key: cord-0984251-48uwo6nh
authors: Lin, Lujian; Yuan, Bo; Hong, Hualong; Li, Hanyi; He, Le; Lu, Haoliang; Liu, Jingchun; Yan, Chongling
title: Post COVID-19 pandemic: Disposable face masks as a potential vector of antibiotics in freshwater and seawater
date: 2022-01-13
journal: Sci Total Environ
DOI: 10.1016/j.scitotenv.2022.153049
sha: 05cc1aded89faae3c04636047c5ffcccc2fc4b1c
doc_id: 984251
cord_uid: 48uwo6nh

With the outbreak and widespread of the COVID-19 pandemic, large numbers of disposable face masks (DFMs) were abandoned in the environment. This study first investigated the sorption and desorption behaviors of four antibiotics (tetracycline (TC), ciprofloxacin (CIP), sulfamethoxazole (SMX), and triclosan (TCS)) on DFMs in the freshwater and seawater. It was found that the antibiotics in the freshwater exhibited relatively higher sorption and desorption capacities on the DFMs than those in the seawater. Here the antibiotics sorption processes were greatly related to their zwitterion species while the effect of salinity on the sorption processes was negligible. However, the desorption processes were jointly dominated by solution pH and salinity, with greater desorption capacities at lower pH values and salinity. Interestingly, we found that the distribution coefficient (K d) of TCS (0.3947 L/g) and SMX (0.0399 L/g) on DFMs was higher than those on some microplastics in freshwater systems. The sorption affinity of the antibiotics onto the DFMs followed the order of TCS > SMX > CIP > TC, which was positively correlated with octanol-water partition coefficient (log K ow) of the antibiotics. Besides, the sorption processes of the antibiotics onto the DFMs were mainly predominated by film diffusion and partitioning mechanism. Overall, hydrophobic interaction regulated the antibiotics sorption processes. These findings would help to evaluate the environmental behavior of DFMs and to provide the analytical framework of their role in the transport of other pollutants.

With the outbreak of the COVID-19 pandemic, citizens around the world are urged to the necessary use disposable face masks (DFMs), especially in public places (Li et al., 2021b) . According to the World Health Organization (WHO) estimated, 89 million DFMs were used globally every day to prevent the widespread of the COVID-19 pandemic and its impact could last for decades (WHO, 2020) . In the long run, DFMs will remain a necessary product to prevent the spread of the COVID-19 pandemic, hence the huge increase in global DFMs production, which will continue to increase in the coming years (Chowdhury et al., 2021; Fernández-Arribas et al., 2021) .

Unfortunately, more than 10 million DFMs are abandoned in the environment each month due to poor handling and management ( Adyel, 2020; Chen et al., 2021b) .

Once DFMs enter the environment, their impact on the environment is inevitable (Patrício Silva et al., 2021) . For example, DFMs can act as carriers of some pollutants and viruses (Mol and Caldas, 2020) in the environment, which likely exacerbates the widespread of those pollutants. Ioannidis et al. (2021) highlighted that the DFMs could act as radionuclide carriers even at ultra-trace concentrations. At solution pH 7, the distribution coefficients (K d ) values for Ra-226 ranged between 80 and 130 L/kg and for U-232 from 60 to 590 L/kg. Besides, Liu and Mabury (2021) and Fernández-Arribas et al. (2021) reported that DFMs were a potential source of synthetic phenolic and organophosphate antioxidants to the environment. Recently, DFMs releasing numerous microplastics (< 5 mm) into the aquatic environment have been reported, J o u r n a l P r e -p r o o f Journal Pre-proof greatly increasing microplastics pollution in the environment (Chen et al., 2021b; Fadare and Okoffo, 2020; Li et al., 2021b; Ma et al., 2021; Morgana et al., 2021; Shen et al., 2021) . It is estimated that 2.37 million tons of plastic waste from improperly disposed of DFMs enter the oceans each year in 46 coastal countries (Chowdhury et al., 2021) . The increase in DFMs waste is recognized as a new source of contamination directly related to the COVID-19 pandemic (Sullivan et al., 2021) .

However, information regarding the underlying roles of DFMs in the environment is extremely lacking.

Antibiotics, as a kind of the emerging pollutants in natural aqueous environments, play an irreplaceable role in the treatment of human and animal diseases as well as aquaculture and land-based agriculture (Li et al., 2021c; Nguyen et al., 2021) .

According to Zhang et al. (2015) , about 92700 tons of antibiotics were used in China in 2013, of which 53800 tons of antibiotics were released into the environment. Therefore, antibiotics were often detected in surface water, groundwater, and coastal water (Jurado et al., 2019; Li et al., 2020; Xie et al., 2019) . Importantly, the residual antibiotics in the environments resulted in the generation and spread of antibiotic-resistant genes and antibiotic-resistance bacteria, posing a potential threat to ecological functions and even human health (Šamanić et al., 2021; Zhao et al., 2021) .

Considering that the environmental impact of DFMs remains largely underestimated and the increasing antibiotic contamination. The objectives of this study were to (1) investigate the sorption and desorption behaviors of tetracycline (TC), ciprofloxacin (CIP), sulfamethoxazole (SMX), and triclosan (TCS) onto DFMs J o u r n a l P r e -p r o o f Journal Pre-proof in the freshwater and seawater systems through batch-type experiments; (2) elucidate the effects of environmental factors include salinity and pH on the sorption and desorption processes; (3) reveal the interaction mechanisms between DFMs and the antibiotics. Overall, our main purpose is to highlight that DFMs could act as a carrier of antibiotics in the environment. To the best of our knowledge, this is the first report on the potential role of DFMs as a vector of antibiotics in the freshwater and seawater systems.

We purchased DFMs from a Chinese e-commerce platform (JD. COM). The DFMs were chosen because they were the No. 1 sales on the platform in June and July 2021.

Details of the DFMs were supplied in Table S1 .

Ciprofloxacin (CIP, ≥ 98%) and sulfamethoxazole (SMX, 98%) were purchased from Shanghai Aladdin Bio-Chem Technology Co. Ltd, China. Triclosan (TCS, 99%) was purchased from Shanghai Yuanye Biological Technology Co., Ltd, China.

Tetracycline (TC, 98%) was purchased from Tianjin Xiensi Biological Technology Co., Ltd, China. The physicochemical properties of the four antibiotics are summarized in Table S2 . All the other chemicals used were analytical grade or higher purity.

The ultrapure water used was prepared from a Milli-Q water purification system (Milli-Q Reference, China). The freshwater was prepared by ultrapure water J o u r n a l P r e -p r o o f Journal Pre-proof containing 0.01 M NaCl and the solution pH was adjusted to 7. To overcome the effects of analysis and seasonal variation of natural seawater in organic matter, we used artificial seawater instead of natural seawater to prepare repeatable seawater solutions of known components. Herein, the artificial seawater was prepared according to the previous paper (Mathew et al., 2016) , as shown in Table S3 , which was almost imitated natural seawater. The 100 mg/L stock solutions of TC were directly prepared in the ultrapure water and the artificial seawater. For SMX stock solutions (100 mg/L), 5 mL methanol was added to enhance their solubility in the above background solutions. The accurately weighed 0.1 g TCS was completely dissolved in 10 mL methanol to obtain a TCS stock solution (10 g/L). Because CIP is soluble in dilute hydrochloric acid solutions, the CIP stock solutions (100 mg/L) were prepared with ultrapure water or the artificial seawater and 1% methanol at pH 3.5 (Peñafiel et al., 2021) . The solution pH was adjusted using HCl or NaOH. All the stock solutions were stored in brown reagent bottles at 4 ℃ for no longer than 7 days.

In this study, the final methanol concentration used in sorption experiments was maintained at < 0.1% (v/v) to avoid co-solvent effects during the experiments. 

J o u r n a l P r e -p r o o f

For sorption kinetics, a DFM (3.1228~3.2665 g) was put into a 250 mL brown glass reagent bottle with a lid. And then, 5 mg/L antibiotics (TC, CIP, SMX, and TCS) solutions with 200 mL volume were added to the flasks. Herein, the solution pH was adjusted to 7 (the freshwater system) or 8.18 (the seawater system) based on the different research backgrounds, respectively. The solution was collected at given time intervals (0. 5, 1, 2, 4, 6, 12, 24, 36, 48, and 60 h) , and the samples were stored in the dark at 4 ℃ until further analysis. For sorption isotherms, the initial concentrations of the four antibiotics of 1, 3, 5, 10, 20, 30, and Desorption experiments were implemented uniformly as the sorption experiments.

Specifically, the DFMs in the sorption kinetic experiments (the initial concentration of antibiotics was 5 mg/L) were taken out at the sorption equilibrium, dried at room temperature, and then transferred to a 250 mL brown reagent flask. Next, 200 mL of the freshwater or the seawater was added to the flask. The solution was collected at given time intervals (0. 5, 1, 2, 4, 6, 12, 24, 36, 48 , and 60 h).

Considering that the main differences in the composition of the freshwater and the seawater are related to NaCl and solution pH, the influences of NaCl content and J o u r n a l P r e -p r o o f Journal Pre-proof solution pH on the sorption and desorption of the antibiotics onto DFMs in ultrapure water system were investigated. The pH of the antibiotic's solution was adjusted to 3, 4, 5, 6, 7, 8, 9 , and 10, respectively. The concentration of NaCl was set (0%, 0.05%, 0.1%, 0.5%, 1%, 2%, and 3.5%) according to the gradient of salinity in the fresh-to-seawater systems (Li et al., 2021a) . Other conditions (like the initial antibiotics concentration, DFMs dosage, etc.) were the same as those used in kinetics experiments. All sorption and desorption experiments were conducted in triplicate and blank experiments (DFMs free) were conducted at the same time. To eliminate the effects of the glass wall on the antibiotic's sorption process, the concentrations measured in the blank control group were used as the initial concentrations.

The DFMs are mainly composed of four parts, namely outer layer, middle layer, inner layer, and ear band. Therefore, the DFMs were disassembled into four parts for the following characterization. The functional groups of DFMs with and without antibiotics sorbed were identified by Fourier transform infrared spectroscopy with attenuated total reflection equipment (ATR-FTIR, Bruker Vertex 70V, Germany). The water contact angles of DFMs were measured by a contact angle goniometer (JY-82B Kruss DSA, Germany). X-ray diffraction (XRD, Ultima IV, Japan) was utilized to analyze the crystal structure and morphology of the DFMs. The surface morphology and elements of DFMs with and without antibiotics sorbed were characterized using a scanning electron microscope equipped with energy dispersive X-ray analysis (SEM-EDX, FEI NANO 450, USA).

The TC, CIP, SMX, and TCS concentrations were detected by UV/Vis spectrophotometer (Agilent 8453, USA) at a wavelength of 356, 277, 275, and 282 nm, respectively.

The amounts of antibiotics sorbed onto the DFMs were calculated according to Eq.

(1):

in which t (μg/g) and t (µg/L) are the sorption capacities and concentrations of adsorbates at time t, respectively; 0 (µg/L) is the initial concentration of antibiotics; m (g) is the weight of DFMs, and V (L) represents the volume of antibiotics solution.

When the sorption reaches equilibrium, e = t , e = t , where e (µg/L) and e (μg/g) are the antibiotics concentration and sorption capacities at equilibrium, respectively.

Details of the mathematical models and the statistical parameters for the antibiotics sorption onto the DFMs (i.e., intra-particle diffusion, Boyd, Langmuir, Freundlich, and Linear models) are shown in the Supplementary Information.

The photographs, SEM images, EDX spectra, contact angles, and XRD spectra of the outer layer, middle layer, inner layer, and ear ribbon of the DFMs are shown in S1 ) (Munoz et al., 2021; Shen et al., 2021) . Besides, FTIR was employed to analyze the changes of the surface functional groups of the DFMs before and after the antibiotics sorption. As can be seen in Fig. S1 , the peaks at 2955~2835 cm -1 represent the C-H stretching vibrations of aliphatic structures (-CH, -CH 2 , and -CH 3 ) (Charles et al., 2009) . The peaks at 1000~720 cm -1 are ascribable to the C-H vibration out of plane (Muñoz-González et al., 2009) . The peaks at 1507~1336 cm -1 are assigned to C-C stretching (Chang et al., 2009) . For the ear ribbon (PET fibers), the peaks at roughly 1717 and 1246 cm -1 are attributed to the C-H and C-H stretching of -COOH J o u r n a l P r e -p r o o f (Park et al., 2005) . After the TC, CIP, SMX, and TCS sorption, no new peaks appeared or disappeared, which indicates that the sorption process is dominated by physical force .

" Fig. 1 insert here" Furthermore, intraparticle diffusion and Boyd (film diffusion) models were employed to investigate the diffusion behaviors of the antibiotics' sorption processes.

In general, intraparticle diffusion model assumes that the sorption process is proceeded by sorbate diffusion into the internal particles . The

Boyd model assumes that the sorption process is mainly controlled by intraparticle diffusion if the data in the Boyd plot are linear and pass through the origin, otherwise, film diffusion is the rate-limiting step of the sorption process (Boyd et al., 1947) . The plots of fitting intraparticle diffusion model (Q t versus t 1/2 ) are displayed in Fig. 4 , and the related parameters of K id , C i , and adj.R 2 are given in Table S4 . As shown in Fig. 4 , the sorption processes of the antibiotics onto the DFMs are divided into two or three stages (multi-linear plots), which likely suggests that multiple sorption behaviors (e.g., film diffusion, intraparticle diffusion, and dynamic equilibrium) exist in the sorption processes (Fierro et al., 2008; Wang et al., 2020) . Generally, the foremost linear stages are related to the film transport (the boundary layer diffusion), and the subsequent linear stages represent intraparticle diffusion and/or sorption dynamic equilibrium Wu et al., 2019; Zhang et al., 2018a) . Here the extending line of the second stage deviates from the origin, that is the C i ≠ 0 (Table S4) 

The dynamics equilibrium relationship between the antibiotics and the DFMs in the freshwater and seawater systems is displayed in Fig. 2 (c & d) . Like the above kinetics analyses, the antibiotics sorption capacities of the DFMs in both the freshwater and seawater systems follow the order of TCS > SMX > CIP > TC, which is positively correlated with the log K ow values of the antibiotics. The sorption capacities of the antibiotics onto the DFMs in the seawater system are significantly lower than those in the freshwater system, which may be related to the difference in salinity and solution pH in these water systems as above mention (Li et al., 2018) .

Furthermore, the sorption isotherms of the antibiotics onto the DFMs appear the same growth trend, that is, the equilibrium sorption capacities increased linearly with the increase of the initial concentration of the antibiotics. Therefore, the fitting curves of sorption isotherms based on linear model are shown in Fig. 2 Table 1 insert here"

The solution pH is a critical parameter in different aqueous solutions, which can influence the proton state and sorption performance of antibiotics. As shown in Fig. 6 (a, c, e, g), the TC, CIP, and SMX can form three species (zwitterion, cation, and anion) due to protonation and deprotonation, while TCS can only form zwitterion and anion. Interestingly, it can be clearly observed that the sorption capacities (Q e ) of the four antibiotics onto the DFMs correspondingly show a similar trend to zwitterion species fraction of the antibiotics with the change of solution pH (Fig. 6) . This phenomenon implies that the four antibiotics exist mainly in the form of molecular sorbed onto the DFMs, which indicates that the hydrophobic interaction is the main interaction mechanism between the antibiotics and the DFMs ( Furthermore, the influence of NaCl content on the sorption process was also investigated. As seen in Fig. S3 , with the increase of NaCl concentration from 0% to 3.5%, no significant variation is observed in the sorption capacities of the four antibiotics sorption onto the DFMs. This result indicates that the interaction mechanisms between the four antibiotics and the DFMs are not dominated by cation exchange and electrostatic interaction (Wang et al., 2015; Xu et al., 2018a; Yu et al., 2020a) . Meanwhile, it further highlights that hydrophobic interaction likely plays a decisive role in the antibiotics sorption process (Lin et al., 2021a) . Similar findings have also been reported in previous studies on the sorption behavior between plastics fibers and antibiotics Lin et al., 2021a Lin et al., , 2020 Xu et al., 2018a,b ; Zhang et al., 2018b) . The result also confirms that solution pH can be regarded as a key factor for the antibiotics sorption in the freshwater and seawater.

" Fig. 6 insert here"

Desorption kinetics of TC, CIP, SMX, and TCS from the DFMs in the freshwater and seawater systems are shown in Fig. S4 . All the desorption process increased drastically during the first 2 h and then reached dynamic equilibrium within 12 h. Fig.   7 summarizes the desorption rate (%) for the release of the antibiotics from the DFMs in the freshwater and seawater, and the results show two interesting phenomena. On the one hand, the desorption rate of TC, CIP, SMX, and TCS in the freshwater is 23.58%, 23.17%, 21.75%, and 12.9%, respectively, which is inversely proportional to J o u r n a l P r e -p r o o f Journal Pre-proof the log K ow of the four antibiotics. Razanajatovo et al. (2018) suggested that the high desorption rates of propranolol and sertraline were due to hydrophobic interactions with PE microplastics. Besides, more hydrophobic antibiotics (e.g., TCS and SMX with higher log K ow ) can be more easily entrapped in the inner spaces of the DFMs due to hydrophobic interactions, but this is not energetically conducive to their subsequent desorption behaviors in freshwater (Wu et al., 2020) . However, less hydrophobic antibiotics (e.g., TC and CIP) seem less likely to enter the internal field of the DFMs and then surface sorption onto the DFMs. Therefore, the desorption processes of TC and CIP from the DFMs are more reversible than that of TCS and SMX. On the other hand, the desorption rates of the antibiotics from the DFMs in the seawater are lower than those in the freshwater, which may be related to solution pH and ionic strength. As shown in Fig. S5 , the increase of NaCl content (0 ~ 3.5%) and solution pH (3 ~ 10) resulted in the decrease of the antibiotics desorption capacity from the DFMs. Herein, the main inhibiting effect of ionic strength (Na + ) can be described by the total free energy change of desorption (ΔG deso ), which is associated with the electrostatic free energy change (ΔG elect ). The ΔG elect value is related to the ionic strength in aqueous solution, can be expressed as (Huang and Smith, 1981; Tang et al., 2020) :

in which ɀ, F, and are the ionic charge, the Faraday constant, and the potential of the plane of the sorbed ion, respectively. As illustrated in Fig. S6 , the thickness of the electrical layer (above all diffusion layer) around DFMs surface is compressed when J o u r n a l P r e -p r o o f the cations are added to the aqueous solution, which causes the antibiotics to be less easily desorbed from the DFMs. The electrostatic potential decreased ( → ′ ) with the increase of cation contents, causing the coulombic energy to decrease. Importantly, the antibiotics are increasingly present as anions in the solution with solution pH values ranging from 7 (the freshwater system) to 8.18 (the seawater system). Overall, we conclude that the desorption processes of the antibiotics from the DFMs are jointly contorted by both the solution pH and ionic strength.

" Fig. 7 insert here"

Herein, the sorption and desorption behaviors of the four antibiotics (TC, CIP, SMX, and TCS) on DFMs in the freshwater and seawater were comprehensively studied. Results revealed that the antibiotics exhibited relatively higher sorption and desorption capacities on the DFMs in the freshwater than those in the seawater. The antibiotics sorption processes were greatly related to their zwitterion species while the effect of salinity on the sorption processes was negligible. However, the desorption 199, 110668. https://doi.org/10.1016/j.ecoenv.2020.110668 Lin, L., Tang, S., Wang, X., Sun, X., Liu, Y., 2021a. Sorption of tetracycline onto hexabromocyclododecane/polystyrene composite and polystyrene microplastics:

Statistical physics models, influencing factors, and interaction mechanisms.

Environ. Pollut. 284, 117164. https://doi.org/10.1016 Pollut. 284, 117164. https://doi.org/10. /j.envpol.2021 Lin, L., Tang, S., Wang, X., Sun, X., Yu, A., 2021b . Hexabromocyclododecane alters malachite green and lead(II) adsorption behaviors onto polystyrene microplastics:

Interaction mechanism and competitive effect. Ma, J., Zhao, J., Zhu, Z., Li, L., Yu, F., 2019. Effect of microplastic size on the Table 1 Comparison of K d (L/g) between the results of this work and others found in the literatures. J o u r n a l P r e -p r o o f 

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The authors would like to acknowledge the editor and three anonymous reviewers whose comments and advice greatly improved the quality of this paper. We gratefully acknowledge the support of this work by National Important Scientific Research J o u r n a l P r e -p r o o f