| Home | E-Submission | Sitemap | Contact Us |  
Environ Eng Res > Volume 28(3); 2023 > Article
Lee, Saifuddin, Park, and Kim: Synergistic effect of total residual oxidants and microplastics on seawater disinfection: An ecotoxicity study

Abstract

As fishery products are increasingly imported, a treatment system is required to prevent the introduction of invasive pathogens into domestic marine environments. In this study, disinfection efficiency was estimated through ozone treatment of Edwardsiella tarda, Streptococcus mitis, and Vibrio harveyi in seawater. The ecotoxicity of total residual oxidants (TROs) generated during the treatment process was evaluated based on the degree of bioluminescence variation in Aliivibrio fischeri. The interaction with microplastics according to their type and weight was compared in terms of bacterial inactivation and ecotoxicity, along with the level of TROs. In addition, morphological and chemical changes in the ozonated microplastics were analyzed through scanning electron microscopy and Fourier transform infrared spectroscopy. The optimal ozone doses for the inactivation rate of > 99.5% for E. tarda, S. mitis, V. harveyi, and the mixed bacteria were 3, 4, 4, and 5 mg O3/min, while the bioluminescence inhibition rates by TROs were 59.7%, 58.8%, 49.7%, and 54.4%, respectively. Furthermore, there was no considerable difference in bacterial disinfection in the presence of microplastics, but the concentrations of TROs and bioluminescence inhibition rates were slightly lowered owing to the consumption of oxidants.

1. Introduction

According to recent data (up to 2019), the annual per capita consumption of aquatic products in Korea gradually increased over the past 10 years and has recently become similar to that of rice and meat. However, while the self-sufficiency rate of aquatic products is gradually decreasing [1], the proportion of fishery product imports has increased [2]. Accordingly, fishery products must be inspected at quarantine stations along with water that may contain invasive pathogens. Moreover, such pathogens can be introduced into domestic waters through a variety of routes such as the distribution of fishery products [3], ballast water, or ocean currents. Pathogenic fish bacteria such as Vibrio, Aeromonas, Flavobacterium, Yersinia, Edwardsiella, and Streptococcus are frequently observed in ballast water and reservoirs [410]. Edwardsiella tarda, Streptococcus mitis, and Vibrio harveyi can infect humans [11, 12], resulting in liver abscesses [13], brain abscesses, infective endocarditis, sepsis, pneumonia, and peritonitis. Therefore, it is necessary to implement a water treatment system after quarantine to prevent the risk of biological factors caused by foreign pathogens in the domestic marine ecosystem.
Water treatment processes, such as antibiotic addition, ozone treatment [14], filtration, heat treatment [15], UV irradiation [16], and electrolytic treatment [17], are currently used for seawater disinfection. Among these, ozone treatment is the most widely used method [18]. However, during the treatment of quarantined seawater, ozone can react with halogen compounds (such as chlorine (Cl) or bromine (Br)) to produce undesirable oxidants that can remain in the water for days [19]. These chlorine-based or bromine-based oxidants are collectively referred to as the total residual oxidants (TROs). In particular, Br reacts rapidly with ozone molecules, which in turn can be converted into bromate (BrO3) [2022]. It is a designated carcinogen and has stringent emission standards owing to its toxicity toward various marine organisms, such as bacteria [23], algae, shrimp [24], and fish [25, 26]. Moreover, it has been demonstrated that the medium lethal concentration (LC50) for Pacific white shrimp Litopenaeus vannamei was estimated to be 0.84, 0.61, 0.54, and 0.50 mg Cl2/L for 24, 48, 72, and 96 h of exposure, respectively [24]. The LC50 of Cyprinodon variegatus was 0.35 mg Br2/L after 4 h of ozonation of seawater, whereas that of Leptocheirus plumulosus was reported to be up to 5.36 mg Br2/L after 5 h of ozonation [23]. In addition, TROs are used as a standard for the disinfection of pathogenic bacteria and discharge standards in ballast water treatment systems because they remain longer in the system than does ozone.
Countless types of pollutants such as plastics, oil, sewage, and chemicals are discharged into the ocean every year [27]. In particular, plastic debris and microplastics are emerging contaminants that rapidly spread through movement [28] due to ocean currents, wind, and tides [29]. Moreover, plastic debris can absorb toxic chemicals, posing a serious threat to various marine organisms and the entire marine system [27].
Many studies have investigated the ecotoxicity of effluents after disinfection in seawater matrices [14, 20, 23, 30]. Nevertheless, few studies have reported synergistic toxic effects of TROs and microplastics. The possible complex reactions between microplastic and disinfection byproducts (i.e., TROs) during seawater ozonation require further investigation. Therefore, this study evaluated the potential impact of microplastics of different types and weights in terms of the ecotoxicity of TROs in ozone disinfection of E. tarda, S. mitis, and V. harveyi. Physicochemical changes on the microplastic surface were analyzed using scanning electron microscopy (SEM) and Fourier transform infrared (FTIR) spectroscopy. The disinfection efficiency of bacterial pathogens was estimated by the colony count method and Adenosine Triphosphate (ATP) assay, and ecotoxicity was examined through a bioluminescence assay using Photobacterium Aliivibrio fischeri.

2. Materials and Methods

2.1. Preparation of Bacterial Pathogens

The target bacterial pathogens of E. tarda (KCTC 12267), S. mitis (KCTC 13047), and V. harveyi (KCTC 12724) were obtained from the Korean Collection for Type Cultures (KCTC, Republic of Korea). E. tarda and S. mitis were inoculated in DifcoTM Nutrient Broth (NB, BD, USA) and DifcoTM Trypticase Soy Broth (TSB, BD, USA) medium containing 3 g/L BactoTM Yeast extract (BD, USA), respectively, and cultivated at 37 °C for 24 h in a shaking incubator (Vision Scientific Co. Ltd., Republic of Korea) at 120 rpm. V. harveyi was incubated in DifcoTM Marine Broth 2216 (BD, USA) at 25 °C for 24 h in a shaking incubator at 120 rpm. Subsequently, the bacterial population density of each culture was adjusted to 7.1 × 106 CFU/mL.
Mixed bacterial suspensions were prepared by stirring each bacterial suspension adjusted to an identical cell population density at a mixing ratio of 1:1:1. For this, 50 mL of bacterial cultures of E. tarda, S. mitis, and V. harveyi were cultured according to the aforementioned culture conditions and transported into sterile conical centrifuge tubes (50 mL, Falcon®, BD, USA). Each conical tube was centrifuged at 1,000 × g for 15 min (HA-1000-3, Hanil Science Industrial Co., Ltd., Korea). The concentrated cell pellet was reconstituted in 50 mL of sterile distilled water and mixed in equal proportions. The final population density of the mixed bacterial suspension was determined by incubation on DifcoTM Trypticase Soy Agar (TSA, BD, USA) at 37 °C for 24 h in an incubator.

2.2. Disinfection of Pure and Mixed Bacterial Cultures

Bacterial pathogens were treated by ozonation in artificial seawater to determine the optimal ozone dose for achieving the desired inactivation efficiency. An ozone generator (Ozone Tech, Republic of Korea) was supplied with pure oxygen gas (99.5–99.9%) at 200 mL/min and produced ozone (O3) gas into the reactor. The ozone concentration was measured in real time using an ozone analyzer (H1-Ozone analyzer, IN USATM, USA). In addition, the gas was transferred using a Teflon tube to reduce ozone loss during ozonation. For disinfection of bacterial pathogens, 100 mL of bacterial suspension prepared as described in Section 2.1 was added to 1 L of artificial seawater (PSU = 35, pH = 8.20) in a 2 L glass beaker. Artificial seawater was prepared by dissolving 35 g of sea salt mixture (mass fractions of 55% chloride (Cl), 31% sodium (Na+), 8% sulfate (SO42−), 4% magnesium (Mg2+), 1% potassium (K+), 1% calcium (Ca2+), and < 1% other; Sigma Aldrich, USA) in 1 L of distilled water. Ozone gas was injected at doses of 1–5 mg O3/min into the aqueous sample for 10 min while being gently mixed at a stirring speed of 130 rpm on a digital magnetic stirrer (MS-20D, Daihan Scientific, Republic of Korea). Subsequently, 100 mL of the water sample was taken from the reactor for further analysis. All experiments were performed in triplicate, and the optimal ozone dose with an inactivation rate of over 99% or more for each bacterial pathogen or mixture was determined based on colony counting and ATP assay.

2.2.1. Colony count method

For the conventional colony count method, the nutrient agar plate was prepared in a sterile petri dish (D 90 mm × H 15 mm, SPL Life Sciences, Republic of Korea) with DifcoTM Nutrient Agar (NA) for E. tarda, DifcoTM TSA with 3 g/L BactoTM Yeast extract for S. mitis, DifcoTM Marine Agar 2216 for V. harveyi, and DifcoTM TSA for mixed bacterial suspension. Samples (20 μL) were serially diluted 10–106 times and were spread on agar plates followed by incubation in an incubator. After 24 h, the bacterial population density and inactivation rate were calculated from colony enumeration.

2.2.2. ATP assay

Colorimetric ATP assay was conducted using a PicoSensTM ATP assay kit (BM-ATP-100, Biomax, Republic of Korea) and a sterile 96-well cell culture plate (SPL Life Sciences, Republic of Korea). Colorimetric analytes were prepared according to the manufacturer’s protocol, and absorbance was measured at a wavelength of 570 nm using a MultiskanTM SkyHigh microplate spectrophotometer (Thermo Fisher Scientific, USA). The blank test was performed using cell-free artificial seawater. The absorbance value was converted to ATP concentration (nM) based on a standard curve. For quality control, the average R2 value of the standard curve was calculated.

2.3. TROs and Their Ecotoxicity

2.3.1. Quantification of TROs generated

To evaluate the behavior of the TROs, ozone was injected at doses of 0.2, 0.5, 0.7, and 1.0 mg O3/min into 1 L of artificial seawater for 10 min without bacterial inoculation. The concentration of TROs in mg Br2/L in ozonated artificial seawater was determined using the colorimetric N, N-diethyl-p-phenylenediamine (DPD) reagent method, as proposed in the US EPA Method 8016. Briefly, 10 mL of the water sample was collected at retention times of 0, 1, 3, 6, 12, 24, 48, 60, 72, 96, 120, 144, 168, and 192 h. Color intensity was measured using a DR 2800TM portable spectrophotometer (Hach, USA) after adding the DPD reagent (DPD Total Chlorine Reagent Powder Pillow, Hach, USA) at room temperature.

2.3.2. Aliivibrio fischeri bioluminescence inhibition assay

In this study, to evaluate the ecotoxicity of TROs toward the marine ecosystem, TROs generated by seawater ozonation were diluted with artificial seawater at concentrations of 0.5, 1, 1.5, 2, 2.5, 3, 4, 5, 6, 7, and 8 mg Br2/L. Ecotoxicity against the photobacterium Aliivibrio fischeri was determined using a BioTox LumoPlate Ultimate Matrix Kit (EBPI, Canada). All the analytical procedures were performed according to the manufacturer’s instructions. Briefly, the lyophilized Aliivibrio fischeri reagent was reconstituted by pouring a chilled (4 °C) reagent diluent (Luminescent Bacterial Assays, BioTox, EBPI, Canada). The reconstituted reagents were stored in a refrigerator (4–6 °C) and then incubated at 15 °C for an additional 30 min before the assay. Meanwhile, 500 μL of osmotic adjustment solution (OAS, EBPI, Canada) was added to 4.5 mL of the sample to optimize the salinity. Next, the samples were serially diluted 5–10 times using a sample diluent reagent, and then 100 μL of the sample was transferred to an opaque 96-well cell culture plate (SPL Life Sciences, Republic of Korea). Aliivibrio fischeri reagent (100 μL) was added immediately before loading into the luminometer. Luminescence intensity was recorded after 0, 15, and 30 min using a SpectraMax® L Microplate Reader (Molecular Devices, USA). Standard linear regression analysis was performed for a linear comparison between the logarithm of the concentration of toxic compounds (i.e., TROs) and the logarithm of the luminescent response.
In addition, the inhibition rate of TROs in the effluent after ozone treatment of the bacterial pathogens was measured in the same manner. The percentage of bioluminescence inhibition was calculated using Equation (1):
(1)
ΔL=(L0-Lt)/Lo×100
where ΔL is bioluminescence inhibition (%), L0 is the initial luminescence (time = 0 min), and Lt is luminescence after time t.

2.4. Ozonation of Quarantined Seawater and Microplastics

Four types of microplastic samples, polyethylene (PE), polystyrene (PS), polypropylene (PP), and polyethylene terephthalate (PET), with sizes ranging from 100–500 μm were used in this study. PE microplastics were procured from Sigma-Aldrich (USA), whereas the other three types of microplastics (PS, PP, and PET) were obtained from the Korea Testing & Research Institute (KTR, Republic of Korea). The samples were stored in a vacuum desiccator at room temperature prior to use. For ozonation, 100 mL of either single bacterial culture or mixed bacterial suspensions, which were adjusted to a population density of 7.1 × 106 CFU/mL, were poured into 1 L of artificial seawater (PSU = 35, pH = 8.20). The microplastic sample, with different weights of 0.05 g, 0.25 g, and 0.50 g, was individually added to the artificial seawater. Artificial seawater containing bacterial pathogens and microplastics was gently stirred using a digital magnetic stirrer at 130 rpm, while ozone gas was constantly injected. After ozonation for 10 min, the concentration of TROs and their ecotoxicity were analyzed along with pH and salinity measurements. In addition, 100 mL of each sample was transferred into sterile conical centrifuge tubes to determine the inactivation efficiency through the ATP assay, while another 500 mL of the sample was filtered onto a glass fiber filter (GF/C, effective pore size of 1.2 μm, Whatman, UK) to recover the microplastics. They were rinsed with distilled water and oven-dried at 50 °C (Vision Scientific Co., Ltd., Korea) for 24 h.
In addition, ecotoxicity was estimated to confirm the interaction between TROs and microplastics using the Aliivibrio fischeri bioluminescence inhibition assay. Different weights of microplastics were added to artificial seawater with various concentrations of TROs (0.5, 1.0, 1.5, 2.0, 2.5, 3, 4, 5, 6, 7, and 8 mg Br2/L). The bioluminescence inhibition rate was determined using Aliivibrio fischeri in the same manner as mentioned before, and the results were compared with those without microplastics.

2.5. Characterization of Microplastics

The morphological changes in the oxidized microplastics were observed using a field emission-scanning electron microscope (FE-SEM, S-4300SE, Hitachi High Technologies Co., Japan) at 15.0 kV. The samples were mounted onto a sample holder surface with carbon tape and then sputter-coated with Pt:Pd for 120 s. Fourier transform infrared (FTIR) spectroscopy (NicoletTM iSTM 5, Thermo Scientific, USA) was used to analyze the surface chemistry of the microplastics. It was measured in the wavenumber range of 4,000–400 cm−1 with a resolution of 4 cm−1. The microplastic samples were then pelletized using potassium bromide (KBr, FTIR grade, Sigma, USA). Subsequently, the obtained spectra were compared with the characteristic peaks of polymers [31] and a standard FTIR spectrum database was used from the OMNIC software.

2.6. Quenching TROs

A reducing agent, sodium thiosulfate (Na2S2O3), was used to quench TROs. After ozone treatment, 50 mM of Na2S2O3 solution was added as a scavenger of TROs to the artificial seawater, and the final TROs concentrations and ecotoxicity were determined using the DPD method and the Aliivibrio fischeri bioluminescence inhibition assay, respectively.

3. Results and Discussion

3.1. Disinfection of Bacterial Pathogens

Ozone generally promotes the oxidation of glutathione, nucleic acids, and amino acids together with cell lysis [32, 33], where the extent of their reaction depends on the cell wall structure and bacterial cell content extravasation [34]. Previous studies have demonstrated that different bacterial species have different resistance to ozone oxidation [35, 36]. Therefore, before evaluating TROs and ecotoxicity, in this study, the degree of ozone disinfection affecting target bacterial pathogens was estimated, from which their optimal ozone doses were subsequently obtained. The inactivation rates under various ozone doses were evaluated using the colony counting method and ATP assay and are shown in Figs. 1 and 2 for each bacterial pathogen and mixture, respectively. First, the inactivation rate of E. tarda was estimated to be less than 96% at 2 mg O3/min for both colonies counting [Fig. 1 (a)] and the ATP assay [Fig. 1 (b)]. Further, it was found to be more than 99% disinfected at 3 mg O3/min. In particular, E. tarda was presumed to be the most ozone-sensitive pathogen among the target bacterial species examined in this study. In contrast, S. mitis had a less detrimental effect than E. tarda [Figs. 1 (c) and (d)], from which the optimal ozone dose required for the inactivation of S. mitis (> 99%) was estimated to be 4 mg O3/min. Conversely, ozone treatment at 3 mg O3/min for 10 min yielded only 94% inactivation, which is ~ 5% lower than that of E. tarda. Similarly, V. harveyi was inactivated by the addition of 4 mg O3/min, as shown in Figs. 1(e) and (f). Therefore, the optimal ozone dosages for E. tarda, S. mitis, and V. harveyi were determined to be 3, 4, and 4 mg O3/min, respectively.
In the meantime, there was a high correlation between the number of cells and the ATP assay showing a correlation coefficient (R2) of 0.995 (p < 0.05), suggesting that the ATP assay, which can be successfully implemented in a much shorter time (within 30 min), is sufficient to replace the colony counting method, even in seawater ozonation. Nevertheless, ATP levels were measured at different levels according to the type of bacterial pathogen. Initial ATP levels were measured at 1,500 nM for E. tarda [Fig. 1 (b)], 1,440 nM for S. mitis [Fig. 1 (d)], and 1,176 nM for V. harveyi [Fig. 1 (f)], although the bacterial population density was maintained at the same level of 7.1 × 106 CFU/mL at the beginning of the test. Therefore, it was considered for disinfecting real seawater containing various bacterial pathogens in the ozone treatment.
For the mixed bacterial suspensions, the ATP level of non-treated cells was 1,304 nM, which was similar to the average value for the three different bacterial pathogens, as previously shown in Fig. 1. However, it has been implied that disinfection of mixed pathogens requires a higher ozone dose of 5 mg O3/min, compared to single bacterial cultures, as shown in Figs. 2(a) and (b).

3.2. Behavior of Generated TROs in the Bacteria-free Seawater and Their Ecotoxicity

TROs could be proportionally generated up to 1.75, 4.97, 5.58, and 6.84 mg Br2/L as the ozone injected at doses of 0.2, 0.5, 0.7, and 1.0 mg O3/min for 10 min in bacteria-free seawater. Moreover, temporal monitoring of TROs without further treatment (Fig. 3) demonstrated that they were gradually decomposed with prolonged retention time, although they remained at concentrations above 0.1 mg Br2/L for up to 8 days. Additionally, in the case of TROs generated after the addition of 1.0 mg O3/min, the concentration of TROs decreased sharply at the beginning for 6 h and then further decreased at a relatively constant rate. It could be suggested that if TROs are continuously released into nearby coastal waters, even at low concentrations, they may be retained for several days with negligible degradation, possibly causing adverse effects in the marine ecosystem [37, 38].
The Photobacterium Aliivibrio fischeri bioluminescence inhibition assay has been widely used for toxicity monitoring because of its shorter test duration, higher sensitivity, and simple manipulation [39], irrespective of the type of environmental matrix. It has been implemented to assess the toxicity of halogenic disinfection by-products generated during the treatment of wastewater [40] and ballast water [41]. Accordingly, the ecotoxicity of TROs after ozone treatment of artificial seawater was evaluated employing the Aliivibrio fischeri. Artificial seawater that was not treated with ozone was first confirmed to be non-toxic to test bacteria. Bioluminescence inhibition occurred immediately upon contact with a minimum of 0.5 mg Br2/L of TROs (Fig. 4). The inhibition reached 36% after 30 min of incubation at 0.5 mg Br2/L of TROs. As the concentration of TROs increased, the bioluminescence of Aliivibrio fischeri was significantly inhibited, with the maximum inhibition rate estimated to be 90% at 8 mg Br2/L of TROs. The half-maximal effective concentration (EC50) was determined according to ISO 11348 [42] Part 3 (method using freeze-dried bacteria). The EC50 value was 3.05 and 1.33 mg Br2/L after the exposure to TROs for 15 and 30 min, respectively, from the standard linear regression analysis (Fig. S1).

3.3. Synergistic Effects of TROs and Microplastics

The coexistence of microplastics can affect the disinfection efficiency and behavior of TROs after ozonation. Therefore, in this study, the ecotoxicity of TROs was estimated using a bioluminescence assay in artificial seawater containing microplastics of various types and weights. Subsequently, the effects of microplastics on bacterial disinfection, ecotoxicity, and TROs monitoring were investigated.

3.3.1. Synergistic effect on ecotoxicity: EC50

Fig. 5 shows the EC50 values of TROs according to the added microplastic types (PE, PS, PP, and PET) and weights (0.05, 0.25, and 0.50 g) in bacteria-free seawater ozonation, which were compared with those of the control sample (i.e., without microplastics). In the control, as previously mentioned, the EC50 was determined to be 3.05 and 1.33 mg Br2/L after 15 and 30 min of incubation. This indicates that high concentrations of toxic compounds can inhibit bioluminescence in a shorter duration. Conversely, if the exposure time was longer, 50% bioluminescence could be suppressed even at a low concentration. Notably, however, when microplastics were added to seawater, there was no significant difference in the EC50 value for TROs according to reaction time (15 and 30 min), and it was found that the value of EC50 for a reaction time of 30 min was commonly increased compared to the control. This can be attributed to the uptake of TROs by the microplastics, thereby reducing the degree of toxicity of TROs to microorganisms. Microplastics can consume oxidizing agents in aquatic environments. It has been suggested that microplastics are weathered by reactive oxygen species (ROS) under different conditions of alteration properties (such as size and surface characteristics), exposure time, and adsorption capacity [43]. In particular, the absorption capacity of microplastics is related to the generation of ROS, depending on the type of oxidant [43, 44]. This was presumed to be the mechanism of ROS generation and ozone uptake by microplastics [45]. Moreover, in our previous study, ozone uptake by PE microplastics was shown to occur through a given ozone treatment system, although it was performed in freshwater [45].
In the presence of TROs and microplastics, the toxic effects were alleviated according to the microplastic type and amount added. The EC50 of samples containing PE microplastic were measured to be 5.43, 1.92, and 2.97 mg Br2/L at 15 min, 2.61, 1.93, and 3.04 mg Br2/L at 30 min of incubation depending on the amounts of 0.05, 0.25, and 0.5 g, respectively. For PS microplastics, they were 3.00, 2.92, and 2.96 mg Br2/L at 15 min and 2.50, 2.95, and 2.81 mg Br2/L at 30 min, while for PP microplastics, they were determined to be 3.19, 2.58, and 2.57 mg Br2/L at 15 min, and 2.48, 2.76, and 2.53 mg Br2/L at 30 min, respectively. In addition, for PET microplastics, the values were 2.63, 2.58, and 2.61 mg Br2/L at 15 min, and 2.50, 2.77, and 2.59 mg Br2/L at 30 min for the different addition conditions. In summary, PS microplastics mostly mitigated the toxic effect of TROs with high EC50 values under the given conditions, followed by PP, PET, and PE, while the amount added was less affected by changing the EC50 value, even with variability in the case of PE microplastics.

3.3.2. Effect of microplastics on the disinfection efficiency

The effect of microplastics on the disinfection efficiency of bacterial pathogens according to the ozone treatment conditions was comparatively studied, as shown in Figs. 6(a), (c), (e), and (g). Herein, the bacterial inactivation rate was determined by the ATP assay, in which the initial ATP concentrations were 1,384 nM, 1,541 nM, 1,019 nM, and 1,274 nM for E. tarda, S. mitis, V. harveyi, and the mixed bacterial suspension, respectively, which were all equivalent to a bacterial population density of 7.1 × 106 CFU/mL. No significant evidence was found that microplastics were involved in the ozone disinfection process. The inactivation rate exceeded 99.5% when the same level of ozone as that of seawater without microplastics was supplied to the reactor for each pathogenic bacterium of E. tarda, S. mitis, V. harveyi, and mixed bacterial pathogens, as shown in Figs. 6(a), (c), (e), and (g), respectively. In addition, there was no significant difference based on the type and weight of microplastics added to the seawater.
Other studies have demonstrated that microplastics can provide a protective barrier against bacteria, which can be resistant to chlorine and ultraviolet disinfection [46]. It has also been reported that microplastics can affect various disinfection processes, including ozonation. Moreover, microplastics have been regarded as an important factor in lowering ozone treatment efficiency owing to ozone consumption by microplastics and other organic pollutants adsorbed on the surface of microplastics [4749]. Microplastics can reduce the efficiency of ozonation for two reasons [48]. First, ozone has a strong affinity for various organic contaminants adsorbed on microplastic surfaces, resulting in the generation of various derivatives [50, 51]. Second, ozone can serve as a strong oxidizing agent that can oxidize microplastics [45]. However, in the present study, the effect of microplastics on the bacterial inactivation efficiency was not observed because ozone treatment was performed for a short duration at low ozone concentrations without giving enough time for them to react with the ozone dose.

3.3.3. Effect of microplastics on the TRO generation and bioluminescence

The concentration of TROs was found to be slightly lower than that of the control (i.e., without microplastics), as shown in Figs. 6(b), (d), (f), and (h). This implies that they were not significantly affected by the type or amount of microplastics but were more dependent on the types of microorganisms to be disinfected. In the disinfection of E. tarda, TROs concentration was shown to be average of (2.21 ± 0.07) mg Br2/L for PE, (2.25 ± 0.07) mg Br2/L for PS, (2.29 ± 0.02) mg Br2/L for PP, and (2.30 ± 0.01) mg Br2/L for PET, while it was slightly higher (2.37 mg Br2/L) for the control. For S. mitis, TROs were generated with the level of (2.21 ± 0.09) mg Br2/L, (2.26 ± 0.08) mg Br2/L, (2.21 ± 0.06) mg Br2/L, and (2.17 ± 0.03) mg Br2/L in the seawater containing PE, PS, PP, and PET microplastics, respectively, which were slightly lower than that of control (2.25 mg Br2/L). Lower TROs were detected in the case of V. harveyi than in other bacteria, at (1.42 ± 0.14) mg Br2/L for PE, (1.50 ± 0.02) mg Br2/L for PS, (1.49 ± 0.03) mg Br2/L for PP, (1.51 ± 0.02) mg Br2/L for PET, and 1.51 mg Br2/L for the control. Meanwhile, for pathogenic consortium, TROs concentration was estimated to be (2.07 ± 0.07) mg Br2/L, (1.79 ± 0.13) mg Br2/L, (2.02 ± 0.08) mg Br2/L, and (2.02 ± 0.05) mg Br2/L in the presence of PE, PS, PP, and PET microplastics, respectively, while it was estimated to be 2.01 mg Br2/L for the control. In the meantime, there were no observed changes in pH and salinity during the test period.
Fig. 7 and Table S1 show the ecotoxicological effects of TROs in the presence of microplastics after ozone disinfection. The bioluminescence inhibition rate after 15 min of incubation was generally less than 40% in the samples containing TROs with or without the addition of microplastics, which was consistent with the EC50 result shown in Fig. 5. As the incubation time increased, the inhibition of bioluminescence increased to 59.7, 58.8, 49.7, and 54.4% for E. tarda, S. mitis, V. harveyi, and mixed bacterial pathogens, respectively, in the absence of microplastics. In the presence of microplastics, the mean bioluminescence inhibition rate tended to decrease slightly or remain similar to that of the control.
Nonetheless, there were no significant differences among the test groups injected with the various types of microplastics. This simply indicates that the bioluminescence inhibition could not occur by the addition of microplastics due to insufficient biological or chemical interactions with far lower concentrations of ozone and for a contact time of less than 10 min. In addition, it might be presumed that keeping its own property equivalent to that of the pristine microplastic, after being oxidized shortly, could have a less toxic impact on the test organisms [52, 53], in contrast to revealing ecotoxicological effects [5456]. It also demonstrates that pristine microplastics have been scarcely altered by ozonation because they possess much fewer oxygen-containing functional groups, such as carbonyl groups, which are more resistant to oxidation than environmentally aged microplastics [57, 58]. This is further discussed in the following section, along with the analysis of SEM and FTIR results.

3.4. Characterization of Microplastics After Ozone Disinfection

In a comparison of the surface images of pristine and ozonated microplastics obtained through SEM analysis (Figs. S2, S3, S4, and S5), there was no significant morphological damage to the microplastic surfaces after customized ozonation with 3–5 mg O3/min for 10 min. No microorganisms were adsorbed on the microplastic surface, possibly because of the short contact time of 10 min and severe cellular destruction by ozone. Several studies have reported that microbial colonization can contribute to microplastic surface modification [5961], after which, the biofilm community is formed over a duration of a few days [6264]. However, in the present study, no biofilm formation by viable bacterial cells was observed.
The FTIR spectra of the microplastics were compared with a standard FTIR spectral database built into OMNIC software (Thermo ScientificTM, USA). FTIR analysis provides vivid information about the intermolecular interactions corresponding to the stretching or bending vibrations of specific bonds, as the positions at which these peaks appear depend on the bond types [65, 66]. The FTIR absorption bands associated with all pristine microplastics of PE, PS, PP, and PET were typically observed in the 4,000–400 cm−1 region (Figs. S6, S7, S8, and S9).
For the PE microplastics, strong CH2 asymmetric stretching and CH2 symmetric stretching could be observed in the 3,000–2,800 cm−1 region of the spectrum [Fig. S6(a)]. Peaks of strong bending deformation and weak symmetric deformation could be observed in the 1,550–1,400 cm−1 and 1,400–1,350 cm−1 regions, respectively. A medium rocking deformation peak can be found in the 750–650 cm−1 spectral region [66]. However, there was no significant change in the peak positions and intensities of the bands of all the treated PE samples compared to those of the pristine PE (Fig. S6).
For PS, as illustrated in Fig. S7, two peaks of aromatic C–H stretching vibrations were seen in the range of 3,100–3,000 cm−1, whereas the peaks in the region of 3,500–3,400 cm−1 indicated the presence of O–H groups. The broad band in the region of 1,600−1,400 cm−1 included three peaks of aromatic C=C stretching vibrations, indicating the presence of benzene rings. In addition, the absorption peaks in the range of 800–650 cm−1 corresponded to the C–H out-of-plane bending vibration absorption and represented only one substituent in the benzene ring. The two peaks in the 3,000–2,800 cm−1 region were attributed to the presence of methylene in the structure of the PS polymer [6769]. Meanwhile, the FTIR spectra of pristine PS and the 12 treated PS microplastics showed identical peaks and intensities (Fig. S7).
In the case of PP (Fig. S8), the vibration in the 3,000–2,900 cm−1 region was due to strong asymmetric stretching by CH2 and CH3, and the stretching in the 2,900–2,800 cm−1 region was assigned to the CH3 group. The band in the region 3,500–3,400 cm−1 was caused by the stretching vibration of the O–H groups. The two peaks formed by the symmetrical bending vibration in the region 1,500–1,350 cm−1 suggest the presence of CH3 groups. A wagging vibration at ~1,200 cm−1 was observed owing to the presence of the C–H bond. The two peaks at 1,000–900 cm−1 were assigned to the stretching and rocking vibrations of the CH3 group and the C–C bond. The weak vibrations in the 850–800 cm−1 range were due to the stretching of the C–C bond, rocking of the C–H bond, and CH3 group [70, 71]. The FTIR spectra of pristine PP and all treated PP samples (Fig. S8) exhibited similar results, indicating no substantial changes in the bands and peak intensities.
Finally, the region from 1,500 cm−1 to 500 cm−1 in PET microplastics contains a complicated series of absorption peaks and is known as the fingerprint region (Fig. S9). The band at 1,725 cm−1 was attributed to the stretching vibration of the C=O bond of the ester group. Two bands exist for C–O stretching in the region 1,250–1,100 cm−1. The absorption bands associated with C=O bond stretching were generally very strong in this mode because a large adjustment occurs in the dipole. On the contrary, there are two types of vibrations in the C–H bond of the ethyl group: (i) the C–H stretching band of the aromatic ring present at 3,055 cm−1 and (ii) the medium C–H stretching bond found at 3,050–2,900 cm−1. In addition, the bands at 1,600–1,550 cm−1 were assigned to the C–H bond stretching vibrations of the phenyl ring [65, 72]. The FTIR spectra of the PET microplastic also exhibited similar results, with no change in the pristine and treated microplastic samples (Fig. S9).
In summary, none of the four different types of microplastic samples used in this study underwent surface modification, including changes in the chemical composition and bonding. Although some studies have suggested that ozonation or other oxidation processes may lead to changes in C-H stretching, C=O groups, and OH stretching [67], ozone treatment did not affect the chemical structure of the microplastic samples under the given conditions.

3.5. Quenching TROs in the Effluent

Disinfection and ecotoxicity assays confirmed that TROs persisted and accumulated in the aquatic environment and might have potential ecological toxicity to aquatic ecosystems. Therefore, post-treatment to quench these oxidative substances before discharge is required to maintain a sustainable marine ecosystem. In this study, the TROs were quenched using a reducing agent, sodium thiosulfate (Na2S2O3) [73]. As a result, the concentration of TROs in the effluent significantly reduced to less than 0.05 mg Br2/L, and the bioluminescence inhibition rate was estimated to be similar to that of artificial seawater without ozone treatment (corresponding to inhibition rate of 3%), indicating that toxicity was completely neutralized.

4. Conclusion

In this study, a seawater ozone treatment system at quarantine stations was investigated to prevent the introduction of invasive bacterial pathogens into domestic coastal environments. The TROs generated during ozone treatment can adversely affect aquatic ecosystems. For this reason, the ecotoxicity of TROs was estimated based on the Aliivibrio fischeri bioluminescence assay under various disinfection conditions. The toxic effect of TROs was further monitored after bacterial disinfection in the presence of microplastics, and the surface modification of microplastics was analyzed. To summarize them,
  1. The optimal ozone dose for a bacterial inactivation rate of > 99% was 3 mg O3/min for E. tarda, 4 mg O3/min for S. mitis, 4 mg O3/min for V. harveyi, and 5 mg O3/min for the mixed bacterial suspension. Concurrently, the concentrations of TROs generated were estimated to be 2.37, 2.25, 1.51, and 2.01 mg Br2/L, and the bioluminescence inhibition rates against TROs were estimated to be 59.7%, 58.8%, 49.7%, and 54.4%, respectively.

  2. EC50 values were determined to be 3.05 mg Br2/L after 15 min and 1.33 mg Br2/L after 30 min of incubation with Aliivibrio fischeri. The addition of microplastics slightly lowered ecotoxicity by consuming oxidants (i.e., ozone and TROs).

  3. The coexistence of microplastics had little effect on bacterial inactivation, irrespective of the type and amount of microplastics. The concentration of TROs and the bioluminescence inhibition rate were slightly lower than those in the control group. Moreover, no noticeable morphological or chemical changes were observed around the microplastic surface during the reaction time.

Nevertheless, these disinfection byproducts can combine with other marine pollutants to create antagonistic effects on the ecosystems. Therefore, they should be controlled by adding a scavenger (e.g., sodium thiosulfate) before being discharged into the environment. Further studies are needed on the ecotoxicity of TROs in combination with different types of microorganisms and various pollutants, particularly environmentally weathered microplastics.

Supplementary Information

Acknowledgment

This research was a part of the project titled “Development of a water treatment system to remove harmful substances from ecological disturbances emitted from quarantine stations such as imported fishery products” (No. 2018341) supported by the Korea Institute of Marine Science and Technology Promotion and partially supported by an INHA University Grant.

Notes

Conflict-of-Interest Statement

The authors declared that they have no conflict of interest.

Author Contribution

Y.R.L (M.S.) performed the conceptualization, methodology, investigation, formal analysis, and data curation and wrote the original draft. M. S (Post-doctoral associate) conducted the visualization and curation of the experimental data and reviewed and edited the manuscript. S.Y.P (Post-doctoral associate) performed visualization and curation of the experimental data, validation, project administration, methodology, and comprehensive revision of the manuscript. C.G.K (Professor) performed conceptualization, supervision, validation, funding acquisition, and comprehensive revision of the manuscript.

References

2. Ministry of Oceans and Fisheries (MOF). Statistics on the seafood import and export [Internet]. MOF; 2019. [cited 27 September 2021]. Availabe from: https://www.data.go.kr/data/15057607/openapi.do


3. Park SY, Kim JH, Kim CG. Assessment of the ozonation against pathogenic bacteria in the effluent of the quarantine station. J. Mar. Biosci. Biotechnol. 2021;13(1)10–19. https://doi.org/10.15433/ksmb.2021.13.1.010
crossref

4. Baeck GW, Kim JH, Gomez DK, Park SC. Isolation and characterization of Streptococcus sp. from diseased flounder (Paralichthys olivaceus) in Jeju Island. J. Vet. Sci. 2006;7(1)53–58. https://doi.org/10.4142/jvs.2006.7.1.53
crossref

5. Han HJ, Kim DH, Lee DC, Kim SM, Park SI. Pathogenicity of Edwardsiella tarda to olive flounder, Paralichthys olivaceus (Temminck & Schlegel). J. Fish Dis. 2006;29(10)601–609. https://doi.org/10.1111/j.1365-2761.2006.00754.x
crossref

6. Roh HJ, Kim A, Kang GS, Kim DH. Photoinactivation of major bacterial pathogens in aquaculture. Fish. Aquat. Sci. 2016;19:28 https://doi.org/10.1186/S41240-016-0029-5
crossref

7. Won KM, Park SI. Pathogenicity of Vibrio harveyi to cultured marine fishes in Korea. Aquaculture. 2008;285(1–4)8–13. https://doi.org/10.1016/j.aquaculture.2008.08.013
crossref

8. Xiao J, Wang Q, Liu Q, Wang X, Liu H, Zhang Y. Isolation and identification of fish pathogen Edwardsiella tarda from mariculture in China. Aquac. Res. 2008;40(1)13–17. https://doi.org/10.1111/j.1365-2109.2008.02101.x
crossref

9. Zhang XH, He X, Austin B. Vibrio harveyi: a serious pathogen of fish and invertebrates in mariculture. Mar. Life Sci. Technol. 2020;2:231–245. https://doi.org/10.1007/s42995-020-00037-z
crossref pmid pmc

10. Weinstein MR, Litt M, Kertesz D, et al. Invasive infections due to a fish pathogen, Streptococcus iniae. N. Engl. J. Med. 1997;337:589–594. https://doi.org/10.1056/NEJM199708283370902
crossref pmid

11. Goh SH, Driedger D, Gillett S, et al. Streptococcus iniae, a human and animal pathogen: Specific identification by the chaperonin 60 gene identification method. J. Clin. Microbiol. 1998;36(7)2164–2166. https://doi.org/10.1128/jcm.36.7.2164-2166.1998
crossref pmid pmc

12. Slaven EM, Lopez FA, Hart SM, Sanders C. Myonecrosis caused by Edwardsiella tarda: A case report and case series of extra-intestinal E. tarda infections. Clin. Infect. Dis. 2001;32(10)1430–1433. https://doi.org/10.1086/320152
crossref pmid

13. Bakirova GH, Alharthy A, Corcione S, et al. Fulminant septic shock due to Edwardsiella tarda infection associated with multiple liver abscesses: A case report and review of the literature. J. Med. Case Rep. 2020;14:144 https://doi.org/10.1186/s13256-020-02469-8
crossref pmid pmc

14. Perrins JC, Cooper WJ, van Leeuwen J, Herwig RP. Ozonation of seawater from different locations: Formation and decay of total residual oxidant-implications for ballast water treatment. Mar. Pollut. Bull. 2006;52(9)1023–1033. https://doi.org/10.1016/j.marpolbul.2006.01.007
crossref pmid

15. Whipple MJ, Rohovec JS. The effect of heat and low pH on selected viral and bacterial fish pathogens. Aquaculture. 1994;123(3–4)179–189. https://doi.org/10.1016/0044-8486(94)90056-6
crossref

16. Chang PS, Chen LJ, Wang YC. The effect of ultraviolet irradiation, heat, pH, ozone, salinity and chemical disinfectants on the infectivity of white spot syndrome baculovirus. Aquaculture. 1998;166(1–2)1–17. https://doi.org/10.1016/S0044-8486(97)00238-X
crossref

17. Jorquera MA, Valencia G, Eguchi M, Katayose M, Riquelme C. Disinfection of seawater for hatchery aquaculture systems using electrolytic water treatment. Aquaculture. 2002;207(3–4)213–224. https://doi.org/10.1016/S0044-8486(01)00766-9
crossref

18. Langlais B, Reckhow DA, Brink DR. Ozone in Water Treatment Processes: Application and Engineering. Lewis Publishers; 1991.


19. Heeb MB, Criquet J, Zimmermann-Steffens SG, Von Gunten U. Oxidative treatment of bromide-containing waters: Formation of bromine and its reactions with inorganic and organic compounds - A critical review. Water Res. 2014;48:15–42. https://doi.org/10.1016/j.watres.2013.08.030
crossref pmid

20. Grguric G, Trefry JH, Keaffaber JJ. Ozonation products of bromine and chlorine in seawater aquaria. Water Res. 1994;28(5)1087–1094. https://doi.org/10.1016/0043-1354(94)90194-5
crossref

21. Herwig RP, Cordell JR, Perrins JC, et al. Ozone treatment of ballast water on the oil tanker S/T Tonsina: Chemistry, biology and toxicity. Mar. Ecol. Prog. Ser. 2006;324:37–55. https://doi.org/10.3354/meps324037
crossref

22. Von Gunten U, Oliveras Y. Advanced oxidation of bromide-containing waters: Bromate formation mechanisms. Environ. Sci Technol. 1998;32(1)63–70. https://doi.org/10.1021/es970477j
crossref

23. Jones AC, Gensemer RW, Stubblefield WA, Van Genderen E, Dethloff GM, Cooper WJ. Toxicity of ozonated seawater to marine organisms. Environ. Toxicol. Chem. 2006;25(10)2683–2691. https://doi.org/10.1897/05-535R.1
crossref pmid

24. Schroeder JP, Gärtner A, Waller U, Hanel R. The toxicity of ozone-produced oxidants to the Pacific white shrimp Litopenaeus vannamei. Aquaculture. 2010;305(1–4)6–11. https://doi.org/10.1016/j.aquaculture.2010.03.030
crossref

25. Powell A, Chingombe P, Lupatsch I, Shields RJ, Lloyd R. The effect of ozone on water quality and survival of turbot (Psetta maxima) maintained in a recirculating aquaculture system. Aquac. Eng. 2015;64:20–24. https://doi.org/10.1016/j.aquaeng.2014.11.005
crossref

26. Reiser S, Wuertz S, Schroeder JP, Kloas W, Hanel R. Risks of seawater ozonation in recirculation aquaculture - Effects of oxidative stress on animal welfare of juvenile turbot (Psetta maxima, L.). Aquat. Toxicol. 2011;105(3–4)508–517. https://doi.org/10.1016/j.aquatox.2011.08.004
crossref pmid

27. Vikas M, Dwarakish GS. Coastal Pollution: A Review. Aquat Procedia. 2015;4:381–388. https://doi.org/10.1016/j.aqpro.2015.02.051
crossref

28. Derraik JGB. The pollution of the marine environment by plastic debris: A review. Mar. Pollut. Bull. 2002;44(9)842–852. https://doi.org/10.1016/S0025-326X(02)00220-5
crossref pmid

29. Wang J, Tan Z, Peng J, Qiu Q, Li M. The behaviors of microplastics in the marine environment. Mar. Environ. Res. 2016;113:7–17. https://doi.org/10.1016/j.marenvres.2015.10.014
crossref pmid

30. Gonçalves AA, Gagnon GA. Seawater ozonation: effects of seawater parameters on oxidant loading rates, residual toxicity, and total residual oxidants/by-products reduction during storage time. Ozone Sci. Eng. 2018;40(5)399–414. https://doi.org/10.1080/01919512.2018.1448705
crossref

31. Veerasingam S, Ranjani M, Venkatachalapathy R, et al. Contributions of Fourier transform infrared spectroscopy in microplastic pollution research : A review. Crit. Rev. Environ Sci. Technol. 2021;51(22)2681–2743. https://doi.org/10.1080/10643389.2020.1807450
crossref

32. Broadwater WT, Hoehn RC, King PH. Sensitivity of three selected bacterial species to ozone. Appl. Microbiol. 1973;26(3)391–393. https://doi.org/10.1128/am.26.3.391-393.1973
crossref pmid pmc

33. Fontes B, Cattani Heimbecker AM, de Souza Brito G, et al. Effect of low-dose gaseous ozone on pathogenic bacteria. BMC Infect. Dis. 2012;12:358 https://doi.org/10.1186/1471-2334-12-358
crossref pmid pmc

34. Scott DB, Lersher EC. Effect of ozone on survival and permeability of Escherichia coli. J. Bacteriol. 1963;85(3)567–576. https://doi.org/10.1128/jb.85.3.567-576.1963
crossref pmid pmc

35. Lezcano I, Pérez Rey R, Baluja C, Sánchez E. Ozone inactivation of Pseudomonas aeruginosa, Escherichia coli, Shigella sonnei and Salmonella typhimurium in water. Ozone Sci. Eng. 1999;21(3)293–300. https://doi.org/10.1080/01919519908547242
crossref

36. Lezcano I, Rey RP, Gutiérrez MS, Baluja C, Sánchez E. Ozone inactivation of microorganisms in water. Gram positive bacteria and yeast. Ozone Sci. Eng. 2001;23(2)183–187. https://doi.org/10.1080/01919510108962001
crossref

37. Lee J, Choi EJ, Rhie K. Validation of algal viability treated with total residual oxidant and organic matter by flow cytometry. Mar. Pollut. Bull. 2015;97(1–2)95–104. https://doi.org/10.1016/j.marpolbul.2015.06.029
crossref pmid

38. Meunpol O, Lopinyosiri K, Menasveta P. The effects of ozone and probiotics on the survival of black tiger shrimp (Penaeus monodon). Aquaculture. 2003;220(1–4)437–448. https://doi.org/10.1016/S0044-8486(02)00586-0
crossref

39. Abbas M, Adil M, Ehtisham-ul-Haque S, et al. Vibrio fischeri bioluminescence inhibition assay for ecotoxicity assessment: A review. Sci. Total Environ. 2018;626:1295–1309. https://doi.org/10.1016/j.scitotenv.2018.01.066
crossref pmid

40. Bayo J, Angosto JM, Gómez-López MD. Ecotoxicological screening of reclaimed disinfected wastewater by Vibrio fischeri bioassay after a chlorination-dechlorination process. J. Hazard Mater. 2009;172(1)166–171. https://doi.org/10.1016/j.jhazmat.2009.06.157
crossref pmid

41. Park C, Cha HG, Lee JH, et al. The effects of chemical additives on the production of disinfection byproducts and ecotoxicity in simulated ballast water. J. Sea Res. 2017;129:80–88. https://doi.org/10.1016/j.seares.2017.07.005
crossref

42. International Organization for Standardization (ISO). ISO-11348 Water quality: Determination of the inhibitory effect of water samples on the light emission of Vibrio fischeri (Luminescent bacteria test). 1998;


43. Liu P, Qian L, Wang H, et al. New insights into the aging behavior of microplastics accelerated by advanced oxidation processes. Environ. Sci. Technol. 2019;53(7)3579–3588. https://doi.org/10.1021/acs.est.9b00493
crossref pmid

44. Hu M, Palic D. Micro- and nano-plastics activation of oxidative and inflammatory adverse outcome pathways. Redox Biol. 2020;37:101620 https://doi.org/https://doi.org/10.1016/j.redox.2020.101620
crossref pmid pmc

45. Zafar R, Park SY, Kim CG. Surface modification of polyethylene microplastic particles during the aqueous-phase ozonation process. Environ. Eng. Res. 2021;26(5)200412 https://doi.org/10.4491/eer.2020.412
crossref

46. Kelkar VP, Rolsky CB, Pant A, Green MD, Tongay S, Halden RU. Chemical and physical changes of microplastics during sterilization by chlorination. Water Res. 2019;163(15)114871 https://doi.org/10.1016/j.watres.2019.114871
crossref

47. Wang Z, Lin T, Chen W. Occurrence and removal of microplastics in an advanced drinking water treatment plant (ADWTP). Sci. Total Environ. 2020;700:134520 https://doi.org/10.1016/j.scitotenv.2019.134520
crossref

48. Wu M, Tang W, Wu S, Liu H, Yang C. Fate and effects of microplastics in wastewater treatment processes. Sci. Total Environ. 2021;757:143902 https://doi.org/10.1016/j.scitotenv.2020.143902
crossref

49. Zhang Z, Chen Y. Effects of microplastics on wastewater and sewage sludge treatment and their removal: A review. Chem Eng. J. 2020;382:122955 https://doi.org/10.1016/j.cej.2019.122955
crossref

50. Ahmed MB, Zhou JL, Ngo HH, Guo W, Thomaidis NS, Xu J. Progress in the biological and chemical treatment technologies for emerging contaminant removal from wastewater: A critical review. J. Hazard. Mater. 2017;323(Part A)274–298. https://doi.org/10.1016/j.jhazmat.2016.04.045
crossref pmid

51. Benner J, Helbling DE, Kohler HPE, et al. Is biological treatment a viable alternative for micropollutant removal in drinking water treatment processes? Water Res. 2013;47(16)5955–5976. https://doi.org/10.1016/j.watres.2013.07.015
crossref pmid

52. Amelia TSM, Khalik WMAWM, Ong MC, Shao YT, Pan HJ, Bhubalan K. Marine microplastics as vectors of major ocean pollutants and its hazards to the marine ecosystem and humans. Prog. Earth Planet. Sci. 2021;8:12 https://doi.org/10.1186/s40645-020-00405-4
crossref

53. Wang Z, Fu D, Gao L, Qi H, Su Y, Peng L. Aged microplastics decrease the bioavailability of coexisting heavy metals to microalga Chlorella vulgaris. Ecotoxicol. Environ. Saf. 2021;217:112199 https://doi.org/10.1016/j.ecoenv.2021.112199
crossref pmid

54. Anbumani S, Kakkar P. Ecotoxicological effects of microplastics on biota: a review. Environ. Sci. Pollut. Res. 2018;25:14373–14396. https://doi.org/10.1007/s11356-018-1999-x
crossref pmid

55. Prokić MD, Radovanović TB, Gavrić JP, Faggio C. Ecotoxicological effects of microplastics: Examination of biomarkers, current state and future perspectives. TrAC - Trends Anal. Chem. 2019;111:37–46. https://doi.org/10.1016/j.trac.2018.12.001
crossref

56. Wang F, Wong CS, Chen D, Lu X, Wang F, Zeng EY. Interaction of toxic chemicals with microplastics: A critical review. Water Res. 2018;139:208–219. https://doi.org/10.1016/j.watres.2018.04.003
crossref pmid

57. Liu G, Zhu Z, Yang Y, Sun Y, Yu F, Ma J. Sorption behavior and mechanism of hydrophilic organic chemicals to virgin and aged microplastics in freshwater and seawater. Environ. Pollut. 2019;246:26–33. https://doi.org/10.1016/j.envpol.2018.11.100
crossref pmid

58. Liu P, Wu X, Liu H, Wang H, Lu K, Gao S. Desorption of pharmaceuticals from pristine and aged polystyrene microplastics under simulated gastrointestinal conditions. J. Hazard Mater. 2020;392:122346 https://doi.org/10.1016/j.jhazmat.2020.122346
crossref pmid

59. Mughini-Gras L, van der Plaats RQJ, van der Wielen PWJJ, Bauerlein PS, de Roda Husman AM. Riverine microplastic and microbial community compositions: A field study in the netherlands. Water Res. 2021;192:116852 https://doi.org/10.1016/j.watres.2021.116852
crossref pmid

60. Vaksmaa A, Knittel K, Abdala Asbun A, et al. Microbial Communities on Plastic Polymers in the Mediterranean Sea. Front. Microbiol. 2021;12:673553 https://doi.org/10.3389/fmicb.2021.673553
crossref pmid pmc

61. Zettler ER, Mincer TJ, Amaral-Zettler LA. Life in the ‘plastisphere’: Microbial communities on plastic marine debris. Environ. Sci. Technol. 2013;47(13)7137–7146. https://doi.org/10.1021/es401288x
crossref pmid

62. Al-Gaashani R, Radiman S, Tabet N, Daud AR. Rapid synthesis and optical properties of hematite (α-Fe2O3) nanostructures using a simple thermal decomposition method. J. Alloys Compd. 2013;550:395–401. https://doi.org/10.1016/j.jallcom.2012.10.150
crossref

63. Gorokhova E, Motiei A, El-Shehawy R. Understanding biofilm formation in ecotoxicological assays with natural and anthropogenic particulates. Front. Microbiol. 2021;12:632947 https://doi.org/10.3389/fmicb.2021.632947
crossref pmid pmc

64. Rummel CD, Jahnke A, Gorokhova E, Kühnel D, Schmitt-Jansen M. Impacts of biofilm formation on the fate and potential effects of microplastic in the aquatic environment. Environ. Sci. Tech Lett. 2017;4(7)258–267. https://doi.org/10.1021/acs.estlett.7b00164
crossref

65. Elfalaky A, Ragheb MS, Zakhary SG. Electron beam induced surface modifications of PET film. Radiat. Phys. Chem. 2014;102:96–102. https://doi.org/10.1016/j.radphyschem.2014.04.025
crossref

66. Gulmine JV, Janissek PR, Heise HM, Akcelrud L. Polyethylene characterization by FTIR. Polym. Test. 2002;21(5)557–563. https://doi.org/10.1016/S0142-9418(01)00124-6
crossref

67. Olmos D, Martín EV, González-Benito J. New molecular-scale information on polystyrene dynamics in PS and PS-BaTiO3 composites from FTIR spectroscopy. Phys. Chem. Chem. Phys. 2014;16:24339–24349. https://doi.org/10.1039/c4cp03516j
crossref pmid

68. Vinodh R, Abidov A, Peng MM, et al. A new strategy to synthesize hypercross-linked conjugated polystyrene and its application towards CO2 sorption. Fibers Polym. 2015;16:1458–1467. https://doi.org/10.1007/s12221-015-5151-y
crossref

69. Fang J, Xuan Y, Li Q. Preparation of polystyrene spheres in different particle sizes and assembly of the PS colloidal crystals. Sci. China Technol. Sci. 2010;53:3088–3093. https://doi.org/10.1007/s11431-010-4110-5
crossref

70. Fang J, Zhang L, Sutton D, Wang X, Lin T. Needleless melt-electrospinning of polypropylene nanofibres. J. Nanomater. 2012;2012:382639 https://doi.org/10.1155/2012/382639
crossref

71. Qi F, Tang M, Wang N, et al. Efficient organic-inorganic intumescent interfacial flame retardants to prepare flame retarded polypropylene with excellent performance. RSC Adv. 2017;7:31696–31706. https://doi.org/10.1039/c7ra04232a
crossref

72. Donelli I, Taddei P, Smet PF, Poelman D, Nierstrasz VA, Freddi G. Enzymatic surface modification and functionalization of PET: A water contact angle, FTIR, and fluorescence spectroscopy study. Biotechnol. Bioeng. 2009;103(5)845–856. https://doi.org/10.1002/bit.22316
crossref pmid

73. Olmez-Hanci T, Arslan-Alaton I, Dursun D. Investigation of the toxicity of common oxidants used in advanced oxidation processes and their quenching agents. J. Hazard. Mater. 2014;278:330–335. https://doi.org/10.1016/j.jhazmat.2014.06.021
crossref pmid

Fig. 1
Comparison of average bacterial population densities, ATP levels and their inactivation rates before and after 10 min of ozone treatment for various ozone doses: (a) E. tarda – cell count, (b) E. tarda – ATP assay, (c) S. mitis – cell count, (d) S. mitis - ATP assay, (e) V. harveyi – cell count, and (f) V. harveyi - ATP assay, respectively.
/upload/thumbnails/eer-2021-612f1.gif
Fig. 2
Comparisons of (a) average bacterial population density and inactivation rates, and (b) ATP levels and inactivation rates between non-treated and ozonated bacterial mixture.
/upload/thumbnails/eer-2021-612f2.gif
Fig. 3
Temporal changes of TROs in artificial seawater after ozone treatment at 0.2, 0.5, 0.7, and 1.0 mg O3/min for 10 min.
/upload/thumbnails/eer-2021-612f3.gif
Fig. 4
Bioluminescence inhibition rates for photobacterium Aliivibrio fischeri at various TROs concentrations (0, 0.5, 1, 1.5, 2, 2.5, 3, 4, 5, 6, 7, and 8 mg Br2/L).
/upload/thumbnails/eer-2021-612f4.gif
Fig. 5
Comparison of EC50 of TROs determined in bioluminescence assay obtained after 15 and 30 min of incubation in bacteria-free seawater containing 0.05, 0.25, and 0.50 g of different microplastic types of PE, PS, PP, and PET.
/upload/thumbnails/eer-2021-612f5.gif
Fig. 6
A comparison of bacterial inactivation rates and the concentration of TROs generated after 10 min of the ozonation toward individuals (a and b: E. tarda; c and d: S. mitis, and e and f: V. harveyi) and mixed pathogens (g and h) according to the presence or absence of microplastics, respectively.
/upload/thumbnails/eer-2021-612f6.gif
Fig. 7
Bioluminescence inhibition rates after ozonation at the customized ozone doses of (a) 3 mg O3/min for E. tarda, (b) 4 mg O3/min for S. mitis, (c) 4 mg O3/min for V. harveyi, and (d) 5 mg O3/min for mixed bacterial pathogens, in seawater containing microplastics of different types (PE, PS, PP, and PET) and weights (0.05, 0.25, and 0.50 g) compared to that without microplastics.
/upload/thumbnails/eer-2021-612f7.gif
TOOLS
PDF Links  PDF Links
PubReader  PubReader
Full text via DOI  Full text via DOI
Download Citation  Download Citation
Supplement  Supplement
  Print
Share:      
METRICS
0
Crossref
0
Scopus
505
View
21
Download
Editorial Office
464 Cheongpa-ro, #726, Jung-gu, Seoul 04510, Republic of Korea
TEL : +82-2-383-9697   FAX : +82-2-383-9654   E-mail : eer@kosenv.or.kr

Copyright© Korean Society of Environmental Engineers.        Developed in M2PI
About |  Browse Articles |  Current Issue |  For Authors and Reviewers