O7 is titrated with ferrous ammonium sulphate to determine the amount of K2Cr2O7
degradation of atenolol with respect to time of sonication is shown in Fig. 1. There is a fast disappearance initially with time of irradiation. The same trend is followed at all the selected frequencies. The first-order nature of the degradation pattern is also shown in Fig. 1. These calculated values of the first-order rate constant are 0.01649 min−1, 0.04476 min−1, 0.03717 min−1 and 0.0372 min−1 for 200, 350, 620 kHz and 1 MHz, respectively (these are summarised in open resource ESM_1). It can be seen from these results that when the frequency is increased from 200 to 350 kHz, there is a sharp increase in the rate of degradation as well as efficiency of degradation. However, nearly similar degradation efficiency is observed thereafter for 620 kHz and 1 MHz. The degradation efficiency observed at various frequencies can be explained in terms of the number of bubbles, size of the bubbles and the availability of hydroxyl radicals in the interface and bulk. Resonance size is the maximum size of the bubble reached by the cavity before collapse. Resonance size of the cavitating bubble is found to be maximum at lower frequency. The resonant size and frequency are related by the equation (KyuichiYasui 2011).
Analysis of primary intermediate product
Rr ¼ 3KP0 ρωr
For determining the intermediate product, a H class UPLC system with a C18 column, 50×2.1 mm, 1.7 μm connected to a Waters Xevo G2 Q TOF by using electrospray ionisation (ESI) source. The column temperature was set at 35 °C. Water containing formic acid (A) and methanol (B) were used as the eluent. A gradient elution was performed at a flow rate of 0.2 ml min−1. The ionisation was carried out with cone voltage of 20 V.
Where, R r is the resonant radius, K polytropic gas coefficient, P 0 hydrostatic pressure, ρ density of the liquid and ω r is the radiant resonant frequency. The number of bubbles formed and hence the cavitation intensity is expected to be rather low at lower frequency (200 kHz). These bubbles will have large size together with low acoustic cycle leading to higher recombination possibility at interface. This will reduce the chances of having enough • OH in the bulk for larger degree of ATL degradation. When the frequency is increased to 350 kHz, the bubble formation
TOC analysis Total organic carbon (TOC) was determined with a total organic carbon analyser, which utilises a non-dispersive IR (NDIR) detector. The instrument used was a HiPer TOC programmed by ThEuS software. COD analysis
Ion chromatography analysis
2
2
ð2Þ
The ions (both anions and cations) released during the sonolysis was analysed by Dionex ICS-1100 ion chromatography using conductivity detection. For the detection of anions, a mixture of sodium carbonate (2.7×10−3 mol dm−3) and sodium bicarbonate (3×10−4 mol dm−3) in water was used as the mobile phase (flow rate 1 ml min−1). Column used for the detection of anion was Ion Pac AS12A (4 mm×200 mm). For cations, a solution of methane sulfonic acid (20 × 10−3 mol dm−3) in water (1 ml min−1) against an Ion Pac CS12A column was used. Injection volume was 25 μ L.
Results and discussion Effect of frequency on degradation An aqueous solution of ATL was exposed to ultrasound of four varying frequencies at a fixed power (50 W) for 1 h. The
Fig. 1 Effect of frequency on the sonochemical degradation, power= 50 W, [ATL] =10−5 mol dm−3. Inset: first-order kinetics of degradation
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will increase with a size reduction. This will effectively increase the surface-to-volume ratio of bubbles (Navarro et al. 2011; Petrier et al. 1992; Sunartio et al. 2007; Xie et al. 2011). The chances of finding free radicals at interface and in bulk region are higher. More and more ATL will come across hydroxyl radicals enhancing the degradation to a larger extent. Further increase in frequencies to 620 and 1,000 kHz does not result in much variation in degradation since larger frequencies result in fragmentary transient cavitation (Ashokkumar 2011). Effective energy will be lower so as to improve further degradation.
mainly due to the increase of number of active cavitation bubbles and their implosive energy. The increased vibration with power moves the ATL more towards the interface resulting in higher observed degradation. It is already clear from Fig. 1 that at a power of 50 W, 350 kHz resulted in higher degradation. But when the power was increased to 80 W, 620 kHz showed better degradation efficiency (Fig. 2). The number of acoustic cycles increases with frequency whereas the resonance size increases with power. The combined effect of these two processes may be responsible for the observed shift of degradation to 620 kHz (Riez et al. 1985).
Effect of applied power
Effect of solute concentration
The degradation studies were also carried out at higher and lower applied powers (80 and 20 W) for frequencies 350, 620 and 1,000 kHz. The results are presented in Fig. 2. It is found that degradation rate increases with power at all selected frequencies. The increase in degradation along with power is as expected and is in agreement with the results discussed in literature (Xie et al. 2011). As the power increases, energy of collapse increases along with the resonance size. The power lowers the threshold limit of cavitation and thereby induces the formation of bubbles. This will increase the number of cavitating bubbles (Xie et al. 2011). The increase in degradation with respect to power is
The degradation experiments were carried out using varying ATL concentrations in the range 1–50×10−6 mol dm−3 (Fig. 3) by keeping a fixed frequency at 350 kHz and power at 50 W. It is observed that the degradation is more efficient at lower concentration. But when the amount degraded is considered, even the higher solute concentration also gives rise to efficient degradation. This is mainly because of the increase in the probability of reaction of the solute molecule with•OH. This mainly depends on the chemistry of •OH in the interface region. When the concentration of the solute is increased, the collapse rate of bubble is decreased. In addition to this, the competition
Fig. 2 Effect of power on the degradation under the frequency (a) 350 kHz, (b) 620 and (c) 1 MHz. Inset: ln C/C0 vs time plot for each frequency
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Fig. 3 Effect of initial solute concentration on the degradation efficiency; frequency=350 kHz and power 50 W, inset: first-order kinetic plots of degradation with concentrations (black circles) 10−6 mol dm−3, (white circles) 5×10−6 mol dm−3, (black triangles) 10−5 mol dm−3 and (white triangles) 5×10−5 mol dm−3
reaction for the OH radical with the solute will increase, and therefore the degradation will reduce (Isariebel et al. 2009). Effect of pH pH is an important parameter in the water treatment process. Sonolytic degradation was carried at varying pH of the solution (Fig. 4). The effect of pH on degradation was more pronounced at frequency 350 kHz and power 50 W. As the figure displays, highest rate of degradation is observed at acidic pH (k =0.0634 min−1) The pK a value of ATL is reported to be in the range 8.6–9.7 (Sandra Babic et al 2007) and this is attributed for the dissociation of the amine functional group. For most of the aliphatic
Fig. 4 Degradation of atenolol at varying pH; [ATL]=10−5 mol dm−3; Freq=350 kHz, power=50 W. Inset: effect of pH on the first-order kinetics
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amines, the hydrogen bonding between the lone pair of electrons in the nitrogen atom and the proton in the water molecule is the main reason for their solubility in aqueous solution (Zaitseva et al. 2012). But, under acidic pH, the nitrogen atom is protonated and therefore the strength of hydrogen bonding and its solubility got reduced. Thus, the ATL molecule can budge into the highly reactive interface region by decreasing the pH. It may enhance the degradation. The kinetic plot (ln C0/C vs Time) gave a straight line which revealed that the degradation pattern obeyed a pseudo first-order kinetics. This is a demonstration of the involvement of a single reactive species. This means that the degradation mainly occurs through the hydroxyl radical attack. It is known that only about 10 % of the total•OH formed in the bubble interface diffuse into the bulk solution state (Goel et al. 2004). Since the recombination reaction at the interface (which is the main reason for loss of •OH) becomes less efficient under acidic pH (Merouani et al. 2010), it is expected that more •OH is available in the liquid region for the reaction. Effect of Fe2+ Fe2+ is an important additive in sonolysis as it can initiate Fenton reaction with H2O2 which is produced as a result of sonolysis (Ghodbane and Hamdaoui 2009). Figure 5 represents the degradation of ATL by varying the Fe2+ concentration at pH 6. No significant change was observable in the degradation compared to that without the addition of Fe2+. However, when the rate constants were compared, a slight reduction of rate constant was observed in the presence of Fe2+. Initially, the rate constant was 0.0449 min−1, but it decreased to k =0.03636 min−1 and 0.0345 min−1 with the
Fig. 5 The degradation of ATL in the presence of (black circles ) 0 mol dm−3 (white circles) 10−6 mol dm−3 and (black triangles) 5× 10−6 mol dm−3 Fe2+; [ATL] =10−5 mol dm−3; frequency 350 kHz; power 50 W and pH 6. Inset: pseudo first-order kinetics
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addition of 10−6 mol dm−3 and 5×10−6 mol dm−3, respectively. This slight decrease of rate constant with the addition of Fe2+ is likely due to scavenging of •OH by Fe2+ as given in the Eq. 3 (Ghodbane and Hamdaoui 2009). This may reduce the concentration of •OH in the solution state.
Effect of inorganic ions (chloride, sulphate, nitrate and phosphate)
Sodium dodecyl sulphate (SDS) is selected as a model surfactant as it is often found as a pollutant in water bodies. The effect of SDS on the degradation of ATL is also studied. The degradation was found to decrease with the addition of SDS as shown in Fig. 6. The surface active reagents have a tendency to accumulate in the liquid–gas interface of the cavitating bubble (Sunartio et al. 2007). Rate of bubble coalescence is greatly affected by the nature of the surface active solute. SDS belongs to the class of anionic surfactant. Therefore, their accumulation in the interface results in the formation of negatively charged bubbles. This will cause the electrostatic repulsion of bubbles, instead of coalescence and hence an increase in the number of bubbles (Sunartio et al. 2007). Therefore, bubble expansion takes place only through rectified diffusion (Ashokkumar 2011). Consequently, the bubble will not achieve the critical size. In addition, the accumulation of SDS in the interface will result in the reaction of •OH with the SDS instead of ATL, as •OH has a high reactivity with many surfactant (Laughrey et al. 2001). These two factors affect the degradation efficiency as can be seen from the figure.
Sonochemical degradation is highly dependent on the surface tension and the bubble potential. Surface tension decreases the bubble formation rate and cavitation. The increase in bubble potential increases the number of bubbles by reducing the bubble coalescence. Inorganic ions present in a solution can generally alter these parameters and hence their presence is critical in sonochemical degradation. Therefore, the effect of five inorganic anions such as chloride, sulphate, nitrate and phosphate has been investigated. The degradation profile of ATL in the presence of 25 ppm of various inorganic ions at 350 kHz and 50 W is given in Fig. 7. The results show that only sulphate ions inhibit the degradation efficiency and all the other ions give similar degradation pattern. There are reports on the positive and negative effects of anions on the degradation of pollutants (Minero et al. 2008). This behaviour is mainly explained by the interference of secondary radicals in the liquid region as well as the difference in the ionic strength. An enhancement of degradation in the presence of various anions was reported for chlorobenzene, pethyl phenol and phenol-type compounds which are mainly attributed to the salting-out mechanism (Seymour and Gupta 1997). An improvement in the degradation efficiency in the case of Acid blue 40 was realised as the reaction of secondary radicals in the liquid (Minero et al. 2008). The decrease in degradation efficiency in the present case can be explained by considering three factors; they are salting out effect of ions and surface tension. The inorganic ions are generally classified according to their salting out effect (Hofmeister series) (Chen et al. 2007). The salting out effect is generally in the order of SO42− >PO43− >NO3− >Cl−. The
Fig. 6 Change of degradation with the addition (black circles ) 0 mol dm−3, (white circles ) 10−6 mol dm−3, (black triangles ) 5× 10−6 mol dm−3 and (white triangles ) 10−5 mol dm−3 of SDS, at 350 kHz and 50 W. Inset: effect of SDS on the first-order kinetics
Fig. 7 Degradation of ATL in the presence of 25 ppm of various inorganic ions, [ATL] =10−5 mol dm−3; Inset: ln C/C0 vs time plot
2þ
●
Fe þ OH→Fe þ OH− 3þ
ð3Þ
Effect of sodium dodecyl sulphate
Environ Sci Pollut Res (2014) 21:4297–4308 Table 1 Intermediate compounds formed during the sonolysis of ATL
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change of surface tension of the solution by the addition of salt is in the same order. The efficiency of degradation in the present case when 25 ppm salts was added, is in the order no salt>PO43− ∼Cl− ≥NO3− ≥SO42−. According to Hofmeister series, we expect a maximum degradation in the presence of sulphate. But the order is seen reversed in the present case
Scheme 1 Degradation pathway of ATL by sonolysis
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(Chen et al. 2007). Change in surface tension with the addition of any salt is prominent only at higher concentration (Jones and Ray 1941). Thus, the effect of surface tension in this case is having less importance. Therefore, the reduction in the degradation can be explained by taking two possible mechanisms. One is change of surface potential and the other is
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radical scavenging effect (reaction 4) (Minero et al. 2008). The bubble potential will decrease with the addition of sulphate ions. Thus, it will cause more efficient bubble coalescence and hence there will be a reduction in the number of active bubbles. This leads to a reduction in the degradation of ATL in the presence of sulphate (Cheng et al. 2009; Jarvis and Scheiman 1968). Based on the reactions (4), it is obvious that part of the •OH will be scavenged by various ions. The resulting radicals may also react with ATL, but with much lower rate compared to • OH. From this data, it is evident that the present methodology is valid even in the water system containing inorganic matrix. ●
2−
OH þ SO4 →H2 O þ SO4
●−
ð4Þ Fig. 8 TOC reduction during sonolysis at (black circles) pH=5 and (white circles) pH=3.6 at 350 kHz and 80 W. [ATL]=10−5 mol dm−3
Transformation products of ATL during sonolysis One of the important considerations in the case of any advanced oxidation processes is the information of possible intermediate or by-products of oxidation. Such products may greatly influence the total degradation of the compound. In the present case, the intermediate products during sonolysis were analysed from a solution from which nearly 80 % of the parent compound was degraded, using a LC-Q-TOF. Nearly eight products were identified and are presented in Table 1. The identity of these products was confirmed using their MS/MS pattern in the multimode ionisation. The Table 1 shows that two masses are obtained corresponding to the m /z value 283.1659 indicating two monohydroxylated products (related to the addition of 16 mass unit as OH) are formed (I and III). This implies that like other •OH reactions, two possible degradation pathways are present in sonolysis also (Song et al. 2008; Tay et al. 2011; Yang et al. 2010). They are hydroxylation of aromatic ring (path 1, Scheme 1) and hydroxylation at 2-hydroxy3(isopropylamino)propoxy group (path 2, Scheme 1). The more polar compound (retention time 1.61) may be the aromatic hydroxylated compound whereas the compound with retention time 5.16 may be the hydroxylation at 2-hydroxy-3 (isopropylamino)propoxy group. The product at compound (II) with [M+H]+ =299.1596 are obtained which represents the dihydroxylation at the aromatic ring. In addition to II, VI and VII are formed as dihydroxilated product. The product VI may assume to be formed from the hydroxylation of either I or III. The product IV having [M+H]+ =152.0684 denote a radical attack at the ipso position of the ATL. An amino diol (V) ([M+H]+ =134.1189) is assumed to be formed from the cleavage of side chain by •OH. The total ion chromatogram and MS/MS pattern for each product is given as Online Resource ESM_2, ESM_3. The complete degradation pattern is given in the Scheme 1.
Mineralisation during sonolysis The TOC reduction during sonolysis under two different experimental conditions is shown in Fig. 8. Even though sonolysis gives a 100 % transformation, only 62 % TOC reduction was observed during 60 min of irradiation. The product analysis, as explained in the preceding section, showed the formation of various stable intermediate compounds during irradiation rather than complete mineralisation. Under acidic pH, the TOC reduction was found to be very fast. This is mainly due to the accumulation of intermediate compound in the interface region which favours radical attack. Generally, all the beta blockers are found to release ammonium ions during the oxidation reaction (Yang et al. 2010). It is noted that in the present case, NH4+ is continuously released during the sonication time (Fig. 9). It shows that the bound
Fig. 9 The evolution of NH4+ during sonolysis at 350 kHz and 80 W at (black circles) pH 5 (white circles) pH 3.6, [ATL]=10−5 mol dm−3
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Fig. 10 Release of NO3− during sonolysis at (black circles) pH=5 and (white circles) pH=3.6 at 350 kHz and 80 W, [ATL]=10−5 mol dm−3
Fig. 11 Degradation of ATL in pure water (black squares) and in river water (black circles)
nitrogen in the compound is converted into ammonium ions. In addition to NH4+, a constant release of nitrate was also observed (Fig. 10). From the degradation pattern, the NO3− release is likely to be the result of oxidation of V, which is having a free –CONH2 group. The •OH may attack at NH2 group like in the case of amide, and this reaction will release nitrite ion and the nitrate ions (Pelizzetti et al. 2004). The entire degradation mechanism based on these product studies and mineralisation studies is presented in Scheme 1. It is also noted that some amount of organic compounds (intermediate products) are still resistant to complete mineralisation.
The degradation of ATL (10−5 mol dm−3) in river water was monitored after varying time of sonication (350 kHz and 50 W). The decrease in the concentration of ATL is presented in Fig. 11. The degradation of ATL in pure water is also presented in the same figure as monitored by HPLC. At a time scale of 60 min, the degradation in river water was nearly 83 %. This is less than that of pure water (in pure water, it was 93 %). The lower degradation of ATL in river water is likely due to the scavenging of OH radical by the inorganic or organic matrix. In order to look at the mineralization, the corresponding COD reduction was also determined in river water and in pure water as given in Fig. 12. It is seen that the COD value got decreased from 380 to 100 mg/l in river water whereas it went down from 50 to 10 mg/l in pure water. The COD of river water was found to be 350 mg O2 l−1. Thus, it is clear that the
Degradation studies in river water In order to study the efficiency of this methodology in real system, we have carried out degradation studies in river water taken from one of the local rivers in Kerala i.e. Periyar. The various parameters of the water were measured initially. The parameters are listed in the Table 2.
Table 2 Various water quality parameter of river water
pH Conductivity TDS Salinity Turbidity COD Na+ K+ Mg2+ Ca2+
6.28 252 125 mg/l 0.12 psu 1.96 NTU 350 mgO2/l 10.9221 mg/l 0.3171 mg/l 2.8498 mg/l 0.3168 mg/l
Fig. 12 COD reduction of ATL in river water (black circles) and in pure water (black squares) during sonolysis
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river water already contains many organic compounds other than ATL. At the same time, the reduction of COD is well comparable in both the cases. This is an encouraging result. The reasonably comparable COD reduction in river water may be due to several factors related to complex radical reactions. Some of the hydrophobic organic compounds can get into the gas–liquid interface which may result an efficient degradation.
Conclusions AOPs are among the important technologies for the mineralisation of organic water pollutants. This technique is widely experimented in recent years particularly in the degradation of emerging pollutants such as pharmaceutical wastes. Atenolol is a β-blocker drug which is identified as an emerging pollutant. Degradation using ultrasound mainly proceeds through radical reaction as well as pyrolysis. In the present study, radical reaction is identified as the major source of degradation which is clear from the product analysis. The degradation depends on the frequency and power of the ultrasound used. The studies in the presence of various additives which are generally present in contaminated water gave a clear idea about the mechanism of degradation in river water. The observed complete degradation of atenolol and nearly 62 % mineralisation under the present conditions demonstrate the potential utility of the sonochemical method in the removal of atenolol from contaminated water. The information of on the production of various intermediate products would open up further challenges to completely mineralise these compounds using other experimental protocols. One could also think of combining more than one AOPs for its complete mineralisation such as sonophotocatalysis. The present report is a direction towards this aspect. To the best of our knowledge, this is the first report on the sonochemical degradation of atenolol in aqueous medium. Acknowledgments Financial support from KSCSTE, Thiruvananthapuram is gratefully acknowledged. Part of the financial support is also from DST (under FIST and PURSE Prgramme), New Delhi.
References Adewuyi YG (2001) Sonochemistry: environmental science and engineering applications. Ind Eng Chem Res 40:4681–4715 Ashokkumar M (2011) The characterization of acoustic cavitation bubbles—an overview. Ultrason Sonochem 18:864–872 Brillas E, Sirés I, Oturan MA (2009) Electro-Fenton process and related electrochemical technologies based on Fenton’s reaction chemistry. Chem Rev 109:6570–6631 Camacho-Muñoz MD, Santos JL, Aparicio I, Alonso E (2010) Presence of pharmaceutically active compounds in Doñana Park (Spain) main watersheds. J Hazard Mater 177:1159–1162
4307 Chen X, Yang T, Kataoka S, Cremer PS (2007) Specific ion effects on interfacial water structure near macromolecules. J Am Chem Soc 129:12272–12279 Chen Y, Hu C, Hu X, Qu J (2009) Indirect photodegradation of amine drugs in aqueous solution under simulated sunlight. Environ Sci Technol 43:2760–2765 Cheng J, Vecitis CD, Park H, Mader BT, Hoffmann MR (2009) Sonochemical degradation of perfluorooctane sulfonate (PFOS) and perfluorooctanoate (PFOA) in groundwater: kinetic effects of matrix inorganics. Environ Sci Technol 44:445–450 Chowdhury P, Viraraghavan T (2009) Sonochemical degradation of chlorinated organic compounds, phenolic compounds and organic dyes—a review. Sci Total Environ 407:2474–2492 De Andrés F, Castañeda G, Ríos Á (2009) Use of toxicity assays for enantiomeric discrimination of pharmaceutical substances. Chirality 21:751–759 Eren Z (2012) Ultrasound as a basic and auxiliary process for dye remediation: a review. J Environ Manag 104:127–141 Fernández C, González-Doncel M, Pro J, Carbonell G, Tarazona JV (2010) Occurrence of pharmaceutically active compounds in surface waters of the Henares-Jarama-Tajo river system (Madrid, Spain) and a potential risk characterization. Sci Total Environ 408:543– 551 Goel M, Hongqiang H, Mujumdar AS, Ray MB (2004) Sonochemical decomposition of volatile and non-volatile organic compounds—a comparative study. Water Res 38:4247–4261 Ghodbane H, Hamdaoui O (2009) Degradation of Acid Blue 25 in aqueous media using 1700 kHz ultrasonic irradiation: ultrasound/ Fe(II) and ultrasound/H2O2 combinations. Ultrason Sonochem 16: 593–598 Isariebel Q-P, Carine J-L, Ulises-Javier J-H, Anne-Marie W, Henri D (2009) Sonolysis of levodopa and paracetamol in aqueous solutions. Ultrason Sonochem 16:610–616 Jarvis NL, Scheiman MA (1968) Surface potentials of aqueous electrolyte solutions. J Phys Chem 72:74–78 Jones G, Ray WA (1941) The surface tension of solutions of electrolytes as a function of the concentration II. J Am Chem Soc 63:288– 294 Kümmerer K (2009a) Antibiotics in the aquatic environment—a review—Part I. Chemosphere 75:417–434 Kümmerer K (2009b) Antibiotics in the aquatic environment—a review—Part II. Chemosphere 75:435–441 KyuichiYasui 2011 (Editor): Fundamentals of acoustic cavitation and sonochemistry theoretical and experimental sonochemistry involving inorganic systems Springer, 1–25 pp Laughrey Z, Bear E, Jones R, Tarr MA (2001) Aqueous sonolytic decomposition of polycyclic aromatic hydrocarbons in the presence of additional dissolved species. Ultrason Sonochem 8:353–357 Liu Q-T, Williams HE (2006) Kinetics and degradation products for direct photolysis of β-blockers in water. Environ Sci Technol 41: 803–810 Merouani S, Hamdaoui O, Saoudi F, Chiha M (2010) Influence of experimental parameters on sonochemistry dosimetries: KI oxidation, Fricke reaction and H2O2 production. J Hazard Mater 178: 1007–1014 Meyer B, Pailler J-Y, Guignard C, Hoffmann L, Krein A (2011) Concentrations of dissolved herbicides and pharmaceuticals in a small river in Luxembourg. Environ Monit Assess 180:127– 146 Minero C, Pellizzari P, Maurino V, Pelizzetti E, Vione D (2008) Enhancement of dye sonochemical degradation by some inorganic anions present in natural waters. Appl Catal B Environ 77:308– 316 Navarro NM, Chave T, Pochon P, Bisel I, Nikitenko SI (2011) Effect of ultrasonic frequency on the mechanism of formic acid sonolysis. J Phys Chem B 115:2024–2029
4308 Pelizzetti E, Calza P, Mariella G, Maurino V, Minero C, Hidaka H (2004) Different photocatalytic fate of amido nitrogen in formamide and urea. Chemical Communications, 1504–1505 Petrier C, Jeunet A, Luche JL, Reverdy G (1992) Unexpected frequency effects on the rate of oxidative processes induced by ultrasound. J Am Chem Soc 114:3148–3150 Riez P, Berdahl D, Christman CL (1985) Free radical generation by ultra sound in aqueous solution and non aqueous solutions. Environ Health Perspect 64:233–252 Sandra B, Horvat AJM, Kastelan-Macan M (2007) Determination of pKa values of active pharmaceutical ingredients. Trends Anal Chem 26: 1045–1061 Seymour JD, Gupta RB (1997) Oxidation of aqueous pollutants using ultrasound: salt-induced enhancement. Ind Eng Chem Res 36:3453– 3457 Song W, de la Cruz AA, Rein K, O’Shea KE (2006) Ultrasonically induced degradation of microcystin-LR and -RR: identification of products, effect of pH, formation and destruction of peroxides. Environ Sci Technol 40:3941–3946 Song W, Cooper WJ, Mezyk SP, Greaves J, Peake BM (2008) Free radical destruction of β-blockers in aqueous solution. Environ Sci Technol 42:1256–1261 Sunartio D, Ashokkumar M, Grieser F (2007) Study of the coalescence of acoustic bubbles as a function of frequency, power, and watersoluble additives. J Am Chem Soc 129:6031–6036
Environ Sci Pollut Res (2014) 21:4297–4308 Sunil Paul MM, Aravind UK, Pramod G, Aravindakumar CT (2013) Oxidative degradation of fensulfothion by hydroxyl radical in aqueous medium. Chemosphere 91:295–301 Tay KS, Rahman NA, Abas MRB (2011) Characterization of atenolol transformation products in ozonation by using rapid resolution highperformance liquid chromatography/quadrupole-time-of-flight mass spectrometry. Microchem J 99:312–326 Wang JL, Xu LJ (2011) Advanced oxidation processes for wastewater treatment: formation of hydroxyl radical and application. Crit Rev Environ Sci Technol 42:251–325 Winter MJ, Lillicrap AD, Caunter JE, Schaffner C, Alder AC, Ramil M, Ternes TA, Giltrow E, Sumpter JP, Hutchinson TH (2008) Defining the chronic impacts of atenolol on embryo-larval development and reproduction in the fathead minnow (Pimephales promelas). Aquat Toxicol 86:361–369 Xie W, Qin Y, Liang D, Song D, He D (2011) Degradation of m-xylene solution using ultrasonic irradiation. Ultrason Sonochem 18:1077– 1081 Yang H, An T, Li G, Song W, Cooper WJ, Luo H, Guo X (2010) Photocatalytic degradation kinetics and mechanism of environmental pharmaceuticals in aqueous suspension of TiO2: a case of βblockers. J Hazard Mater 179:834–839 Zaitseva KV, Varfolomeev MA, Solomonov BN (2012) Hydrogen bonding of aliphatic and aromatic amines in aqueous solution: thermochemistry of solvation. Russ J Gen Chem 8:669–1674
