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Impact of salinity on colloidal ozone aphrons in removing phenanthrene from sediments
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Impact of salinity on colloidal ozone aphrons in removing phenanthrene from sediments Ming Zhang, Yudong Feng, Kaihua Zhang, Yafeng Wang, Xiangliang Pan⁎ Key Laboratory of Microbial Technology for Industrial Pollution Control of Zhejiang Province, College of Environment, Zhejiang University of Technology, Hangzhou, 310014, China
G R A P H I C A L A B S T R A C T
A R T I C LE I N FO
A B S T R A C T
Editor: Deyi Hou
Polycyclic aromatic hydrocarbons (PAHs) tend to adsorb and accumulate on sediments owing to their hydrophobicity and persistence. Salinity is the predominant factor determining the PAH partition between aqueous and solid phases in freshwater, estuaries and seawater. This study focuses on the impact of salinity on the phenanthrene (PHE) removal from sediments using an in situ and targeted remediation technology – colloidal ozone aphrons (COAs). The ozone-encapsulated colloidal aphrons exhibited increasing air holdup but decreasing stability with the salinity increasing from 0.5‰ to 35‰. The hydrophobic attraction between Tween-20-coated bubbles and the hydrophobic solid surface weakened at high salinities. The presence of inorganic ions in the aqueous phase could lead to the salting-out of nonionic compounds (PHE, Tween-20 and even ozone), hindering detaching and degrading PHE from the solid phase. Anyhow, COAs achieved high efficiencies of washing (88.0–90.2%) and oxidative degradation (74.0–76.5%) particularly for the hydrophobic sediments with highly concentrated PHE (200.4 μg/kg) over the investigated salinities. The flushing effect imposed by the bubble flow played an important role, which was not greatly influenced by salinity. Although the dissolved natural organic matter competed with PHE for COAs and led to low PHE removal, the efficiency was improved by successive COA addition.
Keywords: Colloidal ozone aphrons Functionalized microbubbles Salinity Polycyclic aromatic hydrocarbons Sediments
⁎
Corresponding author. E-mail address: [email protected] (X. Pan).
https://doi.org/10.1016/j.jhazmat.2019.121436 Received 27 April 2019; Received in revised form 30 September 2019; Accepted 8 October 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.
Please cite this article as: Ming Zhang, et al., Journal of Hazardous Materials, https://doi.org/10.1016/j.jhazmat.2019.121436
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1. Introduction
changes the stability of CGAs (Zhang et al., 2019b). As a novel remediation technology, COAs have not been explored in terms of salinity impact on the PAH removal to the authors’ knowledge. The objective of this work is to investigate the influence of the important variable (salinity) on the properties of COAs as well as the removal of PAHs from sediments. Phenanthrene (PHE), one of the USEPA priority PAHs (Li et al., 2016), was chosen as representative PAH herein; and the typical nonionic surfactant, polyoxyethylene sorbitol anhydride monolaurate (Tween-20), was used for the COA generation. Different levels of salinity, simulating that in the aquatic environments of freshwater, estuaries and marine water, were adopted in all the experiments. The properties of COAs were characterized in terms of stability, air content, size and surface adherence. Desorption kinetics of PAHs from sediments was then studied for the target capture of PHE by the Tween-20-coated bubbles. Furthermore, the efficiency of PHE removal from sediments was explored using the COAs generated from mineral water and river water, respectively, and the operation was optimized for good performance.
Polycyclic aromatic hydrocarbons (PAHs) are ubiquitously detected in the aquatic environment. These fat-soluble and degradation-resistant organic pollutants accumulate and persist in the solid phase of water. Take the lakes in China as an instance, the concentration of ∑PAHs in sediments ranges from 6.52 μg/kg to 7935.21 μg/kg (Meng et al., 2019). Great concerns have been raised for their teratogenic, carcinogenic and mutagenic properties to living bodies, and the U.S. Environmental Protection Agency (USEPA) has listed them as priority control pollutants (Gong et al., 2018). Physic-chemical technologies are promising in the isolation, separation and even elimination of PAHs from sediments if they combine the technical advantages of sediment washing/solvent extraction and oxidative degradation. Previous studies proposed the concept of increasing the utilization of ozone within the subsurface, by which the ozone micro-nano-bubbles were generated for the site remediation of groundwater contamination (Li et al., 2014; Xia and Hu, 2019). Such bubbles are able to transport with water flow, present rapid mass transfer rates, and provide continuous ozone supply (Hu and Xia, 2018). As for the remediation of sediment contamination, more robust and persistent bubble systems are required for the in situ and targeted removal of contaminants. Noting that the modifiable microbubble system of colloidal gas aprhons (CGAs), being created from the surfactant solution, is of striking features, including large specific bubble surface area, high stability and air hold-up, strong transport capability in a pressurized environment, as well as controllable core and surface properties (Zhang and Guiraud, 2017; Zhang et al., 2018). CGAs are of great potential in contamination remediation. Given that PAHs are of strong hydrophobicity and affinitive adhesion on the surface or inside the pores of sediments (Gitipour et al., 2018), the innovative attempt can be constructing CGAs with nonionic surfactant molecules and encapsulating the gaseous oxidant – ozone into the gas core of bubbles. The so-created bubbles, being named as colloidal ozone aphrons (COAs), are expected to first adsorb on the PAH-contaminated sediments, detach those target hydrophobic contaminants from the sediment surface, and then oxidatively degrade them with the released ozone (Zhang et al., 2019a). Much of work needs to be done to examine and evaluate the COA performance; in particular, the influence of aquatic chemistry on the COA treatment process should be specifically investigated. It has been found that the hydrophobic contaminants present distinct sorption and desorption behaviors for the sediments in marine water, estuaries and freshwater, and the important difference among the abovementioned aquatic environments is salinity (Oh et al., 2013). Therefore, the study of salinity or ionic strength on the COA-remediation of PAH-contaminated sediments is crucial for process interpretation and field applications. On the one hand, the salinity of the surfactant solution may affect the solubility of surfactant molecules because the complex compounds between ions and surfactants can alter micelle formation (Abouseoud et al., 2010; Zhao et al., 2015). Since surfactant has been reported to increase the dissolution of ozone (Ji et al., 2018), the change of surfactant solubility by salinity then probably affects the ozonation during the treatment with COAs. Consequently, the COAs being generated from the surfactant solution may exhibit distinct properties and interfacial behaviors at different levels of salinity. On the other hand, it has been reported that the competitive adsorption between the anion of Cl− and the organic contaminants on the solid surface can make an important influence on the contaminantinvolved chemical reactions (Fu et al., 2019; Qin et al., 2019). Thus, through changing the partition of chemicals in solid and aqueous phases, salinity affects the environmental fate and exposure of chemicals in aquatic environment (Saranjampour et al., 2017; Jonker et al., 2015). Obviously, the interaction between PAH molecules and sediments, such as sorption, desorption and even aging, may vary with salinity. Moreover, it is known that high ionic strength significantly
2. Materials and methods 2.1. Chemicals PHE (C14H10, purity ≥99%) standards for gas chromatography analysis were provided by Aladdin, China. Tween-20 (C58H114O26, 1227.5 g/mol, chemical pure), anhydrous ethanol (analytical reagent, AR) and sodium chloride (NaCl, AR), were purchased from Sinopharm Chemical Reagent Co., Ltd, China. Dichlorodimethylsliane (DCDMS, purity ≥99%) was obtained from Adamas-beta, Switzerland. Acetonitrile (purity ≥99%) and hexane (purity ≥95%) standards for high performance liquid chromatography (HPLC) analysis were provided by Sigma-Aldrich, U.S.A.. All those reagents were used as purchased without further purification. All the solutions for the HPLC analysis were prepared using ultrapure water from a ultrapure water system (Smart-N15VF, Healforce, China) with resistivity of 18.2 MΩ cm. 2.2. Preparation of water samples with different levels of salinity The artificial water was prepared using mineral water (Nongfu Spring Co. Ltd, Hangzhou, China). The pristine properties of water, including dissolved organic carbon (DOC), pH and ion concentrations, were analyzed by total organic carbon analyzer (TOC-L, Shimadzu, Japan), pH meter (Primainst, innoLab 20 P, U.K.) and high performance ion chromatograph (Dionex™ Integrion™ HPIC™, ThermoFisher Scientific Co. Ltd., U.S.A.) and were shown in Table 1. Three levels of salinity (0.5, 10 and 35‰, g/L) were controlled with NaCl, and the corresponding conductivity values were measured to be 0.86, 16.29 and 53.16 mS/cm, respectively, by conductivity meter (DDS-307, INESA Scientific Instrument Co., Ltd., China). The water samples were then adopted in the experimental trials, including COA generation, PHE Table 1 Salient properties of pristine mineral water and river water before salinity adjustment.
2
Parameters
Mineral spring
River water
pH DOC (mg/L) Na+ (mg/L) K+ (mg/L) Mg2+ (mg/L) Ca2+ (mg/L) NH4+ (mg/L) Cl− (mg/L) NO3− (mg/L) SO42− (mg/L)
6.50 ± 0.06 0.08 ± 0.00 5.60 ± 0.01 1.27 ± 0.06 2.44 ± 0.01 20.07 ± 0.37 0.27 ± 0.00 4.28 ± 0.11 4.28 ± 0.02 7.50 ± 0.02
7.34 ± 0.01 1.17 ± 0.01 25.45 ± 0.05 4.44 ± 0.05 4.66 ± 0.02 24.19 ± 5.48 0.20 ± 0.06 23.04 ± 0.54 12.48 ± 0.11 30.91 ± 0.35
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The desorption experiments were conducted in the 500 m L-glass vials with Teflon-lined septa and screw caps at 25 °C. Those pre-contaminated sediments (0.06 kg) were added into the vial which contained 0.3 L of Tween-20-CGA suspension making from the mineral water with different levels of salinity; then, the mixture was shaken for 120 min a rotary agitator at the speed of 150 rpm. Five milliliter of water sample was taken at 10, 30, 60, 90 and 120 min, respectively. The PHE concentration on sediments was then calculated as follows:
desorption and contaminated sediment remediation. The river water was also employed for the generation of COAs which were then used in the experiments for remediating the PHEcontaminated sediments. The river water was collected from the top 30 cm of water column from Shangtang River in Hangzhou, China (gerographic coordinates: 30°17′37″ N, 12°09′44″E). Before use, the sample was passed through 0.45-μm membrane filters to remove the suspended solids and then was sterilized at 121 °C for 35 min via autoclaving. The salinity of river water, being adjusted with mineral salt (YEE® aquarium experts, China), was controlled to be close to that of the mineral water mentioned above: the conductivity was measured to be 0.23, 18.16 and 43.04 mS/cm, individually.
C (PHE )se dim ents, t C (PHE )se dim ents,0 × mse dim ents − C (PHE ) water , t × Vwater , t = mse dim ents
(1)
where, C (PHE )se dim ents, t (μg/kg) and C (PHE ) water , t (μg/L) represent the PHE concentration on sediments and in water phase, individually, at time t; C (PHE )se dim ents,0 (μg/kg) represents the initial PHE concentration on sediments; mse dim ents (kg) is the mass of sediments and has been fixed to 0.06 kg herein; Vwater , t (L) equals to the volume of water at time t which should be corrected from the water loss due to sampling:
2.3. Generation and characterization of COAs The apparatus for continuous ozone production-microbubble generation (HG-WNF-1, Hangzhou Guiguan Company, China) was used for the COA generation: ozone was first produced from air by ozonator; then, the ozone-contained air and the bubble generation solution (surfactant solution) were pumped into the saturator at the flow rate of 2.25 L/min and 360 mL/min, separately. COAs were created by pressuring dissolution of ozone-contained air at 300 kPa into water with different levels of salinity, which was the typical method of generating CGAs (Hashim et al., 2012). The nonionic surfactant, Tween-20, was chosen for bubble production to ensure the hydrophobic attraction between PHE and COAs. The concentration of Tween-20 in the bubble generation solution was determined to approximately 122.8 mg/L, higher than the critical micelle concentration (CMC) of Tween-20 (89.2 mg/L). The dissolved ozone concentration was maintained to be 0.6 mg/L which was measured with ozone analyzer (PC-II, Hach, U.S.A.). As comparison, the ozone microbubbles without surfactant and the Tween-20-CGAs without ozone were generated, individually, for the removal of PHE from sediments. After the COA generator ran for 7 min, the COA suspensions were obtained for bubble characterization or PHE removal. The COAs were analyzed in terms of stability (half-life time), air holdup and size, and the detailed measurement procedures have been elaborated in previous publications (Zhang and Guiraud, 2017; Zhang et al., 2018). The adhesion of colloidal aphrons (Tween-20-CGAs) onto the surface with different degrees of hydrophobicity was tested. One side of the glass slide (Sail Brand, China) was surface-modified with the PHE stock solution at the pre-determined PHE concentrations to reach the specific levels of hydrophobicity. The PHE concentration on the surfacetreated glass slides was measured by HPLC (LC-20AT, Shimadzu, Japan). DCDMS was also used to obtain the superhydrophobic surface. The detailed procedure of surface treatment is given in Section 1 of Supporting Information (SI). The glass slides were put in a transparent glass tank (15 × 10 × 13 cm3), and the untreated side was attached firmly onto the inner wall of the glass tank while the other side was immersed into the newly generated suspension of COAs. The images of the bubble-attached glass slides were taken at 0, 1, 2, 3, 5, 7 and 24 h with optical microscope equipped with a digital camera (IX71, Olympus, Japan). The number of bubbles per 25 mm2 on slide was counted, which was repeated at least three times for each slide. Each batch was made in triplicate.
Vwater , t = Vwater ,0 − T × ΔV
(2)
where, Vwater,0 (mL) and ΔV (mL) are the initial volume of water and the sample volume for each sampling, individually, which have been fixed to 0.3 L and 0.005 L, respectively. T represents the times of sampling at time t. 2.5. COA treatment of PHE-contaminated sediments The experiments of contaminated sediment remediation were carried out in semi-batch mode. The generated bubble suspension was pumped out at the flow rate of 252 mL/min by peristaltic pump (BT3002 J, Longer Pump, China) to the reactor for the remediation of PHEcontaminated sediments; the rest of the bubble suspension was recycled for bubble generation through a three-way valve. The COA suspension (0.1 L) was pumped into the glass reactor from the bottom and was used for remediating 0.02 kg of PHE-contaminated sediments. The reactor was sealed during the treatment for 2 h, after which the liquid and sediments were collected, respectively, to measure the residual PHE concentration. The chemical extraction process and HPLC analysis of PHE can be found in Section S3 of SI. Both of the mineral water and the river water (described in Section 2.2) with different salinities were used to explore the removal of PHE from sediments by COAs in this part of study. The efficiencies of sediment washing and PHE oxidation degradation indicated the capacity of COAs in detaching and removing PHE from the sediment matrix, respectively. The efficiencies could be defined and calculated as Eqs. (3) and (4):
PHE desorption efficiency (%) C (PHE )se dim ents,0 − C (PHE )se dim ents, t = × 100% C (PHE )se dim ents,0
(3)
where, C(PHE)sediment,0 (μg/kg) and C(PHE)sediments,t (μg/kg) represent the amount of PHE adsorbed on the sandy sediments before and after the COA treatment, individually.
2.4. Desorption kinetics of PHE from sediments
PHE oxidation degradation efficiency
(%)
(C (PHE )se dim ents,0 − C (PHE )se dim ents, t ) × mse dim ents
The model sandy sediments (6–9 mm, Helei, China), free from any contamination, were selected for the experiments of PHE desorption kinetics and contaminated sediment remediation. The artificial contamination of PHE followed the method in Section S2 of SI. The sandy sediments with and without hydrophobic surface-treatment were spiked with the PHE stock solution to obtain different homogeneous contamination levels. The characteristics of the artificial PHE-contaminated sandy sediments are shown in Table S1.
=
− C (PHE ) water , t × Vwater × 100% (C (PHE )se dim ents,0 − C (PHE )se dim ents, t ) × mse dim ents
(4)
where, C(PHE)water,t (mg/L) represents the PHE concentration in liquid after treatment; msediments (g) and Vwater (mL) are the mass of sandy sediments (0.02 kg) and the volume of liquid being collected after the treatment for 2 h. 3
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Fig. 1. Characteristics of COAs generated from Tween-20 solutions with different levels of salinity: (a) half-life time, (b) air hold-up, and (c) size distribution.
3. Results and discussion
greatly but the air holdup increased with the increase of salinity. Barrut et al. found in their study that salinity imposed significant impact on gas holdup (Barrut et al., 2012). Compared with the scenario without electrolytes, the hydrophilic repulsive force between bubbles might rise at the high salinity, particularly in the marine water, which could lead to the inhibition of Laplace pressure. Laplace pressure is known to cause the gas diffusion from small bubbles to larger ones due to the pressure difference (Ruen-ngam et al., 2008; Bera et al., 2013). As a result, the air content of the bubble system increased with the salinity increasing from 0.5‰ to 35‰ in this study.
The PHE removal by COAs depends on the properties of COAs, the PHE desorption from sediment by the hydrophobic surface of COAs, and the PHE oxidation by the released ozone from COAs. In this study, the salinity of freshwater (0.5‰), estuaries (10‰) and marine water (35‰) was specifically investigated for its impact on all the above mentioned processes. This helps to clarify and interpret the characteristic COA behaviors during the remediation of PHE-contaminated sediments in different aquatic environments. 3.1. Characteristic COA properties
3.2. Target adsorption of Tween-20-CGAs on hydrophobic surfaces
The most significant properties of colloidal aphrons include half-life time, air holdup and average size (Molaei and Waters, 2015), which indicate stability, bubble density and specific surface area of the bubble system. Experimental results denote that, with the salinity of Tween-20 solution increasing from 0.5‰ to 35‰, the half-life time of COAs dropped from 221 s to 163 s (Fig. 1(a)), the air holdup increased from 19.2% to 30.8% (Fig. 1(b)), and the average size of bubble fluctuated between 130–160 μm (Fig. 1(c)). The addition of inorganic electrolytes, such as NaCl, created a high concentration of ions in the vicinity of the micelle surface and thus favored the micelle formation. In this regard, the CMC tended to be reduced due to the presence of salt (PérezGramatges et al., 2013). Particularly, at the salinity close to that in the marine water (35‰), NaCl might play the role of salting-out agent for the non-electrolyte (Pérez-Gramatges et al., 2013; Salabat and Alinoori, 2008) which is Tween-20 in this study. In consequence, the concentration of effective surfactant molecules was reduced; and accordingly, the foamability of bubble generation solution and the stability of bubbles decreased (Wang and Li, 2016). This agreed well with what was found elsewhere (Feng et al., 2009; Luo et al., 2009). It has been reported that the bubble size is closely associated with the variation of air holdup (Bournival et al., 2014), which was, however, not consistent with the results in this work. Herein, the bubble size did not change
The COAs generated at the three levels of salinity adsorbed on the surface-modified glass slide, and the bubble density on the slide is shown in Fig. 2. For the DCDMS-treated slide which was of the highest hydrophobicity, the initial CGA density reduced from 246 bubble/ 25 mm2 to 101 bubble/25 mm2 with the salinity increasing from 0.5‰ to 35‰. Under the condition of high ionic strength, the hydrophobic attraction between the slide surface and the CGA surface with Tween20 molecules got weakened, which played the essential role in the bubble adsorption. For the CGA suspension at the fixed salinity, the bubble density decreased on the slide surface in the order of DCDMS-, PHE-modified- and untreated-slide. The difference of bubble density was found to be quite small between the high salinities of 10‰ and 35‰. At such high ion concentrations, the adsorption of the nonionic surfactant-coated bubbles onto the slide surface was hard to be further increased. The adsorbed Tween-20-CGAs drained with time, and thus, the bubble density was less than 30 bubble/25 mm2 at 24 h for all the three levels of salinity. The bubble density reduced faster with time at the lowest salinity of 0.5‰ than that at the other two higher salinities, which could be ascribed to the faster movement of Tween-20 surfactant molecules at the lower ion concentration.
4
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Fig. 2. Density of Tween-20-CGAs varying with time on hydrophobically-treated surfaces at the salinity of (a) 0.5‰, (b) 10‰ and (c) 35‰.
aggregates, which resulted in difficulty in desorption diffusion (Li et al., 2017). The comparison of desorption kinetics with (Fig. 3(a)–(c)) and without (Fig. 3(d)–(f)) Tween-20-CGAs reveals that the colloidal bubbles were apt to capture and detach the PHE molecules from the hydrophobic surface. Apparently, the addition of Tween-20-CGAs facilitated the release of PHE at different levels of salinity. At the high PHE concentration on sediment of 200.4 μg/kg (Fig. 3(a)), with the increasing addition of NaCl, the cation of Na+ adsorbed to the molecules of Tween-20, particularly, at the oxygen in
3.3. Desorption kinetics of PHE from hydrophobic surfaces using Tween-20CGAs In addition to the bubble properties, the role of salinity was also explored with respect to the PHE desorption from sediments with the assistance of Tween-20-coated colloidal aphrons. Previous study has pointed out that the peculiar molecular structure of PHE, three rings sequentially connected in a non-straight shape, could be intercepted by pores as well as twined with other PHE molecules forming larger
Fig. 3. Desorption kinetics of PHE from aquifer sediments in the presence ((a), (b) and (c)) and absence ((d), (e) and (f)) of Tween-20-CGAs at the salinity of 0.5‰, 10‰ and 35‰. The initial PHE concentrations on sediments were approximately 200.4 μg/kg ((a) and (d)), 55.8 μg/kg ((b) and (e)) and 30.0 μg/kg ((d) and (f)), individually. 5
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water at different levels of salinity, and the corresponding treatment effect can be found in Fig. 4(a)–(d), respectively. For the COA-involved experiments, the specific salinity gradient between 0.5‰ and 35‰ was tested to verify the trend of PHE removal varying with salinities (Fig. S1). The PHE desorption efficiency by COAs kept stable in 88.0–90.2% in the tested salinity range of 0.5-35‰ when PHE adsorbed on the hydrophobic sediment surface at the high concentration of 200.4 μg/kg. Compared with the PHE molecules in the inner PHE layer which firmly adhered to the surface and pores of sediments, the outer PHE might loosely adsorb and be easily washed out by COAs. Thus, it can be inferred that the desorption of highly concentrated PHE from sediments was mainly ascribed to the flow flushing caused by the bubble suspension as well as the COA capture due to the hydrophobic attraction between PHE and Tween-20. The former mechanism may be predominant and not greatly impacted by salt concentration. For the sediments with low concentration of PHE – 55.8 μg/kg for the hydrophobic surface and 24.5 μg/kg for the hydrophilic surface, the increased salinity of the COA suspension slightly decreased the desorption of PHE: with the salinity increasing from 0.5‰ to 35‰, the desorption for the former and latter scenario reduced from 68.1% to 61.7% and from 75.3% to 69.3%, respectively. Most of those hydrophobic organic molecules should be trapped by the porous structure of sediments and/or by the hydrophobic attachment with the hydrophobic sediment surface. This would conspicuously increase the difficulty for COAs to capture PHE from sediment in spite of the hydrophobic attraction of hydrophobic molecules between Tween-20 and PHE. The flushing effect which was caused by the bubble suspension flow imposed little effect at such a low PHE concentration in contrast with that at the high PHE concentration as mentioned above. Moreover, the presence of salt, particularly at high concentrations, could impede the desorption process as explained in Section 3.3. As shown in Fig. 4(b), the oxidative degradation of PHE did not vary significantly with the salinity increasing from 0.5‰ to 35‰ for the
the ethylene oxide groups. This might change the configuration through the coordination of the oxygen atoms to the Na+ in a tetragonal-pyramidal manner and form some ordered structure like microcages, which would interact preferably with the planar molecules like PHE (Li and Chen, 2002). But it should also be highlighted that the salting-out effect caused by the too high concentrations of inorganic salt (35‰), typically in of marine water, could result in increasing chemical distribution from aqueous to solid phase by reducing the solubility of target hydrophobic contaminant (such as PHE) and the nonionic surfactant molecules (Tween-20) (Wu and Sun, 2010). This was directly reflected by the results shown in Fig. 3(b) and (c). The residual PHE concentration on sediment first decreased and then increased with time for the hydrophobically (Fig. 3(a), (b), (d) and (e)) and hydrophilically (Fig. 3(c) and (f)) surface-treated sediments with different initial PHE concentrations. Hence, the desorption kinetics within the investigated 120 min could be divided into two stages – desorption and re-adsorption. This phenomenon was conspicuous for the desorption process with the assistant of Tween-20-CGAs (Fig. 3(b) and (c)). The release of PHE is kinetically controlled by the desorption itself and the diffusion through sorbent to water (Barnier et al., 2014). The latter mechanism was strongly influenced by colloidal aphrons. After the Tween-20-CGAs collapsed, the free molecules of the nonionic surfactant might adsorb onto the surface of sediments, resulting in the re-adsorption of PHE. In that case, the diffusion of PHE into the aquatic phase would be impeded.
3.4. PHE removal from sediments using COAs 3.4.1. Comparison between COAs and ozone microbubbles The performance of the ozone-encapsulated colloidal aphrons in in situ and targetedly removing PHE from sediments was studied in terms of desorption efficiency and oxidative degradation efficiency within 2 h. The COAs and ozone microbubbles were generated from the mineral
Fig. 4. Comparison of PHE removal between COAs ((a) and (b)) and ozone microbubbles ((c) and (d)): (a) and (c) desorption efficiency, (b) and (d) oxidative degradation efficiency. 6
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Fig. 5. (a) Desorption and (b) oxidative degradation of PHE from real water-based aquifer sediments.
Fig. 6. Improvement of treatment efficiency by adding COAs twice: (a) desorption and (b) oxidative degradation.
PHE from sediments. Nevertheless, as much as ∼73.8% of PHE desorbed from sediments by ozone bubbles, which was not influenced by salinity. This confirmed the speculation that, at the high PHE concentration, the flow flushing (or, the washing effect) contributed greatly to the desorption process. When the hydrophobic or hydrophilic sediment surface was loaded with smaller amount of PHE (55.8 μg/kg and 24.5 μg/kg, individually), the PHE removal by ozone microbubbles was higher than that induced by COAs. For the ozone microbubble-treated sediments, the addition of NaCl led to the increase of surface tension, internal pressure of bubbles, and high mass transfer efficiency of ozone in the absence of nonionic surfactant (Xia and Hu, 2019). As for the COA-involved treatment, the high amount of ions in aqueous phase resulted in the salting-out of the nonionic compounds (El-Nahhal and Safi, 2004) – PHE and Tween-20 as well as the encapsulated ozone, and hence, the nonionic compounds were apt to adsorb on, rather than desorb from, the solid phase. Obviously, when the ozone dosage was fixed and the salinity was controlled below 10‰, COAs performed better on the sediments with heavy PHE contamination (i.e. 200 μg/kg) whereas the ozone microbubbles took greater advantages in removing PHE from sediments with relatively low PHE concentration.
sediments with high and low PHE concentrations. Salinity was found to affect ozone oxidation of hydrophobic organic components in some contrasting aspects (Ji et al., 2018): negatively, (1) NaCl may compete for dissolved ozone and radicals, and (2) the raised ionic strength pushes more PAHs to the Tween-20/water cages; positively, (1) the nonionic surfactant of Tween-20 may entrain more PHE to the top layer of the water, which is conducive to the contact and reactions with gaseous ozone in the headspace, and (2) the elevated ionic strength probably increases the interfacial concentration of the hydrophobic hydrocarbons of PHE at the ozone-solution interface due to the “salting-out” effect; but this effect can be partially offset because of the lowered surface tension and the increased solubilization of PHE. Besides, it is noted that the oxidative degradation of PHE was 74.0–76.5%, 54.0–56.7% and 33.8–36.1% for the sediments with high and low concentrations of PHE, individually. During ozonation, the direct attack by ozone molecules is the predominant mechanism of PHE oxidation (Gong and Zhao, 2017). As for the COA treatment process, the PHE molecules desorbed from the sediments would preferably attach on the COA surface due to the hydrophobic attraction between the molecules of PHE and Tween-20. Then, those PHE molecules could be directly attacked by the ozone molecules released from the drained COAs. Therefore, the efficiency of PHE oxidation degradation by COAs was consistent with that of desorption (shown in Fig. 4(a)), particularly for the PHE adsorbed on the hydrophobic sediment surface. In comparison of the treatment with COAs, the ozone microbubbles, being generated in the absence of surfactant, were adopted to remove PHE from sediments (see in Fig. 4(c) and (d)). For the hydrophobically surface-treated sediments with the PHE concentration of 200.4 μg/kg, the desorption and oxidative degradation efficiencies induced by ozone microbubbles were around 12–15% and 4–10%, respectively, being lower than those caused by COAs. In contrast to the bubbles without surfactant coating, the Tween-20-layer of COAs facilitated the release of
3.4.2. River water-based COAs for PHE removal To further investigate the COA performance under natural aqueous conditions, the river water, rather than mineral water, was employed for bubble generation as described in Section 2.2. By comparing the results in Fig. 5(a) and Fig. 4(a), it can be found out that the PHE desorption efficiency decreased by 6–20% in the river water-involved experimental trials no matter that the PHE concentration was high or low and that the sediment surface was hydrophobic or hydrophilic. As presented in Table 1, the DOC in river water was approximately 14 times higher than that in mineral water; thereby, the most significant difference between the two waters was the content of dissolved natural 7
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DNOM. For the improvement of remediation effect, the successive treatment of sediments (totally, twice) was adopted and in excess of 90.0% and 89.0% of desorption and oxidative degradation efficiency was achieved. The study sheds light on adapting the operation conditions of COA technology to specific salinity environments for the effective remediation of PAH-contaminated sediments.
organic matter (DNOM). As a diverse and complex mixture with a wide range of molecular weight and chemical structures, DNOM has multiple interaction with organic pollutants, affecting their distribution in waters with varied salinities (Lou et al., 2012; Yuan et al., 2017). The addition of mineral salt prevented the DNOM (particularly that with hydrophobic moieties) from releasing from sediments. This would further enhance the adsorption of highly concentrated PHE molecules on sediments through the hydrophobic attraction between PHE and DNOM. It is worth noting that, even though the salinity influenced the desorption process induced by the COAs generated from river water, the impact was different for the hydrophobic and hydrophilic sediment surface. Higher salinities (10‰ and 35‰) tended to hinder the PHE release from the hydrophobic surface whilst the opposite result was obtained for the hydrophilic surface. The PHE molecules were supposed to loosely adsorb on the hydrophilic sediment surface, and the partition of PHE in the aqueous phase could be enhanced at high ionic strength. The results of oxidative degradation in Figs. 4(b) and 5 (b) indicate the drop of efficiency for the river water-involved tests. This could be explained by the consumption of ozone by DNOM. In particular, for the hydrophobically surface-treated sediments, the increase of salinity gave rise to the decrease of oxidative degradation efficiency. The Tween-20 molecules might not only salt out but also be entrained by the undissolved DNOM, which negatively affected the encapsulation of ozone by the surfactant molecules during the formation of COAs. Additionally, in excess of 90.0% and 89.0% of desorption and oxidative degradation efficiencies could be achieved via the successive treatment of sediments with the same volume of COA suspension (0.1 L for each time, and totally, twice) as revealed in Fig. 6. On the one hand, the adsorption of recalcitrant PHE on sediment could get weakened by the first added COAs; and thus, it would be easier for the later added COAs to detach and oxidize the PHE molecules. On the other hand, the desorbed but not oxidized PHE molecules in the first treatment would be removed with ozone during the following COA treatment. Noting that the PHE removal was still negatively affect by high salinity in the twice COA treatment.
Acknowledgements This research was supported by the National Science Foundation of China (No. 51608373, No. U1703243 and No. U1503281), the Zhejiang Provincial Natural Science Foundation of China (LY19E080018), and the Scientific Starting Foundation of Zhejiang University of Technology (2017129008229). Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jhazmat.2019.121436. References Abouseoud, M., Yataghene, A., Amrane, A., Maachi, R., 2010. Effect of pH and salinity on the emulsifying capacity and naphthalene solubility of a biosurfactant produced by Pseudomonas fluorescens. J. Hazard. Mater. 180, 131–136. https://doi.org/10.1016/ j.jhazmat.2010.04.003. Barnier, C., Ouvrard, S., Robin, C., Morel, J.L., 2014. Desorption kinetics of PAHs from aged industrial soils for availability assessment. Sci. Total Environ. 470–471, 639–645. https://doi.org/10.1016/j.scitotenv.2013.10.032. Barrut, B., Blancheton, J.-P., Champagne, J.-Y., Grasmick, A., 2012. Mass transfer efficiency of a vacuum airlift-application to water recycling in aquaculture systems. Aquacult. Eng. 46, 18–26. https://doi.org/10.1016/j.aquaeng.2011.10.004. Bera, A., Ojha, K., Mandal, A., 2013. Synergistic effect of mixed surfactant systems on foam behavior and surface tension. J. Surfactants Deterg. 16, 621–630. https://doi. org/10.1007/s11743-012-1422-4. Bournival, G., Du, Z., Ata, S., Jameson, G.J., 2014. Foaming and gas dispersion properties of non-ionic surfactants in the presence of an inorganic electrolyte. Chem. Eng. Sci. 116, 536–546. https://doi.org/10.1016/j.ces.2014.05.011. El-Nahhal, Y.Z., Safi, J.M., 2004. Adsorption of phenanthrene on organoclays from distilled and saline water. J. Colloid Interface Sci. 269, 265–273. https://doi.org/10. 1016/S0021-9797(03)00607-6. Feng, W., Singhal, N., Swift, S., 2009. Drainage mechanism of microbubble dispersion and factors influencing its stability. J. Colloid Interface Sci. 337, 548–554. https://doi. org/10.1016/j.jcis.2009.05.054. Fu, Y., Qin, L., Huang, D., Zeng, G., Lai, C., Li, B., He, J., Yi, H., Zhang, M., Cheng, M., Wen, X., 2019. Chitosan functionalized activated coke for Au nanoparticles anchoring: green synthesis and catalytic activities in hydrogenation of nitrophenols and azo dyes. Appl. Catal. B 255, 117740. https://doi.org/10.1016/j.apcatb.2019.05. 042. Gitipour, S., Sorial, G.A., Ghasemi, S., Bazyari, M., 2018. Treatment technologies for PAHcontaminated sites: a critical review. Environ. Monit. Assess. 190, 546. https://doi. org/10.1007/s10661-018-6936-4. Gong, X., Xiao, L., Zhao, Z., Li, Q., Feng, F., Zhang, L., Deng, Z., 2018. Spatial variation of polycyclic aromatic hydrocarbons (PAHs) in surface sediments from rivers in hilly regions of Southern China in the wet and dry seasons. Ecotoxicol. Environ. Saf. 156, 322–329. https://doi.org/10.1016/j.ecoenv.2018.03.004. Gong, Y., Zhao, D., 2017. Effects of oil dispersant on ozone oxidation of phenanthrene and pyrene in marine water. Chemosphere 172, 468–475. https://doi.org/10.1016/j. chemosphere.2017.01.007. Hashim, Mohd. A., Mukhopadhyay, S., Gupta, B.S., Sahu, J.N., 2012. Application of colloidal gas aphrons for pollution remediation. J. Chem. Technol. Biotechnol. 87, 305–324. https://doi.org/10.1002/jctb.3691. Hu, L., Xia, Z., 2018. Application of ozone micro-nano-bubbles to groundwater remediation. J. Hazard. Mater. 342, 446–453. https://doi.org/10.1016/j.jhazmat. 2017.08.030. Ji, H., Gong, Y., Duan, J., Zhao, D., Liu, W., 2018. Degradation of petroleum hydrocarbons in seawater by simulated surface-level atmospheric ozone: reaction kinetics and effect of oil dispersant. Mar. Pollut. Bull. 135, 427–440. https://doi.org/10. 1016/j.marpolbul.2018.07.047. Jonker, M.T.O., van der Heijden, S.A., Kotte, M., Smedes, F., 2015. Quantifying the effects of temperature and salinity on partitioning of hydrophobic organic chemicals to silicone rubber passive samplers. Environ. Sci. Technol. 49, 6791–6799. https://doi. org/10.1021/acs.est.5b00286. Li, J.-L., Chen, B.-H., 2002. Solubilization of model polycyclic aromatic hydrocarbons by nonionic surfactants. Chem. Eng. Sci. 57, 2825–2835. https://doi.org/10.1016/ S0009-2509(02)00169-0. Li, H., Hu, L., Song, D., Lin, F., 2014. Characteristics of micro-nano bubbles and potential application in groundwater bioremediation. Water Environ. Res. 86, 844–851. https://doi.org/10.2175/106143014X14062131177953.
4. Conclusions As a distinct feature of freshwater, estuaries and marine water, salinity was explored for its impact on the in situ and targeted removal of PHE from sediments by the COA system in this work. With the increase of salinity from 0.5‰ to 35‰, the stability of COAs decreased, the air holdup increased, and the bubble size was almost unaffected. At high salinities, the salting-out effect reduced the solubility of Tween-20, which accelerated the drainage of COAs; meanwhile, the raised interfacial and surface tension gave rise to the increased foamability and thus the improved air content. Those Tween20-coated COAs preferably adsorbed on the hydrophobic surface; but the most essential contributor – the hydrophobic attraction weakened at high ionic strength. Compared with the ozone microbubbles, the COAs showed better performance in desorption (88.0–90.2%) and oxidative degradation (74.0–76.5%), particularly for the hydrophobic sediment surface with the high PHE concentration of 200.4 μg/kg. The effect remained stable over the investigated salinity range. In this scenario, the PHE molecules in the outer PHE layer might loosely adsorb on the sediment surface and easily desorb due to both of the bubble flushing flow and the COA capture. The former effect was not greatly impacted by salinity. As for the desorption of lowly concentrated PHE, the efficiency was not high particularly at high salinities. Herein, the salting-out effect upon PHE played the dominant role in impeding the release of PHE. When DNOM was presented in the system, it could consume the COAs and lead to low PHE removal. The addition of mineral salt inhibited the DNOM (particularly that with hydrophobic moieties) from releasing from sediments, which further enhanced the adsorption of highly concentrated PHE on sediment through the hydrophobic attraction between PHE and 8
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Salabat, A., Alinoori, M., 2008. Salt effect on aqueous two-phase system composed of nonylphenyl ethoxylate non-ionic surfactant. Calphad 32, 611–614. https://doi.org/ 10.1016/j.calphad.2008.06.004. Saranjampour, P., Vebrosky, E.N., Armbrust, K.L., 2017. Salinity impacts on water solubility and n-octanol/water partition coefficients of selected pesticides and oil constituents. Environ. Toxicol. Chem. 36, 2274–2280. https://doi.org/10.1002/etc. 3784. Wang, C., Li, H.A., 2016. Stability and mobility of foam generated by gas-solvent/surfactant mixtures under reservoir conditions. J. Nat. Gas Sci. Eng. 34, 366–375. https://doi.org/10.1016/j.jngse.2016.06.064. Wu, W., Sun, H., 2010. Sorption–desorption hysteresis of phenanthrene—effect of nanopores, solute concentration, and salinity. Chemosphere 81, 961–967. https://doi. org/10.1016/j.chemosphere.2010.07.051. Xia, Z., Hu, L., 2019. Treatment of organics contaminated wastewater by ozone micronano-bubbles. Water 11, 55. https://doi.org/10.3390/w11010055. Yuan, L., Han, L., Bo, W., Chen, H., Gao, W., Chen, B., 2017. Simulated oil release from oil-contaminated marine sediment in the Bohai Sea, China. Mar. Pollut. Bull. 118, 79–84. https://doi.org/10.1016/j.marpolbul.2017.01.065. Zhang, M., Guiraud, P., 2017. Surface-modified microbubbles (colloidal gas aphrons) for nanoparticle removal in a continuous bubble generation-flotation separation system. Water Res. 126, 399–410. https://doi.org/10.1016/j.watres.2017.09.051. Zhang, M., Xie, L., Wang, Z., Lu, X., Zhou, Q., 2018. Using Fe(III)-coagulant-modified colloidal gas aphrons to remove bio-recalcitrant dissolved organic matter and colorants from cassava distillery wastewater. Bioresour. Technol. 268, 346–354. https:// doi.org/10.1016/j.biortech.2018.08.010. Zhang, M., Feng, Y., Zhang, D., Dong, L., Pan, X., 2019a. Ozone-encapsulated colloidal gas aphrons for in situ and targeting remediation of phenanthrene-contaminated sediment-aquifer. Water Res. 160, 29–38. https://doi.org/10.1016/j.watres.2019.05. 043. Zhang, M., Lu, X., Zhou, Q., Xie, L., Shen, C., 2019b. Polyaluminum chloride-functionalized colloidal gas aphrons for flotation separation of nanoparticles from water. J. Hazard. Mater. 362, 196–205. https://doi.org/10.1016/j.jhazmat.2018.09.022. Zhao, X., Gong, Y., O’Reilly, S.E., Zhao, D., 2015. Effects of oil dispersant on solubilization, sorption and desorption of polycyclic aromatic hydrocarbons in sediment–seawater systems. Mar. Pollut. Bull. 92, 160–169. https://doi.org/10.1016/j.marpolbul. 2014.12.042.
Li, R., Feng, C., Wang, D., Li, B., Hu, L., Shen, Z., 2016. Role of salinity in the multiphase redistribution of polycyclic aromatic hydrocarbons (PAHs) in sediment suspension. Environ. Earth Sci. 75, 116. https://doi.org/10.1007/s12665-015-5014-2. Li, Z., Liu, Y., Yang, X., Xing, Y., Yang, Q., Yang, R.T., 2017. Adsorption thermodynamics and desorption properties of gaseous polycyclic aromatic hydrocarbons on mesoporous adsorbents. Adsorpt.-J. Int. Adsorpt. Soc. 23, 361–371. https://doi.org/10. 1007/s10450-017-9863-8. Lou, L., Luo, L., Yang, Q., Cheng, G., Xun, B., Xu, X., Chen, Y., 2012. Release of pentachlorophenol from black carbon-inclusive sediments under different environmental conditions. Chemosphere 88, 598–604. https://doi.org/10.1016/j.chemosphere. 2012.03.039. Luo, J., Li, J., Huang, P., Huang, M., 2009. Kinetic rate constant of liquid drainage from colloidal gas aphrons. Chin. J. Chem. Eng. 17, 955–959. https://doi.org/10.1016/ S1004-9541(08)60302-X. Meng, Y., Liu, X., Lu, S., Zhang, T., Jin, B., Wang, Q., Tang, Z., Liu, Y., Guo, X., Zhou, J., Xi, B., 2019. A review on occurrence and risk of polycyclic aromatic hydrocarbons (PAHs) in lakes of China. Sci. Total Environ. 651, 2497–2506. https://doi.org/10. 1016/j.scitotenv.2018.10.162. Molaei, A., Waters, K.E., 2015. Aphron applications—a review of recent and current research. Adv. Colloid Interface Sci. 216, 36–54. https://doi.org/10.1016/j.cis.2014. 12.001. Oh, S., Wang, Q., Shin, W.S., Song, D.-I., 2013. Effect of salting out on the desorptionresistance of polycyclic aromatic hydrocarbons (PAHs) in coastal sediment. Chem. Eng. J. 225, 84–92. https://doi.org/10.1016/j.cej.2013.03.069. Pérez-Gramatges, A., Matheus, C.R.V., Lopes, G., da Silva, J.C., Nascimento, R.S.V., 2013. Surface and interfacial tension study of interactions between water-soluble cationic and hydrophobically modified chitosans and nonylphenol ethoxylate. Colloids Surf. A 418, 124–130. https://doi.org/10.1016/j.colsurfa.2012.11.035. Qin, L., Huang, D., Xu, P., Zeng, G., Lai, C., Fu, Y., Yi, H., Li, B., Zhang, C., Cheng, M., Zhou, C., Wen, X., 2019. In-situ deposition of gold nanoparticles onto polydopaminedecorated g-C3N4 for highly efficient reduction of nitroaromatics in environmental water purification. J. Colloid Interface Sci. 534, 357–369. https://doi.org/10.1016/j. jcis.2018.09.051. Ruen-ngam, D., Wongsuchoto, P., Limpanuphap, A., Charinpanitkul, T., Pavasant, P., 2008. Influence of salinity on bubble size distribution and gas–liquid mass transfer in airlift contactors. Chem. Eng. J. 141, 222–232. https://doi.org/10.1016/j.cej.2007. 12.024.
9
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Impact of salinity on colloidal ozone aphrons in removing phenanthrene from sediments Ming Zhang, Yudong Feng, Kaihua Zhang, Yafeng Wang, Xiangliang Pan⁎ Key Laboratory of Microbial Technology for Industrial Pollution Control of Zhejiang Province, College of Environment, Zhejiang University of Technology, Hangzhou, 310014, China
G R A P H I C A L A B S T R A C T
A R T I C LE I N FO
A B S T R A C T
Editor: Deyi Hou
Polycyclic aromatic hydrocarbons (PAHs) tend to adsorb and accumulate on sediments owing to their hydrophobicity and persistence. Salinity is the predominant factor determining the PAH partition between aqueous and solid phases in freshwater, estuaries and seawater. This study focuses on the impact of salinity on the phenanthrene (PHE) removal from sediments using an in situ and targeted remediation technology – colloidal ozone aphrons (COAs). The ozone-encapsulated colloidal aphrons exhibited increasing air holdup but decreasing stability with the salinity increasing from 0.5‰ to 35‰. The hydrophobic attraction between Tween-20-coated bubbles and the hydrophobic solid surface weakened at high salinities. The presence of inorganic ions in the aqueous phase could lead to the salting-out of nonionic compounds (PHE, Tween-20 and even ozone), hindering detaching and degrading PHE from the solid phase. Anyhow, COAs achieved high efficiencies of washing (88.0–90.2%) and oxidative degradation (74.0–76.5%) particularly for the hydrophobic sediments with highly concentrated PHE (200.4 μg/kg) over the investigated salinities. The flushing effect imposed by the bubble flow played an important role, which was not greatly influenced by salinity. Although the dissolved natural organic matter competed with PHE for COAs and led to low PHE removal, the efficiency was improved by successive COA addition.
Keywords: Colloidal ozone aphrons Functionalized microbubbles Salinity Polycyclic aromatic hydrocarbons Sediments
⁎
Corresponding author. E-mail address: [email protected] (X. Pan).
https://doi.org/10.1016/j.jhazmat.2019.121436 Received 27 April 2019; Received in revised form 30 September 2019; Accepted 8 October 2019 0304-3894/ © 2019 Elsevier B.V. All rights reserved.
Please cite this article as: Ming Zhang, et al., Journal of Hazardous Materials, https://doi.org/10.1016/j.jhazmat.2019.121436
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1. Introduction
changes the stability of CGAs (Zhang et al., 2019b). As a novel remediation technology, COAs have not been explored in terms of salinity impact on the PAH removal to the authors’ knowledge. The objective of this work is to investigate the influence of the important variable (salinity) on the properties of COAs as well as the removal of PAHs from sediments. Phenanthrene (PHE), one of the USEPA priority PAHs (Li et al., 2016), was chosen as representative PAH herein; and the typical nonionic surfactant, polyoxyethylene sorbitol anhydride monolaurate (Tween-20), was used for the COA generation. Different levels of salinity, simulating that in the aquatic environments of freshwater, estuaries and marine water, were adopted in all the experiments. The properties of COAs were characterized in terms of stability, air content, size and surface adherence. Desorption kinetics of PAHs from sediments was then studied for the target capture of PHE by the Tween-20-coated bubbles. Furthermore, the efficiency of PHE removal from sediments was explored using the COAs generated from mineral water and river water, respectively, and the operation was optimized for good performance.
Polycyclic aromatic hydrocarbons (PAHs) are ubiquitously detected in the aquatic environment. These fat-soluble and degradation-resistant organic pollutants accumulate and persist in the solid phase of water. Take the lakes in China as an instance, the concentration of ∑PAHs in sediments ranges from 6.52 μg/kg to 7935.21 μg/kg (Meng et al., 2019). Great concerns have been raised for their teratogenic, carcinogenic and mutagenic properties to living bodies, and the U.S. Environmental Protection Agency (USEPA) has listed them as priority control pollutants (Gong et al., 2018). Physic-chemical technologies are promising in the isolation, separation and even elimination of PAHs from sediments if they combine the technical advantages of sediment washing/solvent extraction and oxidative degradation. Previous studies proposed the concept of increasing the utilization of ozone within the subsurface, by which the ozone micro-nano-bubbles were generated for the site remediation of groundwater contamination (Li et al., 2014; Xia and Hu, 2019). Such bubbles are able to transport with water flow, present rapid mass transfer rates, and provide continuous ozone supply (Hu and Xia, 2018). As for the remediation of sediment contamination, more robust and persistent bubble systems are required for the in situ and targeted removal of contaminants. Noting that the modifiable microbubble system of colloidal gas aprhons (CGAs), being created from the surfactant solution, is of striking features, including large specific bubble surface area, high stability and air hold-up, strong transport capability in a pressurized environment, as well as controllable core and surface properties (Zhang and Guiraud, 2017; Zhang et al., 2018). CGAs are of great potential in contamination remediation. Given that PAHs are of strong hydrophobicity and affinitive adhesion on the surface or inside the pores of sediments (Gitipour et al., 2018), the innovative attempt can be constructing CGAs with nonionic surfactant molecules and encapsulating the gaseous oxidant – ozone into the gas core of bubbles. The so-created bubbles, being named as colloidal ozone aphrons (COAs), are expected to first adsorb on the PAH-contaminated sediments, detach those target hydrophobic contaminants from the sediment surface, and then oxidatively degrade them with the released ozone (Zhang et al., 2019a). Much of work needs to be done to examine and evaluate the COA performance; in particular, the influence of aquatic chemistry on the COA treatment process should be specifically investigated. It has been found that the hydrophobic contaminants present distinct sorption and desorption behaviors for the sediments in marine water, estuaries and freshwater, and the important difference among the abovementioned aquatic environments is salinity (Oh et al., 2013). Therefore, the study of salinity or ionic strength on the COA-remediation of PAH-contaminated sediments is crucial for process interpretation and field applications. On the one hand, the salinity of the surfactant solution may affect the solubility of surfactant molecules because the complex compounds between ions and surfactants can alter micelle formation (Abouseoud et al., 2010; Zhao et al., 2015). Since surfactant has been reported to increase the dissolution of ozone (Ji et al., 2018), the change of surfactant solubility by salinity then probably affects the ozonation during the treatment with COAs. Consequently, the COAs being generated from the surfactant solution may exhibit distinct properties and interfacial behaviors at different levels of salinity. On the other hand, it has been reported that the competitive adsorption between the anion of Cl− and the organic contaminants on the solid surface can make an important influence on the contaminantinvolved chemical reactions (Fu et al., 2019; Qin et al., 2019). Thus, through changing the partition of chemicals in solid and aqueous phases, salinity affects the environmental fate and exposure of chemicals in aquatic environment (Saranjampour et al., 2017; Jonker et al., 2015). Obviously, the interaction between PAH molecules and sediments, such as sorption, desorption and even aging, may vary with salinity. Moreover, it is known that high ionic strength significantly
2. Materials and methods 2.1. Chemicals PHE (C14H10, purity ≥99%) standards for gas chromatography analysis were provided by Aladdin, China. Tween-20 (C58H114O26, 1227.5 g/mol, chemical pure), anhydrous ethanol (analytical reagent, AR) and sodium chloride (NaCl, AR), were purchased from Sinopharm Chemical Reagent Co., Ltd, China. Dichlorodimethylsliane (DCDMS, purity ≥99%) was obtained from Adamas-beta, Switzerland. Acetonitrile (purity ≥99%) and hexane (purity ≥95%) standards for high performance liquid chromatography (HPLC) analysis were provided by Sigma-Aldrich, U.S.A.. All those reagents were used as purchased without further purification. All the solutions for the HPLC analysis were prepared using ultrapure water from a ultrapure water system (Smart-N15VF, Healforce, China) with resistivity of 18.2 MΩ cm. 2.2. Preparation of water samples with different levels of salinity The artificial water was prepared using mineral water (Nongfu Spring Co. Ltd, Hangzhou, China). The pristine properties of water, including dissolved organic carbon (DOC), pH and ion concentrations, were analyzed by total organic carbon analyzer (TOC-L, Shimadzu, Japan), pH meter (Primainst, innoLab 20 P, U.K.) and high performance ion chromatograph (Dionex™ Integrion™ HPIC™, ThermoFisher Scientific Co. Ltd., U.S.A.) and were shown in Table 1. Three levels of salinity (0.5, 10 and 35‰, g/L) were controlled with NaCl, and the corresponding conductivity values were measured to be 0.86, 16.29 and 53.16 mS/cm, respectively, by conductivity meter (DDS-307, INESA Scientific Instrument Co., Ltd., China). The water samples were then adopted in the experimental trials, including COA generation, PHE Table 1 Salient properties of pristine mineral water and river water before salinity adjustment.
2
Parameters
Mineral spring
River water
pH DOC (mg/L) Na+ (mg/L) K+ (mg/L) Mg2+ (mg/L) Ca2+ (mg/L) NH4+ (mg/L) Cl− (mg/L) NO3− (mg/L) SO42− (mg/L)
6.50 ± 0.06 0.08 ± 0.00 5.60 ± 0.01 1.27 ± 0.06 2.44 ± 0.01 20.07 ± 0.37 0.27 ± 0.00 4.28 ± 0.11 4.28 ± 0.02 7.50 ± 0.02
7.34 ± 0.01 1.17 ± 0.01 25.45 ± 0.05 4.44 ± 0.05 4.66 ± 0.02 24.19 ± 5.48 0.20 ± 0.06 23.04 ± 0.54 12.48 ± 0.11 30.91 ± 0.35
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The desorption experiments were conducted in the 500 m L-glass vials with Teflon-lined septa and screw caps at 25 °C. Those pre-contaminated sediments (0.06 kg) were added into the vial which contained 0.3 L of Tween-20-CGA suspension making from the mineral water with different levels of salinity; then, the mixture was shaken for 120 min a rotary agitator at the speed of 150 rpm. Five milliliter of water sample was taken at 10, 30, 60, 90 and 120 min, respectively. The PHE concentration on sediments was then calculated as follows:
desorption and contaminated sediment remediation. The river water was also employed for the generation of COAs which were then used in the experiments for remediating the PHEcontaminated sediments. The river water was collected from the top 30 cm of water column from Shangtang River in Hangzhou, China (gerographic coordinates: 30°17′37″ N, 12°09′44″E). Before use, the sample was passed through 0.45-μm membrane filters to remove the suspended solids and then was sterilized at 121 °C for 35 min via autoclaving. The salinity of river water, being adjusted with mineral salt (YEE® aquarium experts, China), was controlled to be close to that of the mineral water mentioned above: the conductivity was measured to be 0.23, 18.16 and 43.04 mS/cm, individually.
C (PHE )se dim ents, t C (PHE )se dim ents,0 × mse dim ents − C (PHE ) water , t × Vwater , t = mse dim ents
(1)
where, C (PHE )se dim ents, t (μg/kg) and C (PHE ) water , t (μg/L) represent the PHE concentration on sediments and in water phase, individually, at time t; C (PHE )se dim ents,0 (μg/kg) represents the initial PHE concentration on sediments; mse dim ents (kg) is the mass of sediments and has been fixed to 0.06 kg herein; Vwater , t (L) equals to the volume of water at time t which should be corrected from the water loss due to sampling:
2.3. Generation and characterization of COAs The apparatus for continuous ozone production-microbubble generation (HG-WNF-1, Hangzhou Guiguan Company, China) was used for the COA generation: ozone was first produced from air by ozonator; then, the ozone-contained air and the bubble generation solution (surfactant solution) were pumped into the saturator at the flow rate of 2.25 L/min and 360 mL/min, separately. COAs were created by pressuring dissolution of ozone-contained air at 300 kPa into water with different levels of salinity, which was the typical method of generating CGAs (Hashim et al., 2012). The nonionic surfactant, Tween-20, was chosen for bubble production to ensure the hydrophobic attraction between PHE and COAs. The concentration of Tween-20 in the bubble generation solution was determined to approximately 122.8 mg/L, higher than the critical micelle concentration (CMC) of Tween-20 (89.2 mg/L). The dissolved ozone concentration was maintained to be 0.6 mg/L which was measured with ozone analyzer (PC-II, Hach, U.S.A.). As comparison, the ozone microbubbles without surfactant and the Tween-20-CGAs without ozone were generated, individually, for the removal of PHE from sediments. After the COA generator ran for 7 min, the COA suspensions were obtained for bubble characterization or PHE removal. The COAs were analyzed in terms of stability (half-life time), air holdup and size, and the detailed measurement procedures have been elaborated in previous publications (Zhang and Guiraud, 2017; Zhang et al., 2018). The adhesion of colloidal aphrons (Tween-20-CGAs) onto the surface with different degrees of hydrophobicity was tested. One side of the glass slide (Sail Brand, China) was surface-modified with the PHE stock solution at the pre-determined PHE concentrations to reach the specific levels of hydrophobicity. The PHE concentration on the surfacetreated glass slides was measured by HPLC (LC-20AT, Shimadzu, Japan). DCDMS was also used to obtain the superhydrophobic surface. The detailed procedure of surface treatment is given in Section 1 of Supporting Information (SI). The glass slides were put in a transparent glass tank (15 × 10 × 13 cm3), and the untreated side was attached firmly onto the inner wall of the glass tank while the other side was immersed into the newly generated suspension of COAs. The images of the bubble-attached glass slides were taken at 0, 1, 2, 3, 5, 7 and 24 h with optical microscope equipped with a digital camera (IX71, Olympus, Japan). The number of bubbles per 25 mm2 on slide was counted, which was repeated at least three times for each slide. Each batch was made in triplicate.
Vwater , t = Vwater ,0 − T × ΔV
(2)
where, Vwater,0 (mL) and ΔV (mL) are the initial volume of water and the sample volume for each sampling, individually, which have been fixed to 0.3 L and 0.005 L, respectively. T represents the times of sampling at time t. 2.5. COA treatment of PHE-contaminated sediments The experiments of contaminated sediment remediation were carried out in semi-batch mode. The generated bubble suspension was pumped out at the flow rate of 252 mL/min by peristaltic pump (BT3002 J, Longer Pump, China) to the reactor for the remediation of PHEcontaminated sediments; the rest of the bubble suspension was recycled for bubble generation through a three-way valve. The COA suspension (0.1 L) was pumped into the glass reactor from the bottom and was used for remediating 0.02 kg of PHE-contaminated sediments. The reactor was sealed during the treatment for 2 h, after which the liquid and sediments were collected, respectively, to measure the residual PHE concentration. The chemical extraction process and HPLC analysis of PHE can be found in Section S3 of SI. Both of the mineral water and the river water (described in Section 2.2) with different salinities were used to explore the removal of PHE from sediments by COAs in this part of study. The efficiencies of sediment washing and PHE oxidation degradation indicated the capacity of COAs in detaching and removing PHE from the sediment matrix, respectively. The efficiencies could be defined and calculated as Eqs. (3) and (4):
PHE desorption efficiency (%) C (PHE )se dim ents,0 − C (PHE )se dim ents, t = × 100% C (PHE )se dim ents,0
(3)
where, C(PHE)sediment,0 (μg/kg) and C(PHE)sediments,t (μg/kg) represent the amount of PHE adsorbed on the sandy sediments before and after the COA treatment, individually.
2.4. Desorption kinetics of PHE from sediments
PHE oxidation degradation efficiency
(%)
(C (PHE )se dim ents,0 − C (PHE )se dim ents, t ) × mse dim ents
The model sandy sediments (6–9 mm, Helei, China), free from any contamination, were selected for the experiments of PHE desorption kinetics and contaminated sediment remediation. The artificial contamination of PHE followed the method in Section S2 of SI. The sandy sediments with and without hydrophobic surface-treatment were spiked with the PHE stock solution to obtain different homogeneous contamination levels. The characteristics of the artificial PHE-contaminated sandy sediments are shown in Table S1.
=
− C (PHE ) water , t × Vwater × 100% (C (PHE )se dim ents,0 − C (PHE )se dim ents, t ) × mse dim ents
(4)
where, C(PHE)water,t (mg/L) represents the PHE concentration in liquid after treatment; msediments (g) and Vwater (mL) are the mass of sandy sediments (0.02 kg) and the volume of liquid being collected after the treatment for 2 h. 3
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Fig. 1. Characteristics of COAs generated from Tween-20 solutions with different levels of salinity: (a) half-life time, (b) air hold-up, and (c) size distribution.
3. Results and discussion
greatly but the air holdup increased with the increase of salinity. Barrut et al. found in their study that salinity imposed significant impact on gas holdup (Barrut et al., 2012). Compared with the scenario without electrolytes, the hydrophilic repulsive force between bubbles might rise at the high salinity, particularly in the marine water, which could lead to the inhibition of Laplace pressure. Laplace pressure is known to cause the gas diffusion from small bubbles to larger ones due to the pressure difference (Ruen-ngam et al., 2008; Bera et al., 2013). As a result, the air content of the bubble system increased with the salinity increasing from 0.5‰ to 35‰ in this study.
The PHE removal by COAs depends on the properties of COAs, the PHE desorption from sediment by the hydrophobic surface of COAs, and the PHE oxidation by the released ozone from COAs. In this study, the salinity of freshwater (0.5‰), estuaries (10‰) and marine water (35‰) was specifically investigated for its impact on all the above mentioned processes. This helps to clarify and interpret the characteristic COA behaviors during the remediation of PHE-contaminated sediments in different aquatic environments. 3.1. Characteristic COA properties
3.2. Target adsorption of Tween-20-CGAs on hydrophobic surfaces
The most significant properties of colloidal aphrons include half-life time, air holdup and average size (Molaei and Waters, 2015), which indicate stability, bubble density and specific surface area of the bubble system. Experimental results denote that, with the salinity of Tween-20 solution increasing from 0.5‰ to 35‰, the half-life time of COAs dropped from 221 s to 163 s (Fig. 1(a)), the air holdup increased from 19.2% to 30.8% (Fig. 1(b)), and the average size of bubble fluctuated between 130–160 μm (Fig. 1(c)). The addition of inorganic electrolytes, such as NaCl, created a high concentration of ions in the vicinity of the micelle surface and thus favored the micelle formation. In this regard, the CMC tended to be reduced due to the presence of salt (PérezGramatges et al., 2013). Particularly, at the salinity close to that in the marine water (35‰), NaCl might play the role of salting-out agent for the non-electrolyte (Pérez-Gramatges et al., 2013; Salabat and Alinoori, 2008) which is Tween-20 in this study. In consequence, the concentration of effective surfactant molecules was reduced; and accordingly, the foamability of bubble generation solution and the stability of bubbles decreased (Wang and Li, 2016). This agreed well with what was found elsewhere (Feng et al., 2009; Luo et al., 2009). It has been reported that the bubble size is closely associated with the variation of air holdup (Bournival et al., 2014), which was, however, not consistent with the results in this work. Herein, the bubble size did not change
The COAs generated at the three levels of salinity adsorbed on the surface-modified glass slide, and the bubble density on the slide is shown in Fig. 2. For the DCDMS-treated slide which was of the highest hydrophobicity, the initial CGA density reduced from 246 bubble/ 25 mm2 to 101 bubble/25 mm2 with the salinity increasing from 0.5‰ to 35‰. Under the condition of high ionic strength, the hydrophobic attraction between the slide surface and the CGA surface with Tween20 molecules got weakened, which played the essential role in the bubble adsorption. For the CGA suspension at the fixed salinity, the bubble density decreased on the slide surface in the order of DCDMS-, PHE-modified- and untreated-slide. The difference of bubble density was found to be quite small between the high salinities of 10‰ and 35‰. At such high ion concentrations, the adsorption of the nonionic surfactant-coated bubbles onto the slide surface was hard to be further increased. The adsorbed Tween-20-CGAs drained with time, and thus, the bubble density was less than 30 bubble/25 mm2 at 24 h for all the three levels of salinity. The bubble density reduced faster with time at the lowest salinity of 0.5‰ than that at the other two higher salinities, which could be ascribed to the faster movement of Tween-20 surfactant molecules at the lower ion concentration.
4
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Fig. 2. Density of Tween-20-CGAs varying with time on hydrophobically-treated surfaces at the salinity of (a) 0.5‰, (b) 10‰ and (c) 35‰.
aggregates, which resulted in difficulty in desorption diffusion (Li et al., 2017). The comparison of desorption kinetics with (Fig. 3(a)–(c)) and without (Fig. 3(d)–(f)) Tween-20-CGAs reveals that the colloidal bubbles were apt to capture and detach the PHE molecules from the hydrophobic surface. Apparently, the addition of Tween-20-CGAs facilitated the release of PHE at different levels of salinity. At the high PHE concentration on sediment of 200.4 μg/kg (Fig. 3(a)), with the increasing addition of NaCl, the cation of Na+ adsorbed to the molecules of Tween-20, particularly, at the oxygen in
3.3. Desorption kinetics of PHE from hydrophobic surfaces using Tween-20CGAs In addition to the bubble properties, the role of salinity was also explored with respect to the PHE desorption from sediments with the assistance of Tween-20-coated colloidal aphrons. Previous study has pointed out that the peculiar molecular structure of PHE, three rings sequentially connected in a non-straight shape, could be intercepted by pores as well as twined with other PHE molecules forming larger
Fig. 3. Desorption kinetics of PHE from aquifer sediments in the presence ((a), (b) and (c)) and absence ((d), (e) and (f)) of Tween-20-CGAs at the salinity of 0.5‰, 10‰ and 35‰. The initial PHE concentrations on sediments were approximately 200.4 μg/kg ((a) and (d)), 55.8 μg/kg ((b) and (e)) and 30.0 μg/kg ((d) and (f)), individually. 5
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water at different levels of salinity, and the corresponding treatment effect can be found in Fig. 4(a)–(d), respectively. For the COA-involved experiments, the specific salinity gradient between 0.5‰ and 35‰ was tested to verify the trend of PHE removal varying with salinities (Fig. S1). The PHE desorption efficiency by COAs kept stable in 88.0–90.2% in the tested salinity range of 0.5-35‰ when PHE adsorbed on the hydrophobic sediment surface at the high concentration of 200.4 μg/kg. Compared with the PHE molecules in the inner PHE layer which firmly adhered to the surface and pores of sediments, the outer PHE might loosely adsorb and be easily washed out by COAs. Thus, it can be inferred that the desorption of highly concentrated PHE from sediments was mainly ascribed to the flow flushing caused by the bubble suspension as well as the COA capture due to the hydrophobic attraction between PHE and Tween-20. The former mechanism may be predominant and not greatly impacted by salt concentration. For the sediments with low concentration of PHE – 55.8 μg/kg for the hydrophobic surface and 24.5 μg/kg for the hydrophilic surface, the increased salinity of the COA suspension slightly decreased the desorption of PHE: with the salinity increasing from 0.5‰ to 35‰, the desorption for the former and latter scenario reduced from 68.1% to 61.7% and from 75.3% to 69.3%, respectively. Most of those hydrophobic organic molecules should be trapped by the porous structure of sediments and/or by the hydrophobic attachment with the hydrophobic sediment surface. This would conspicuously increase the difficulty for COAs to capture PHE from sediment in spite of the hydrophobic attraction of hydrophobic molecules between Tween-20 and PHE. The flushing effect which was caused by the bubble suspension flow imposed little effect at such a low PHE concentration in contrast with that at the high PHE concentration as mentioned above. Moreover, the presence of salt, particularly at high concentrations, could impede the desorption process as explained in Section 3.3. As shown in Fig. 4(b), the oxidative degradation of PHE did not vary significantly with the salinity increasing from 0.5‰ to 35‰ for the
the ethylene oxide groups. This might change the configuration through the coordination of the oxygen atoms to the Na+ in a tetragonal-pyramidal manner and form some ordered structure like microcages, which would interact preferably with the planar molecules like PHE (Li and Chen, 2002). But it should also be highlighted that the salting-out effect caused by the too high concentrations of inorganic salt (35‰), typically in of marine water, could result in increasing chemical distribution from aqueous to solid phase by reducing the solubility of target hydrophobic contaminant (such as PHE) and the nonionic surfactant molecules (Tween-20) (Wu and Sun, 2010). This was directly reflected by the results shown in Fig. 3(b) and (c). The residual PHE concentration on sediment first decreased and then increased with time for the hydrophobically (Fig. 3(a), (b), (d) and (e)) and hydrophilically (Fig. 3(c) and (f)) surface-treated sediments with different initial PHE concentrations. Hence, the desorption kinetics within the investigated 120 min could be divided into two stages – desorption and re-adsorption. This phenomenon was conspicuous for the desorption process with the assistant of Tween-20-CGAs (Fig. 3(b) and (c)). The release of PHE is kinetically controlled by the desorption itself and the diffusion through sorbent to water (Barnier et al., 2014). The latter mechanism was strongly influenced by colloidal aphrons. After the Tween-20-CGAs collapsed, the free molecules of the nonionic surfactant might adsorb onto the surface of sediments, resulting in the re-adsorption of PHE. In that case, the diffusion of PHE into the aquatic phase would be impeded.
3.4. PHE removal from sediments using COAs 3.4.1. Comparison between COAs and ozone microbubbles The performance of the ozone-encapsulated colloidal aphrons in in situ and targetedly removing PHE from sediments was studied in terms of desorption efficiency and oxidative degradation efficiency within 2 h. The COAs and ozone microbubbles were generated from the mineral
Fig. 4. Comparison of PHE removal between COAs ((a) and (b)) and ozone microbubbles ((c) and (d)): (a) and (c) desorption efficiency, (b) and (d) oxidative degradation efficiency. 6
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Fig. 5. (a) Desorption and (b) oxidative degradation of PHE from real water-based aquifer sediments.
Fig. 6. Improvement of treatment efficiency by adding COAs twice: (a) desorption and (b) oxidative degradation.
PHE from sediments. Nevertheless, as much as ∼73.8% of PHE desorbed from sediments by ozone bubbles, which was not influenced by salinity. This confirmed the speculation that, at the high PHE concentration, the flow flushing (or, the washing effect) contributed greatly to the desorption process. When the hydrophobic or hydrophilic sediment surface was loaded with smaller amount of PHE (55.8 μg/kg and 24.5 μg/kg, individually), the PHE removal by ozone microbubbles was higher than that induced by COAs. For the ozone microbubble-treated sediments, the addition of NaCl led to the increase of surface tension, internal pressure of bubbles, and high mass transfer efficiency of ozone in the absence of nonionic surfactant (Xia and Hu, 2019). As for the COA-involved treatment, the high amount of ions in aqueous phase resulted in the salting-out of the nonionic compounds (El-Nahhal and Safi, 2004) – PHE and Tween-20 as well as the encapsulated ozone, and hence, the nonionic compounds were apt to adsorb on, rather than desorb from, the solid phase. Obviously, when the ozone dosage was fixed and the salinity was controlled below 10‰, COAs performed better on the sediments with heavy PHE contamination (i.e. 200 μg/kg) whereas the ozone microbubbles took greater advantages in removing PHE from sediments with relatively low PHE concentration.
sediments with high and low PHE concentrations. Salinity was found to affect ozone oxidation of hydrophobic organic components in some contrasting aspects (Ji et al., 2018): negatively, (1) NaCl may compete for dissolved ozone and radicals, and (2) the raised ionic strength pushes more PAHs to the Tween-20/water cages; positively, (1) the nonionic surfactant of Tween-20 may entrain more PHE to the top layer of the water, which is conducive to the contact and reactions with gaseous ozone in the headspace, and (2) the elevated ionic strength probably increases the interfacial concentration of the hydrophobic hydrocarbons of PHE at the ozone-solution interface due to the “salting-out” effect; but this effect can be partially offset because of the lowered surface tension and the increased solubilization of PHE. Besides, it is noted that the oxidative degradation of PHE was 74.0–76.5%, 54.0–56.7% and 33.8–36.1% for the sediments with high and low concentrations of PHE, individually. During ozonation, the direct attack by ozone molecules is the predominant mechanism of PHE oxidation (Gong and Zhao, 2017). As for the COA treatment process, the PHE molecules desorbed from the sediments would preferably attach on the COA surface due to the hydrophobic attraction between the molecules of PHE and Tween-20. Then, those PHE molecules could be directly attacked by the ozone molecules released from the drained COAs. Therefore, the efficiency of PHE oxidation degradation by COAs was consistent with that of desorption (shown in Fig. 4(a)), particularly for the PHE adsorbed on the hydrophobic sediment surface. In comparison of the treatment with COAs, the ozone microbubbles, being generated in the absence of surfactant, were adopted to remove PHE from sediments (see in Fig. 4(c) and (d)). For the hydrophobically surface-treated sediments with the PHE concentration of 200.4 μg/kg, the desorption and oxidative degradation efficiencies induced by ozone microbubbles were around 12–15% and 4–10%, respectively, being lower than those caused by COAs. In contrast to the bubbles without surfactant coating, the Tween-20-layer of COAs facilitated the release of
3.4.2. River water-based COAs for PHE removal To further investigate the COA performance under natural aqueous conditions, the river water, rather than mineral water, was employed for bubble generation as described in Section 2.2. By comparing the results in Fig. 5(a) and Fig. 4(a), it can be found out that the PHE desorption efficiency decreased by 6–20% in the river water-involved experimental trials no matter that the PHE concentration was high or low and that the sediment surface was hydrophobic or hydrophilic. As presented in Table 1, the DOC in river water was approximately 14 times higher than that in mineral water; thereby, the most significant difference between the two waters was the content of dissolved natural 7
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DNOM. For the improvement of remediation effect, the successive treatment of sediments (totally, twice) was adopted and in excess of 90.0% and 89.0% of desorption and oxidative degradation efficiency was achieved. The study sheds light on adapting the operation conditions of COA technology to specific salinity environments for the effective remediation of PAH-contaminated sediments.
organic matter (DNOM). As a diverse and complex mixture with a wide range of molecular weight and chemical structures, DNOM has multiple interaction with organic pollutants, affecting their distribution in waters with varied salinities (Lou et al., 2012; Yuan et al., 2017). The addition of mineral salt prevented the DNOM (particularly that with hydrophobic moieties) from releasing from sediments. This would further enhance the adsorption of highly concentrated PHE molecules on sediments through the hydrophobic attraction between PHE and DNOM. It is worth noting that, even though the salinity influenced the desorption process induced by the COAs generated from river water, the impact was different for the hydrophobic and hydrophilic sediment surface. Higher salinities (10‰ and 35‰) tended to hinder the PHE release from the hydrophobic surface whilst the opposite result was obtained for the hydrophilic surface. The PHE molecules were supposed to loosely adsorb on the hydrophilic sediment surface, and the partition of PHE in the aqueous phase could be enhanced at high ionic strength. The results of oxidative degradation in Figs. 4(b) and 5 (b) indicate the drop of efficiency for the river water-involved tests. This could be explained by the consumption of ozone by DNOM. In particular, for the hydrophobically surface-treated sediments, the increase of salinity gave rise to the decrease of oxidative degradation efficiency. The Tween-20 molecules might not only salt out but also be entrained by the undissolved DNOM, which negatively affected the encapsulation of ozone by the surfactant molecules during the formation of COAs. Additionally, in excess of 90.0% and 89.0% of desorption and oxidative degradation efficiencies could be achieved via the successive treatment of sediments with the same volume of COA suspension (0.1 L for each time, and totally, twice) as revealed in Fig. 6. On the one hand, the adsorption of recalcitrant PHE on sediment could get weakened by the first added COAs; and thus, it would be easier for the later added COAs to detach and oxidize the PHE molecules. On the other hand, the desorbed but not oxidized PHE molecules in the first treatment would be removed with ozone during the following COA treatment. Noting that the PHE removal was still negatively affect by high salinity in the twice COA treatment.
Acknowledgements This research was supported by the National Science Foundation of China (No. 51608373, No. U1703243 and No. U1503281), the Zhejiang Provincial Natural Science Foundation of China (LY19E080018), and the Scientific Starting Foundation of Zhejiang University of Technology (2017129008229). Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.jhazmat.2019.121436. References Abouseoud, M., Yataghene, A., Amrane, A., Maachi, R., 2010. Effect of pH and salinity on the emulsifying capacity and naphthalene solubility of a biosurfactant produced by Pseudomonas fluorescens. J. Hazard. Mater. 180, 131–136. https://doi.org/10.1016/ j.jhazmat.2010.04.003. Barnier, C., Ouvrard, S., Robin, C., Morel, J.L., 2014. Desorption kinetics of PAHs from aged industrial soils for availability assessment. Sci. Total Environ. 470–471, 639–645. https://doi.org/10.1016/j.scitotenv.2013.10.032. Barrut, B., Blancheton, J.-P., Champagne, J.-Y., Grasmick, A., 2012. Mass transfer efficiency of a vacuum airlift-application to water recycling in aquaculture systems. Aquacult. Eng. 46, 18–26. https://doi.org/10.1016/j.aquaeng.2011.10.004. Bera, A., Ojha, K., Mandal, A., 2013. Synergistic effect of mixed surfactant systems on foam behavior and surface tension. J. Surfactants Deterg. 16, 621–630. https://doi. org/10.1007/s11743-012-1422-4. Bournival, G., Du, Z., Ata, S., Jameson, G.J., 2014. Foaming and gas dispersion properties of non-ionic surfactants in the presence of an inorganic electrolyte. Chem. Eng. Sci. 116, 536–546. https://doi.org/10.1016/j.ces.2014.05.011. El-Nahhal, Y.Z., Safi, J.M., 2004. Adsorption of phenanthrene on organoclays from distilled and saline water. J. Colloid Interface Sci. 269, 265–273. https://doi.org/10. 1016/S0021-9797(03)00607-6. Feng, W., Singhal, N., Swift, S., 2009. Drainage mechanism of microbubble dispersion and factors influencing its stability. J. Colloid Interface Sci. 337, 548–554. https://doi. org/10.1016/j.jcis.2009.05.054. Fu, Y., Qin, L., Huang, D., Zeng, G., Lai, C., Li, B., He, J., Yi, H., Zhang, M., Cheng, M., Wen, X., 2019. Chitosan functionalized activated coke for Au nanoparticles anchoring: green synthesis and catalytic activities in hydrogenation of nitrophenols and azo dyes. Appl. Catal. B 255, 117740. https://doi.org/10.1016/j.apcatb.2019.05. 042. Gitipour, S., Sorial, G.A., Ghasemi, S., Bazyari, M., 2018. Treatment technologies for PAHcontaminated sites: a critical review. Environ. Monit. Assess. 190, 546. https://doi. org/10.1007/s10661-018-6936-4. Gong, X., Xiao, L., Zhao, Z., Li, Q., Feng, F., Zhang, L., Deng, Z., 2018. Spatial variation of polycyclic aromatic hydrocarbons (PAHs) in surface sediments from rivers in hilly regions of Southern China in the wet and dry seasons. Ecotoxicol. Environ. Saf. 156, 322–329. https://doi.org/10.1016/j.ecoenv.2018.03.004. Gong, Y., Zhao, D., 2017. Effects of oil dispersant on ozone oxidation of phenanthrene and pyrene in marine water. Chemosphere 172, 468–475. https://doi.org/10.1016/j. chemosphere.2017.01.007. Hashim, Mohd. A., Mukhopadhyay, S., Gupta, B.S., Sahu, J.N., 2012. Application of colloidal gas aphrons for pollution remediation. J. Chem. Technol. Biotechnol. 87, 305–324. https://doi.org/10.1002/jctb.3691. Hu, L., Xia, Z., 2018. Application of ozone micro-nano-bubbles to groundwater remediation. J. Hazard. Mater. 342, 446–453. https://doi.org/10.1016/j.jhazmat. 2017.08.030. Ji, H., Gong, Y., Duan, J., Zhao, D., Liu, W., 2018. Degradation of petroleum hydrocarbons in seawater by simulated surface-level atmospheric ozone: reaction kinetics and effect of oil dispersant. Mar. Pollut. Bull. 135, 427–440. https://doi.org/10. 1016/j.marpolbul.2018.07.047. Jonker, M.T.O., van der Heijden, S.A., Kotte, M., Smedes, F., 2015. Quantifying the effects of temperature and salinity on partitioning of hydrophobic organic chemicals to silicone rubber passive samplers. Environ. Sci. Technol. 49, 6791–6799. https://doi. org/10.1021/acs.est.5b00286. Li, J.-L., Chen, B.-H., 2002. Solubilization of model polycyclic aromatic hydrocarbons by nonionic surfactants. Chem. Eng. Sci. 57, 2825–2835. https://doi.org/10.1016/ S0009-2509(02)00169-0. Li, H., Hu, L., Song, D., Lin, F., 2014. Characteristics of micro-nano bubbles and potential application in groundwater bioremediation. Water Environ. Res. 86, 844–851. https://doi.org/10.2175/106143014X14062131177953.
4. Conclusions As a distinct feature of freshwater, estuaries and marine water, salinity was explored for its impact on the in situ and targeted removal of PHE from sediments by the COA system in this work. With the increase of salinity from 0.5‰ to 35‰, the stability of COAs decreased, the air holdup increased, and the bubble size was almost unaffected. At high salinities, the salting-out effect reduced the solubility of Tween-20, which accelerated the drainage of COAs; meanwhile, the raised interfacial and surface tension gave rise to the increased foamability and thus the improved air content. Those Tween20-coated COAs preferably adsorbed on the hydrophobic surface; but the most essential contributor – the hydrophobic attraction weakened at high ionic strength. Compared with the ozone microbubbles, the COAs showed better performance in desorption (88.0–90.2%) and oxidative degradation (74.0–76.5%), particularly for the hydrophobic sediment surface with the high PHE concentration of 200.4 μg/kg. The effect remained stable over the investigated salinity range. In this scenario, the PHE molecules in the outer PHE layer might loosely adsorb on the sediment surface and easily desorb due to both of the bubble flushing flow and the COA capture. The former effect was not greatly impacted by salinity. As for the desorption of lowly concentrated PHE, the efficiency was not high particularly at high salinities. Herein, the salting-out effect upon PHE played the dominant role in impeding the release of PHE. When DNOM was presented in the system, it could consume the COAs and lead to low PHE removal. The addition of mineral salt inhibited the DNOM (particularly that with hydrophobic moieties) from releasing from sediments, which further enhanced the adsorption of highly concentrated PHE on sediment through the hydrophobic attraction between PHE and 8
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