Science of the Total Environment 506–507 (2015) 631–643

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Iodinated contrast media electro-degradation: Process performance and degradation pathways Guido Del Moro, Carlo Pastore, Claudio Di Iaconi, Giuseppe Mascolo ⁎ Consiglio Nazionale delle Ricerche, Istituto di Ricerca Sulle Acque, Viale F. De Blasio 5, Bari 70132, Italy

H I G H L I G H T S • • • •

The electrochemical degradation of six iodinated contrast media were investigated. Treatment feasibility as well as reaction by-products and toxicity were investigated. In all the investigated cases, the removal efficiency was higher than 80%. Two main degradation pathways were identified.

a r t i c l e

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Article history: Received 22 July 2014 Received in revised form 25 October 2014 Accepted 31 October 2014 Available online 26 November 2014 Thomas Kevin V Keywords: Iodinated contrast media Electro-degradation By-products Pharmaceutical wastewater Toxicity

a b s t r a c t The electrochemical degradation of six of the most widely used iodinated contrast media was investigated. Batch experiments were performed under constant current conditions using two DSA® electrodes (titanium coated with a proprietary and patented mixed metal oxide solution of precious metals such as iridium, ruthenium, platinum, rhodium and tantalum). The degradation removal never fell below 85% (at a current density of 64 mA/cm2 with a reaction time of 150 min) when perchlorate was used as the supporting electrolyte; however, when sulphate was used, the degradation performance was above 80% (at a current density of 64 mA/cm2 with a reaction time of 150 min) for all of the compounds studied. Three main degradation pathways were identified, namely, the reductive deiodination of the aromatic ring, the reduction of alkyl aromatic amides to simple amides and the de-acylation of Naromatic amides to produce aromatic amines. However, as amidotrizoate is an aromatic carboxylate, this is added via the decarboxylation reaction. The investigation did not reveal toxicity except for the lower current density used, which has shown a modest toxicity, most likely for some reaction intermediates that are not further degraded. In order to obtain total removal of the contrast media, it was necessary to employ a current intensity between 118 and 182 mA/cm2 with energy consumption higher than 370 kWh/m3. Overall, the electrochemical degradation was revealed to be a reliable process for the treatment of iodinated contrast media that can be found in contaminated waters such as hospital wastewater or pharmaceutical waste-contaminated streams. © 2014 Elsevier B.V. All rights reserved.

1. Introduction Iodinated contrast media (ICM) are used for imaging soft tissues, internal organs and blood vessels and can be given to humans at doses of up to 200 g per diagnostic session (Weissbrodt et al., 2009). In medical diagnostics, ICM are applied in high amounts, accounting for over 3.5 × 106 kg per year worldwide (Pérez and Barceló, 2007). As they are composed of a benzene ring and three atoms of iodine, these compounds are the best combination of stability and high x-ray absorption, together with low toxicity, even when decomposed. The elimination of the contrast material from the body involves the reduction of the concentration of the blood with plasma. In extravascular regions, contrast agent returns to the plasma. This phase lasts for a few hours until all of the ⁎ Corresponding author. Tel.: +39 080 5820519; fax: +39 080 5313365. E-mail address: [email protected] (G. Mascolo).

http://dx.doi.org/10.1016/j.scitotenv.2014.10.115 0048-9697/© 2014 Elsevier B.V. All rights reserved.

ICM has been ultra-filtered by the kidney (Verburg et al., 2013). While in some countries (e.g., Japan, China, Greece) wastewater from large hospitals is pre-treated or biologically treated on site, in many other countries, it is directly connected to a municipal sewer and treated at municipal wastewater treatment plants (Pauwels and Verstraete, 2006). Owing to their high persistence, ICM pass through conventional wastewater treatment systems (Kümmerer, 2001; Putschew et al., 2000). Accordingly, ICM are commonly detected at quite elevated concentrations in wastewaters, surface waters, groundwater, bank infiltrate and soil leachates because of their high polarity and persistence (Kormos et al., 2010). Trace amounts of ICM have even been found in drinking waters been produced directly from surface water (Seitz et al., 2006). It is not surprising, therefore, that the concentration of pharmaceuticals in effluents and surface waters are in the μg/L range (Anquandah et al., 2011). Moreover, Forrez et al. (2011) have found secondary effluent ICM concentrations in the range of mg/L. Also, little is

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known about the occurrence and fate of these drug metabolites. In general, it is likely that metabolites will be less biologically active than their parent drugs (Carballa et al., 2004; Drewes et al., 2001; Pérez and Barceló, 2007; Nita and Soltis, 2011). However, metabolites could be cleaved by bacteria that are present in wastewater treatment plants or rivers, hence regenerating the active parent pharmaceutical (Touraud et al., 2011). Busetti et al. (2010) placed particular emphasis on environmental risk due to concentrations of ICM in the water matrix. The presence of pharmaceutical compounds in the environment has led to growing concern because relatively little is known about their impact on the health of humans and ecosystems, especially concerning chronic toxicity from continuous exposure to multiple compounds at doses far below medicinal ones (Jeong et al., 2010). The elimination of ICM from drinking water matrices by biotransformation, direct photolysis, ozonation and advanced oxidation processes (AOPs) has been reported in the literature (Kormos et al., 2011; Hapeshi et al., 2013; Chan et al., 2010; Kovalova et al., 2013; Kwon et al., 2012). However, the performance of ICM degradation by these technologies is still under debate. Highly substituted aromatic compounds such as ICM are reported to be difficult to oxidise and even recalcitrant towards ozonation (Hollender et al., 2009). It was also found that only partial oxidation of ICM can be achieved by ozonation, while TiO2 photocatalysis was effective for quickly removing iopamidol, iopromide and diatrizoate (Murgolo et al., 2015) while it did not lead to complete mineralization of iomeprol (Doll and Frimmel, 2004). In terms of ICM removal, the use of UV/H2O2, UV/S2O2− 8 , UV/TiO2 and O3/H2O2 has been investigated (Chu et al., 2011; Huber et al., 2003). Nevertheless, among the proposed ICM treatment technologies, AOPs have shown great potential to degrade these bio-refractory organics and recalcitrant compounds (Migliorini et al., 2011). Thus, it is essential that alternative treatment technologies are developed that can effectively degrade these compounds. In this context, electrochemical technologies, which have been widely recognised as highly efficient for the treatment of recalcitrant wastewaters, offer an alternative solution to other AOP treatments (Abbas et al., 2009; Bashir et al., 2009; Panizza and Cerisola, 2009; Vlyssides et al., 2003). Typically, electrochemical degradation is superior to other treatment technologies due to its versatility, energy efficiency, absence of sludge production and possibility for automation (Zhao et al., 2010). This work investigates the electrochemical degradation performance of six ICM compounds in bench-scale with synthetic water matrices as well as the occurrence of degradation by-products. To date, the identity and relevance of ICM degradation by-products, which can be formed during the treatment of water matrices, have rarely been investigated (Kormos et al., 2010). To the best of the authors' knowledge, there are few works dealing with the topic at the hand. Zwiener et al. (2009) have studied the iomeprol degradation using a working electrode in nickel foam with an electrolytic cell in which the anodic and cathodic sections are separated by a Nafion septum. Also, Radjenovic et al. (2013) have employed a cell divided by an exchange membrane with a palladium nanoparticles-doped graphite felt working electrode for the study of diatrizoate degradation. While, Mu et al. (2011) have implemented a bioelectrochemical system for the treatment of iopromide with the anode compartment fed with acetate. Furthermore, Eversloh et al. (2014) studied the iopromide in split cells employing a niobium plate coated with a boron-doped diamond film working electrode in order to treat reverse osmosis concentrates from wastewater. In this paper, authors study a wider range of molecules, including iopamidol, iobitridol, iodixanol and iopromide in conventional electrolytic cell. To the best of the authors' knowledge, these molecules have not yet been studied. Molecules have been compared by removal efficiency, energy consumption, degradative pathways and toxicity for each experimental conditions tested. In addition, in this work, we used two DSA® electrodes (titanium coated with a proprietary and patented mixed metal oxide solution of precious metals such Iridium, Ruthenium, Platinum, Rhodium, Tantalum) provided by De Nora Spa (Italy) that have never before been used for applications like that.

Thanks to their specific characteristics, the coating formulation and the coating techniques, the use of DSA® anodes can range from the long established application in the chlor-alkali industry to a wider range of other uses. Moreover, potential future changeovers involve to obtain efficient systems easily extensible to applications at full scale. 2. Materials and methods 2.1. Chemicals Iopamidol (IOP), iobitridol (IOB), iodixanol (IOX), iopromide (IOPR), amidotrizoate (AMI) and iomeprol (IOMPR) were used as model compounds (see Table 1). The compounds were purchased from Sigma Aldrich and reported to be of high purity (at least 99%). The solvents and buffers used in liquid chromatography with mass spectrometric detection, namely, methanol, acetonitrile and acetic acid were of analytical grade or LC-MS grade. 2.2. ICM containing test solutions Stock solutions of iopamidol (IOP), iobitridol (IOB), iodixanol (IOX), iopromide (IOPR), amidotrizoate (AMI) and iomeprol (IOMPR) were singularly prepared using deionised water, sodium sulphate or sodium perchlorate in order to obtain a conductivity of 9 mS/cm and to obtain a final concentration of 1 mg/L or 70 mg/L. The higher contrast medium concentration has enabled analysis of parameters that, for the intrinsic sensitivity of the analytic methods, would not be possible to determine at lower concentrations. The test solutions were treated in a lab-scale electro-degradation setup that is described in the following section. 2.3. Electro-degradation reactor The experimental trials were conducted using a working volume equal to 1 L of a solution containing 1 mg/L or 70 mg/L of the contrast medium to be studied and the supporting electrolyte (sodium perchlorate or sodium sulphate) in order to get a conductivity of 9 mS/cm. Batch electrochemical experiments were performed in a rectangular electrolytic reactor (glass) (see Figure S1) containing 1 L solution. Electrolysis tests were performed under constant current conditions using a direct current (DC) power supply (model SPS-1820, Good Will Instrument Co., Ltd, Taiwan). Two DSA® electrodes (titanium coated with a proprietary and patented mixed metal oxide solution of precious metals such as iridium, ruthenium, platinum, rhodium and tantalum) provided by De Nora S.p.a. (Italy) were arranged parallel to each other and submerged in the synthetic wastewater. Electrode polarity was reversed after each experiment in order to avoid electrode deterioration phenomena. The geometric surface area of each electrode was 110 cm2. A magnetic stirrer (Model SM 26, Stuart Scientific, Stone, Staffordshire ST15 0SA, UK) was used for mixing of the solution. The interelectrode distance was 0.5 cm. The temperature was maintained constant (25 °C) by dipping the reactor in a thermostatic bath. 2.4. Analytical methods Iodide released by the electrochemical reaction of ICM, sulphate and perchlorate was quantified by ion-chromatography (DX-600, Dionex) with a UV or conductivity detector. Molecular iodine was determined using an amperometric titration method (Rice et al., 2012). Iodate ion determination was performed by UPLC-MS analysis (Applied Biosystems/ MSD Sciex). DOC was measured by a TOC-V carbon analyser (Shimadzu). Conductivity and pH were measured on-line using selective probes. The specific energy consumption (Ec, in kWh m−3) was calculated as follows: Ec = (Ucell × I × t) / (V × 3600), where Ucell is the average cell voltage (V), I is the current (A), t is the electrolysis time (s) and V is the volume of the treated wastewater (L) (Migliorini et al., 2011; Panizza et al., 2010). For

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Table 1 Investigated iodinated contrast media. Chemical

CAS number

Formula

Iopamidol

62883-00-5

C17H22I3N3O8 MW:777.09

Iobitridol

136949-58-1

C20H28I3N3O9 MW: 835.15

Iodixanol

92339-11-2

C35H44I6N6O15 MW: 1550.18

Iopromide

73334-07-3

C18H24I3N3O8 MW:791.11

Amidotrizoate

Iomeprol

737-31-5

78649-41-9

C11H8I3N2NaO4 MW: 635.9

C17H22I3N3O8 MW:777.09

the Vibrio fischeri test, a commercial assay marketed as Biofix®Lumi-10 was employed using a freeze-dried specially selected strain of the marine bacterium (NRRL number B-11177). Toxicity was evaluated in undiluted samples. The decrease in light emission of the bacteria after a contact period of 30 min was measured and compared with a toxicantfree control (2% NaCl solution). The temperature was kept at 15 °C by a thermo block, and sample salinity was altered to 2% after adjusting the sample pH between 6 and 8. EC50 values, expressed as dilution percentage, were determined using a Model 500 analyser by SDI (Newark, Delaware, USA) employing the ISO 11348-3 method (ISO, 2008). The residual concentration of the studied compounds at various reaction times and by-product determination was performed using UPLC-MS analysis, with an Acquity UPLC chromatographic system equipped with an auto-sampler and a photodiode array detector (Waters), interfaced to an API 5000 mass spectrometer (Applied Biosystems/MSD Sciex) through an electrospray interface. Analyses were carried out using an Acquity BEH Phenyl 2.1 × 150 mm, 1.7 μm chromatographic column, with a flow of 0.15 mL/min and a methanol/ water gradient elution. The solvents A1 (95% H2O, 5% methanol, 1.5 mM ammonium acetate) and B1 (1.5 mM ammonium acetate in methanol) were mixed according to Table S1. The samples were introduced into the system by means of an autosampler equipped with a six-way motorised valve (Rheodyne) and a 5 μL loop. Table S2 shows the experimental conditions used for the acquisition in full scan mode

analysis. Table S3 shows the experimental conditions used for the multiple reaction monitoring (MRM) mode analysis. Table S4 lists the experimental conditions detailed for the product ion scan (PIS) analysis. 2.5. Data analysis The UPLC-MS/MS analyses data are returned by the instrumentation as raw files that are processed by Analyst software (AB SCIEX) version 1.5.1. These raw files were converted into “.txt” files using the MS Processor software included in the ACDLABS suite version 12.0. Subsequently, the resulting files were processed by the software EnviMass (Target Screening workflow) version 1.2 developed by Eawag (Switzerland) in order to subtract the control runs (blanks) from each sample run. At the end of this processing, and for each contrast medium studied, a mass list of compounds potentially generated during the reactions was obtained. 3. Results and discussion 3.1. Electro-degradation performance using naturally occurring concentrations In Figs. 1 and 2, the decays of the six studied organic pollutants using perchlorate (Fig. 1) or sulphate (Fig. 2) as a supporting electrolyte are

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Fig. 1. Contrast media (1 mg/L initial concentration) degradation using perchlorate as a supporting electrolyte: (a) IOP, (b) IOB, (c) IOX, (d) IOPR, (e) AMI and (f) IOMPR.

represented. When perchlorate was used, also for the lower current intensity (64 mA/cm2), the degradation effectiveness was never below 85% (percentage obtained for the iomeprol and shown in Fig. 1f) with a reaction time of 150 min. When sulphate was used as a supporting electrolyte, the degradation performance decreased slightly. However, it always remained above 80% for all of the compounds studied (at a current density of 64 mA/cm2 with a reaction time of 150 min). It is worth noting that the worst removal yields were obtained for iopromide, and this result was more marked when the current intensity was lowered (Fig. 2d). Using the same parameters (electrodes composition, current and voltage), we then checked whether the effect of the supporting electrolyte was negligible or not. In fact, it is known that many oxidants are produced on the electrode surface. In the case of sulphate, the production of peroxosulphates on the anode is well known (Souza et al., 2014). Also, in electrochemical literature, ClO− 4 is often considered to be very stable. Therefore, perchlorate-containing solutions are widely used as supporting electrolytes in electrochemical studies with various electrodes (Ujvari and Lang, 2011). Preliminary reaction experiments performed at various concentration of sulphate or perchlorate showed that the support electrolyte does not play an active role in the degradation reactions (data not shown). Taking everything into account, the experiments conducted using the aforementioned support electrolytes led

to no significant differences being detected in terms of degradation yields. In fact, degradation of organics was found to take place in two steps: diffusion of compounds from the bulk solution to the electrode surface and reaction on the surface. Thus, the efficiency of the electrochemical process depends on the relationship between the mass transfer of the substrate and electron transfer at the electrode surface. The rate of electron transfer is determined by the electrode activity and current density (Anglada et al., 2009). In this condition, contrast media removal, due to mass-transport limitation, followed an exponential trend. This is in keeping with that which was found by Panizza et al. (2001). Values of the degradation rate constants k and of the regression coefficients R2 for all of the studied compounds and for both supporting electrolytes are listed in Table 2. First-order rate constants were evaluated from linear plots of log[C/Co] against time. The regression coefficient (R2) was always higher than 0.981 for all of the studied cases. The value of k increased with the current intensity, which confirmed that this factor considerably influences the rate of reaction. It should be highlighted that an increase in current density does not necessarily result in an increase in reaction efficiency or degradation rate and that the effect of current density on the treatment efficiency for a given electrode material depends on the characteristics of the effluent to be treated (Anglada et al., 2009). In general, during these experiments, when

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Fig. 2. Contrast media (1 mg/L initial concentration) degradation using sulphate as a supporting electrolyte: (a) IOP, (b) IOB, (c) IOX, (d) IOPR, (e) AMI and (f) IOMPR.

using perchlorate, it was found that the k was slightly bigger than with the use of sulphate at equal current densities. In any case, these values are higher than those found by Eversloh et al. (2014) for iopromide that range between 0.57 × 10−4 s−1 and 1.62 × 10−4 s−1 that however

worked at lower currents equal to 20 mA. Also, Radjenovic et al. (2013) found for diatrizoate lower values of k comprised between 0.12 and 0.41 h− 1 working potentiostatically at cathode potentials ranging from -1.1 to -1.7 V vs SHE.

Table 2 Kinetics of contrast media (1 mg/L initial concentration) degradation using perchlorate (a) or sulphate (b) as a supporting electrolyte. Values are expressed as the mean of two experiments. Current density (mA/cm2)

IOP

IOB

IOX

IOPR

AMI

IOMPR

k (min−1)

R2

k (min−1)

R2

k (min−1)

R2

k (min−1)

R2

k (min−1)

R2

k (min−1)

R2

a) 64 91 118 145 182

0.019 0.020 0.023 0.028 0.042

0.991 0.995 0.991 0.994 0.997

0.018 0.021 0.029 0.039 0.050

0.996 0.996 0.995 0.984 0.985

0.018 0.023 0.025 0.033 0.041

0.992 0.984 0.990 0.994 0.987

0.021 0.028 0.038 0.043 0.048

0.981 0.985 0.984 0.985 0.992

0.019 0.024 0.027 0.033 0.046

0.999 0.996 0.999 0.997 0.987

0.015 0.022 0.029 0.032 0.043

0.996 0.997 0.991 0.981 0.997

b) 64 91 118 145 182

0.015 0.020 0.021 0.041 0.045

0.989 0.998 0.996 0.995 0.989

0.016 0.021 0.026 0.027 0.039

0.989 0.995 0.983 0.997 0.996

0.016 0.019 0.025 0.031 0.044

0.997 0.991 0.989 0.989 0.994

0.012 0.014 0.015 0.020 0.028

0.998 0.991 0.993 0.990 0.989

0.013 0.016 0.019 0.027 0.033

0.991 0.992 0.991 0.997 0.997

0.014 0.018 0.023 0.029 0.038

0.988 0.992 0.985 0.984 0.986

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In order to better clarify some aspects of the efficiency of the electrodegradation process, Figs. 3 and 4 show the evolution of removal percentage for the six studied compounds, with the specific electrical charge passed during the electrolysis. It should be mentioned that these figures provide a measurement of the efficiency of the electrodegradation process (Anglada et al., 2009). The trends represented in Figs. 3 and 4 outline similar behaviours between the compounds and between the two supporting electrolytes used. For all organics and for both supporting electrolytes, a removal percentage greater than or equal to 80% was obtained using a specific electrical charge “Q” between 1000 and 1200 A min/L. For all of the cases studied, it is possible to obtain complete removal by current intensities equal to or greater than 118 mA/cm2 using a specific electric charge of about 2000 A min/L. The removal percentage increased with current density, but there were no significant differences between the amount of electric charge consumed at different current density values (Figs. 3 and 4). This indicates that the faradic yield of the degradation process does not depend on current density. This behaviour is characteristic of mass-transfercontrolled processes and is similar to results obtained by other authors

3.2. Set-up energy consumption using naturally occurring concentrations Figs. 5 and 6 show the specific energy consumption versus the removal percentage of the six studied compounds for the two supporting electrolytes employed. Energy consumption is clearly proportional to the increase in the reaction time, with the current density remaining constant for each experiment. The only variable was the voltage necessary to maintain the constant current. The energy consumptions for the six studied organics and for the final reaction time, almost equal for both supporting electrolytes, were 135.9 ± 0.9, 239.5 ± 2.9, 367.6 ± 2.6, 509.8 ± 8.3 and 705.7 ± 10.0 kWh/m3, respectively, for the experiments conducted at 64, 91, 118, 145 and 182 mA/cm2. Figs. 5 and 6 show that at low removal percentage (up to a range between 20% and 40%), the trend of specific energy consumption shows no apparent differences for all the employed current densities but the scale factor of the graph does not make evident that there are variations ranging on

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for the degradation of different pollutants (Panizza and Cerisola, 2009); (Bedoui et al., 2009).

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Fig. 3. Removal percentage (1 mg/L initial concentration) versus the specific electrical charge passed, using perchlorate as a supporting electrolyte, for (a) IOP, (b) IOB, (c) IOX, (d) IOPR, (e) AMI and (f) IOMPR.

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Fig. 4. Removal percentage (1 mg/L initial concentration) versus the specific electrical charge passed, using sulphate as a supporting electrolyte, for (a) IOP, (b) IOB, (c) IOX, (d) IOPR, (e) AMI and (f) IOMPR.

average between 10 and 60 kWh/m3. This behaviour is similar for all the molecules and for the two supporting electrolytes studied. With increasing removal percentages, consumption was shown to diverge depending on the current intensity applied. For all of the considered cases and already at 80% removal, the energy consumption diverged between the lowest and the highest current density, wavering at around 100 kWh/m3. Obviously, at greater current densities, lower reaction times can be used to achieve the same removal percentage. If, however, the aim is to achieve total compound removal, it is necessary to employ current intensity between 118 and 182 mA/cm2 with energy consumption higher than 370 kWh/m3. In fact, with low current density, a high current efficiency and a low energy consumption can be achieved even if organic mineralisation required a longer electrolysis time (Panizza et al., 2008). This can also be deduced by studying Figures S4 and S5, which report the DOC removal of experiments conducted at higher contrast media concentrations, with the aim of identifying byproducts. It can be seen that, at 64 and 145 mA/cm2, complete mineralisation was never obtained. In conclusion, we need to take into account not only energy costs but also the actual mineralisation or

degradation of compounds in order to perform the effective treatment of wastewater contaminated by ICM. In addition, the cost as function of the possible formation of potentially harmful by-products has to be considered. 3.3. Toxicity evaluations of treatments involving naturally occurring concentrations As ICM are given to people, they are very well controlled by drug inspection agencies. In fact, ICM besides potential kidney intoxication problems should not present any environmental toxicity. However, although generally reversible, contrast media toxicity often induces nephropathy (Quiros et al., 2013). It is quite well known that ICM give rise to direct cytotoxicity, resulting in apoptosis/necrosis of renal tubular cells and vascular endothelial cells, in both in vitro and in vivo studies, with ionic ICM showing the highest toxicity (Scoditti et al., 2013). However, there is lack of information regarding the environmental toxicity, while there is no information on the toxicological consequences of treatments for the removal of ICM from wastewater. Steger-Hartmann

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specific energy consumption (kWh/m3)

specific energy consumption (kWh/m3)

c

40

2

182 mA/cm 182 mA/cm2

400 300 200 100 0

9191 mA/cm2 mA/cm2 118 mA/cm2 118 mA/cm2

600

145 mA/cm2 145 mA/cm2 182 mA/cm2 182 mA/cm2

500 400 300 200 100 0

0

20

40

60

80

100

removal (%)

0

20

40

60

removal (%)

Fig. 5. Evolution of specific energy consumption against the removal percentage (for the experiment with 1 mg/L of ICM initial concentration), using perchlorate as a supporting electrolyte, for (a) IOP, (b) IOB, (c) IOX, (d) IOPR, (e) AMI and (f) IOMPR.

et al. (1999) reported that ICM may significantly contribute to the presence of absorbable organic halogens (AOX) in municipal wastewaters. Also, contrast media were identified as a major source of increased AOX concentrations in hospital sewage water. Therefore, it is useful to have information on the toxicity of wastewater, contaminated with ICM, following any treatments. In this work, the inhibition of luminescence in the V. fischeri test was used. The inhibition of bioluminescence is used as an end point to determine the effects of a chemical substance on an effluent. The synthetic wastewater which was used prior to

electro-degradation treatment (0 min of reaction time) caused no relevant inhibition of luminescence. This is not surprising as any drug that is administered to people cannot present evident toxicity. This lack of toxicity is also in agreement with its characteristics as a well-tolerated contrast medium that has been optimised for pharmacodynamics neutrality and metabolic stability (Steger-Hartmann et al., 1999). This finding is not present in Fig. 7, however, where the values of EC50 are only shown for the experimental cases that showed even a minimal toxicity. As can be seen from Fig. 7, the toxicity, as EC50, is around 50% for IOB

G. Del Moro et al. / Science of the Total Environment 506–507 (2015) 631–643

a

639

b 800

specific energy consumption (kWh/m3)

specific energy consumption (kWh/m3)

64 64mA/cm2 mA/cm 2

700

700 64 mA/cm2 64 mA/cm2

600

91 mA/cm2 91 mA/cm2

118 mA/cm2 118 mA/cm2

500

145 mA/cm2 145 mA/cm2 182 mA/cm2 182 mA/cm2

400 300 200 100 0

91 91mA/cm2 mA/cm 2 118 mA/cm2 118 mA/cm 2

600

145 mA/cm2 145 mA/cm 2

182 mA/cm 2 182 mA/cm2

500 400 300 200 100 0

0

20

40

60

80

0

100

20

removal (%)

60

80

100

d

800

800 6464 mA/cm2 mA/cm2

700

specific energy consumption (kWh/m3)

specific energy consumption (kWh/m3)

c

40

removal (%)

9191 mA/cm2 mA/cm2

600

118 mA/cm2 118 mA/cm2

500

182 mA/cm2 182 mA/cm2

145 mA/cm2 145 mA/cm2

400 300 200 100 0

64 mA/cm2 64 mA/cm2

700

91 mA/cm2 91 mA/cm2

600

118 mA/cm2 118 mA/cm2

145 mA/cm2 145 mA/cm2

500

182 mA/cm2 182 mA/cm2

400 300 200 100 0

0

20

40

60

80

0

100

20

removal (%)

40

60

80

100

80

100

removal (%)

e

f 700 91 mA/cm2 91 mA/cm2

specific energy consumption (kWh/m3)

specific energy consumption (kWh/m3)

64 mA/cm2 64 mA/cm2

600

118 mA/cm2 118 mA/cm2 145 mA/cm2 145 mA/cm2

500

182 mA/cm2 182 mA/cm2

400 300 200 100

800 64 mA/cm2 64 mA/cm2

700

91 mA/cm2 91 mA/cm2 118 mA/cm2 118 mA/cm2

600

145 mA/cm2 145 mA/cm2

182 mA/cm2 182 mA/cm2

500 400 300 200 100

0

0 0

20

40

60

80

100

removal (%)

0

20

40

60

removal (%)

Fig. 6. Evolution of specific energy consumption against the removal percentage (for the experiment with 1 mg/L of ICM initial concentration), using sulphate as a supporting electrolyte, for (a) IOP, (b) IOB, (c) IOX, (d) IOPR, (e) AMI and (f) IOMPR.

using both sulphate and perchlorate, for AMI with sulphate and for IOPR with sulphate. In contrast, in the other cases, this value is less than 30% with IOX-S, IOX-P, IOPR-P and IOMPR-S showing the strongest toxicity equal to 4%, 3%, 8% and 11% respectively. The data showed that toxicity had a mutable pattern after the electro-degradation process increasing from a non-toxic to a toxic level throughout 150 min at the lower employed current intensity (64 mA/cm2). As for IOP, IOB and IOX, there seems to be no substantial difference in toxicity between the

two used electrolytes, this is not the case for the other studied contrast media. For IOPR and AMI, it can be seen less toxicity for samples with sulphate while the result is the opposite for IOMPR where samples with perchlorate is less toxic. Generally, the exact source of toxicity in the by-product mixture is difficult to determine conclusively. As described in the section devoted to degradation mechanisms, it is assumed that the used different supporting electrolytes do not affect in a different way on the contrast media degradation pathways since it suggests that

640

G. Del Moro et al. / Science of the Total Environment 506–507 (2015) 631–643

70 60

EC50 (%)

50 40 30 20

IOMPR–P

IOMPR–S

AMI–P

AMI–S

IOPR–P

IOPR–S

IOX–P

IOX–S

IOB–P

IOB–S

IOP–P

0

IOP–S

10

Fig. 7. Graph of only those samples showing toxicity corresponding to the trials at 64 mA/cm2 for 150 min reaction time (S and P stand for sulphate and perchlorate, respectively). Error bars represent standard error (n = 2).

direct electro-reduction on the cathode surface is the main method of contrast media electro-degradation. Fig. 7 shows that, likely, after long-term oxidation (150 min), some chemical degradation intermediates (as described in Section 3.4) with slight toxicity, such as compounds that have had de-iodination of the aromatic ring or compounds that have gone through the N-aromatic amides de-acylation and/or alkyl aromatic amides reduction, are formed and remain in solution for a variable time. The toxicity variability is attributed to the concentration variations of the intermediates. Many by-products generated were most likely not further transformed (as can be deduced by the DOC trends shown in figures S4 and S5 that highlight the non-complete mineralisation at 64 mA/cm2); therefore, by-products and their associated toxicity remained in solution at the end of the reaction. Also, it should be pointed out that synergistic effects among the by-products might also strengthen the toxicity of the solution (Scheurell et al., 2009). The increased toxicity after treatment was previously explained by the changed substituent group of the parent compound (Yuan et al., 2011). The changed substituent group might reduce steric resistance due to a reduction in size, resulting in membrane greater ease of cell penetration than the parent compound (Lu et al., 2002) and consequently leading to increased toxicity. It remains to understand the possible role, if it exists, played by the two different supporting electrolytes employed. Anyway, further work is still needed to elucidate the exact toxicity source. According to the toxicity data obtained, it can be assumed that at the lower current density (the only one to which it was detected toxicity), reaction intermediates have been formed that either remain in the studied medium (wastewater) for longer periods or that are not further degraded, thereby leading to an overall increase in toxicity. This suggests that an optimal current density dosage would be required to completely remove temporary by-products. However, energy consumption should still be assessed. 3.4. Degradation mechanisms A by-products study of the compounds generated during the electrolysis process was carried out that allowed the degradation mechanism of ICM electro-reduction to be determined. Two basic types of de-functionalization have been identified on IOP, IOPR, IOMPR, IOB and IOX: (i) the reductive de-iodination of the aromatic ring and (ii) the reduction of alkyl aromatic amides to simple amides (Fig. 8 shows these two reactions for iopamidol). Reductive de-iodination

was found to be the fastest reaction, since that already after 5 min of electrolysis, the signals attributable to the respective products of first de-iodination were easily detected for all studied ICMs (see Figure S6 for the specific case of IOP). On the other hand, results reported in Table 2 show that in fixing a current intensity, the values of kinetic constants of degradation k determined for the six iodinated compounds (AMI included) are quite similar, even when very different molecules have been compared. This experimental data could be interpreted only by considering a very “similar” reaction mechanism for degradation processes for all the tested molecules. Reasonably, the aromatic ring may play a key role in capturing an electron from the cathode, and an iodide anion was expelled away from the molecule with a global rate that depended essentially from the nature of the substituents of the aromatic ring, which practically are always the same for all the tested ICMs. At the end, being the fastest reaction, it resulted to pivot the kinetic of degradation. On the contrary, process ii was interestingly found dependent by the nature of the alkyl group bound to the nitrogen of the aromatic amide. In detail, only for the iopamidol case, for which the above-mentioned alkyl group is –CH(CH2OH)2, the de-alkylation process, with the formation of the simple aromatic amide, was easily ascertained already after 5 min of electrolysis. Figure S7 clearly shows the extracted chromatogram of the 702 m/z ion correspondent to the simple amide obtained by the de-alkylation of starting IOP. In the other 4 above-mentioned molecules, for which the respective alkyl group bounded to the N-amide is the isomer –CH2CH(OH)CH2OH, the formation of the simple aromatic amide was obtained more slowly (only after 60 min). More presumably, supposing a preliminary formation of a radical carbon that then evolves into a loss of the branched alkyl group (see Supporting Information), the different hindrance effect on the carbon bounded to the N of the aromatic amide, and the nature of the same carbon (secondary for the IOP and primary for all the other similar ICMs), could justify these different kinetic evidences. The different by-products obtained during electrolysis was identified and determined, through LC/MS-scan analysis carried out on samples treated with 145 mA/cm2 for 150 min (see Figure S8-S20, always for the IOP case). However, the evolution of the so identified intermediates during the electrolytic time was followed by extracting their correspondent ions from the LC-scan analysis conducted on the several solutions taken at different treatment time. The different degradation

G. Del Moro et al. / Science of the Total Environment 506–507 (2015) 631–643

641

aromatic ring reductive de-iodination alkyl aromatic amides reduction

Fig. 8. Proposed main degradation mechanisms for the iopamidol electro-chemical treatment. (For interpretation of the references to colour in this figure, the reader is referred to the web version of this article.)

pathways, specific for each compound of studied ICMs, are shown in Figures S21 to S33. For the specific case of the AMI (see Figure S32), being the only aromatic carboxylate, in addition to the de-iodination reaction, a decarboxylation reaction was verified in its general degradation pathway. And always for this compound, the reaction of de-acylation of N- aromatic amides to produce aromatic amines was further detected, only after that no other groups were remained on the molecule, as effect of the advancement of reactions of de-iodination and decarboxylation. Considering that all ICMs are aromatic amide and that only for AMI were determined appreciable development of such a de-acetylation process, most probably, this different reactivity could be attributable to the simpler nature of the acylic group of the AMI (-C(O)-CH3) with respect to those present in the other ICMs (-C(O)CH2OCH3). From the conclusions drawn so far on the possible mechanisms, it is plausible that direct electro-reduction on the cathode surface is the main method of contrast media degradation as suggested by Eversloh et al. (2014) even though they found that in additions to iopromide, de-iodination also a conversion of amide carbonyls to OH-groups and reductive side-chain cleavage occurs at the cathode. In order to further confirm the proposed mechanisms, measurements of the different forms of iodine present in the solution were carried out. In Tables 3 and 4, the concentrations for the different iodine species which are assumed to be in the reaction medium are represented. The values refer to the final reaction time at the current density of 64 mA/cm2 (Table 3) and at the current density of 145 mA/cm2 (Table 4). The various iodine forms were measured in samples from the experiments used to identify the by-products, which were conducted using high ICM concentrations (70 mg/L). Molecular iodine, iodide and iodate were measured as the more stable forms [39]. In fact, many different iodine species can exist in aqueous solutions, since iodine can form compounds in all oxidation states from –1 to + 7. The hydrolysis of

molecular iodine to form hypoiodous acid (HOI), following rapid equilibrium including intermediate species as shown in Eqs. (1), (2) and (3). ‐

þ

3I2 þ 3H2 O↔ 3I þ 3HOI þ 3H ‐

ð1Þ

þ



3HOI ↔ 2I þ IO3 þ 3H ‐



ð2Þ þ

3I2 þ 3H2 O ↔ 5I þ IO3 þ 6H ‐

I2 þ I ↔ I3

ð3Þ



ð4Þ

Table 3 Concentrations of the iodine species that can be formed during electro-degradation. The values refer to experiments at higher starting concentrations (70 mg/L) and at final reaction times (150 min) with a current density of 64 mA/cm2 for perchlorate (a) and sulphate (b). The reported values are the average of two replicates and are expressed in mg/L. IOP

IOB

IOX

IOPR

AMI

IOMPR

34.3

31.9

34.4

33.7

41.9

34.3

a) I− I2 IO− 3 Parental residual as I−

2.2 4.3 15.7 9.4

1.1 2.1 12.1 15.9

0.73 2.2 12.8 9.1

1.2 2.6 13.9 14.1

2.5 4.4 16.2 15.1

0.93 5.6 15.7 6.3

b) I− I2 IO− 3 Parental residual as I−

1.47 7.1 9.4 14.3

2.23 5.5 6.1 16.4

0.39 11.1 6.9 10.8

1.73 4.1 8.7 16.9

2.9 7.9 10.8 19.4

1.58 12.7 10.1 8.2

Parental as I



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G. Del Moro et al. / Science of the Total Environment 506–507 (2015) 631–643

Table 4 Concentrations of the iodine species that can be formed during electro-degradation. The values refer to experiments at higher starting concentrations (70 mg/L) and at final reaction times (150 min) with a current density of 145 mA/cm2 for perchlorate (a) and sulphate (b). The reported values are an average of two replicates and are expressed in mg/L.

Parental as I



IOP

IOB

IOX

IOPR

AMI

IOMPR

34.3

31.9

34.4

33.7

41.9

34.3

a) I− I2 IO− 3 Parental residual as I−

0.52 9.8 15.9 0.7

1.2 12.5 14.6 2.4

2.4 12.1 14.9 0.9

0.27 11.8 15.8 4.8

3.1 6.4 17.7 2.6

2.4 12.8 17.3 0.2

b) I− I2 IO3− Parental residual as I−

0.23 9.9 9.7 1.4

1.9 9.9 6.9 2.7

0.92 12.3 8.1 1.3

4.4 7.3 10.1 6.6

3.3 18.1 12.4 4.6

3.4 18.9 10.8 0.1

iodinated compounds. The processes economy will be improved through researches into the electrode materials and on the optimisation of cell parameters as well as by the market economic laws. Acknowledgements The authors are grateful to De Nora S.p.A. for providing the electrodes used in this experimentation and by the Italian Ministry of Foreign Affairs, in the framework of the bilateral cooperation Italy-Quebec (grant no. PGR00357). Appendix A. Supplementary data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.10.115. References

Also, in rapid equilibrium with iodide, molecular iodine is able to form the triiodide complex I–3 (Eq. (4)). Overall, these reactions are controlled by the experimental conditions used. The sum of the various iodine forms (i.e., iodide, molecular iodine, iodate and that from the contrast medium remaining that has not degraded following the reaction (see Figures S2 and S3)) must tend towards the iodine content at the beginning of the reaction, which is known as the parental content. Any deficit in the balance is due to compounds produced by the degradation of contrast medium compounds that still contain iodine but which it was not possible to quantify. These compounds, for each contrast medium and for each hypothesised degradation mechanism that can occur, are depicted in Figures S21 to S33. These results are in agreement with what was found by Zwiener et al. (2009) that, for iomeprol, observe a quantitative recovery of iodide at the end of the electrochemical treatment even it has been used a different cell. In conclusion, the picture taken at the end of the reactions of the iodine species shows a scenario that is in agreement with that which was expected based on the proposed mechanisms types. 4. Conclusions The electrochemical degradation was shown to be a reliable process for the treatment of ICM that can be found in contaminated waters such as hospital wastewaters or pharmaceutical waste-contaminated streams. The six most widely used active compounds have been studied, and two basic degradation reactions have been identified: (i) the reductive de-iodination of the aromatic ring and (ii) the reduction of alkyl aromatic amides to simple amides. For the specific case of amidotrizoate, the de-acylation of N-aromatic amides to produce aromatic amines and the decarboxylation process were also detected. Employing concentrations close to those found in natural environments, the degradation effectiveness was never below 85% when perchlorate was used (even at the lower current density of 64 mA/cm2 with a reaction time of 150 min) as a supporting electrolyte, while when sulphate was used, the degradation performance was above 80% for all of the studied compounds (even at the lower current density of 64 mA/cm2 with a reaction time of 150 min). Also, the high energy costs observed are partly justified by the refractoriness of these types of compounds to other treatment types (for example biological methods, which are well-known as inexpensive). In addition, the investigation of potentially harmful byproducts did not reveal any toxicity, except for the lower current density, which showed a modest toxicity, most likely for some reaction intermediates that either remain in wastewater for longer periods or which are not further degraded. Due to the peculiarity of the compounds that are required to be degraded and their massive use in the medical field, it is worth continuing the study of electro-degradation treatments of

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Iodinated contrast media electro-degradation: process performance and degradation pathways.

The electrochemical degradation of six of the most widely used iodinated contrast media was investigated. Batch experiments were performed under const...
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