Microplastics as Both a Sink and a Source of Bisphenol A in the

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Environmental Processes

Microplastics as both a Sink and a Source of Bisphenol A in the Marine Environment Xuemin Liu, Huahong Shi, Bing Xie, Dionysios D. Dionysiou, and yaping Zhao Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.9b02834 • Publication Date (Web): 08 Aug 2019 Downloaded from pubs.acs.org on August 10, 2019

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Environmental Science & Technology

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Microplastics as both a Sink and a Source of Bisphenol A in the Marine Environment

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Xuemin Liu,† Huahong Shi,‡ Bing Xie,† Dionysios D. Dionysiou,§,* and Yaping Zhao†,*

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†School

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Process and Eco-Restoration, East China Normal University, Shanghai 200241, China

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‡State

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200241, China

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§Environmental

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Engineering (DChEE), 705 Engineering Research Center, University of Cincinnati, Cincinnati, Ohio, USA

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ABSTRACT:

of Ecological and Environmental Sciences, Shanghai Key Laboratory for Urban Ecological

Key Laboratory of Estuarine and Coastal Research, East China Normal University, Shanghai,

Engineering and Science Program, Department of Chemical and Environmental

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Microplastics were demonstrated to be an environmental sink for hydrophobic organic

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pollutants, while they can also serve as a potential source of such pollutants. In this study, the

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sorption and release of bisphenol A in marine water was investigated through laboratory

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experiments. Sorption and desorption isotherms were developed and the results reveal that

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sorption and desorption depend on the crystallinity, elasticity and hydrophobicity of the

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polymer concerned. The adsorption and partition of bisphenol A can be quantified using a

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dual-mode model of the sorption mechanisms. Polyamide and polyurethane were found to

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exhibit the highest sorption capacity for bisphenol A, and it was almost irreversible, probably

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due to hydrogen bonding. Polyethylenes and polypropylene exhibited high and reversible

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sorption without noticeable desorption hysteresis. Glassy polystyrene, polyvinylchloride,

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polymethyl methacrylate and polyethylene terephthalate exhibited low sorption capacity and

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only partial reversibility. Low-density polyethylene and polycarbonate microplastic particles

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were for the first time proved to be a persistent source releasing bisphenol A into aquatic

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environments. Salinity, pH, co-existing estrogens and water chemistry influence the

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sorption/desorption behaviors to different degree. Plastic particles can serve as transportation 1

ACS Paragon Plus Environment


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vectors for bisphenol A, which may constitute an ecological risk.

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Key words: Microplastics, Bisphenol A, sorption, sources, sinks, environmental risks,

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marine pollution

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

INTRODUCTION

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Plastic debris is widely regarded as a global environmental threat exacerbated by the

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increasing mass production of waste plastic over the last few decades and its poor

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management.1, 2 Microplastics (MPs, with diameter of > PS > PEs (LLD, MD and HD) ≈ PP > PMMA ≈ PVC and PET. The

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enrichment factor of BPA on MPs could reach up to 105 and the calculated distribution

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coefficient Kd varied from 0.41 L kg-1 (with PVC) to 76,287 L kg-1 (for PA) (at 0.005 Sw,

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Table S6). The huge differences in BPA’s solid-liquid distribution can be ascribed to the

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different structural properties of the diverse MPs tested.

180

Usually, the sorption of HOCs by organic matter in soil is thermodynamically driven,

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involving specific hydrophobic interactions, hydrogen bonding and π-π interaction.40 Plastics

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can be simplified by analogy with organic matter in soil to study specific MP-HOC

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interactions.15, 41 The static water contact angle is widely used to describe the hydrophobicity

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of MPs (Figure S2, Table S2). A larger contact angle indicates greater hydrophobicity.42, 43 A

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significant positive correlation between sorption capacity for BPA (Qe) and the water contact

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angle of MPs was observed in this study (Table S3, p≤ 0.05 at Ce= 0.005 Sw). The MPs’ BPA

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sorption capacity was highly dependent on their hydrophobicity. Similarly, about 66% of the

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BPA absorbed by non-woven PP can be attributed to hydrophobic interaction, indicating the

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crucial role of hydrophobic interaction played in BPA sorption.44,

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tested, PA and PU exhibited the largest sorption affinity towards BPA. That should be

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ascribed to the formation of hydrogen bonds between H-bond-donating BPA and the H-bond

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accepting amide groups in PA or PU.47, 48 T-tests confirmed that the BPA sorption capacity

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(Qe) on PA or PU through H-bond interaction was significantly greater than that of the other

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MPs tested (Table S4, p≤ 0.001 at Ce= 0.005 Sw). PS particles exhibited a medium level of

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BPA sorption capacity comparing with that of MDPE, PP, PVC, PMMA or PET. That can

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mainly be ascribed to the contribution of π−π interactions between BPA’s benzene rings and

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those of PS.49, 50 In addition, the benzene ring in PS’s polymeric backbone restricts segmental

45, 46

Among the MPs

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mobility within the chains and increases the distance between adjacent chains, allowing BPA

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to more easily diffuse into a PS matrix.18

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The polymers’ structure consists of highly ordered crystalline domains and less structured

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amorphous domains. Crystallinity can influence the sorption of HOCs because the loosely

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packed polymer chains in the amorphous domains increase the accessibility of HOCs into the

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inner-space of polymer compared to the crystalline domains.15,

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calorimeter (DSC) data can characterize the abundance of amorphous and crystalline domains

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in a polymer (Figure S3, Table S2). The data show that crystallinity is significantly correlated

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with hydrophobicity (r= 0.599, *p≤ 0.05, Table S3), so greater BPA sorption capacity would

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be expected from MPs with higher crystallinity. However, no significant correlation between

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the BPA sorption capacities of the MPs and their crystallinity was observed at Ce= 0.005 Sw

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(Table S3), suggesting that crystallinity has at best a minor effect on MPs’ sorption of BPA.

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Despite the effect of crystallinity, MPs’ hydrophobicity was also affected by other factors,

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such as surface functional groups and surface roughness.51 Therefore, the effect of

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hydrophobicity derived from crystallinity on MPs’ sorption of BPA might decrease, resulting

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in a weak correlation between Qe and crystallinity.

25

Differential scanning

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PEs and PP particles with a lower fraction of amorphous domains were found to have

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greater BPA sorption capacity. This result is consistent with previous reports that the rubbery

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plastic PE has higher sorption capacity towards phenanthrene and perfluorooctanesulfonates

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(PFOS) than that of glassy PS and PVC.10,

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domains rather than its glassy domains predominantly determine its BPA sorption capacity.24

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The relatively expanded, flexible structure of rubbery plastics like PEs and PP should thus

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allow greater mobility, diffusivity, and accessibility for BPA than the more condensed and

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less flexible structure of glassy MPs. T-tests confirmed this explanation (at Ce= 0.005 Sw, *p≤

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0.05, Table S5). Similar studies have indicated that rubbery PEs and PP possess a higher free

25

It can be inferred that a polymer’s rubbery

9



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volume of internal cavities in the rubbery domains compared with that of glassy PET and

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PVC, contributing to greater preferential accumulation of PAHs and PCBs in PEs and PP.8, 11,

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52

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The importance or the effect of specific surface area and particle size on the BPA sorption

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by MPs should also be stressed. Smaller particles with greater specific surface area (Sarea)

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have been shown to have greater sorption capacity of BPA on PP because the larger surface

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area increases the availability of sorption sites.53,

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preferential sorption of tyosin and pyrene on PE compared to PS and PVC is mainly due to

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PE’s larger specific surface area.55, 56 In this study, however, no significant correlation was

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observed between Sarea or the particle size of the various MPs and their BPA sorption

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capacity (at Ce= 0.005 Sw, Table S3), indicating that the influence of Sarea or particle size was

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negligible compared with that of crystallinity, hydrophobicity and a polymer’s rubbery

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nature. It is obvious, though, that for the same type of polymer (take MDPE as an example),

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smaller particles and larger specific surface area should indeed contribute to the greater

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sorption of BPA. This has been confirmed in a previous study.36

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Sorption Mechanisms.

54

Some researchers have suggested that

239

Sorption of HOCs onto polymers has often been described in terms of a dual-mode

240

polymer sorption model involving both adsorption and partition mechanisms. The adsorption

241

of HOCs in amorphous domains is identified as solid–phase dissolution described using a

242

partition model. In the polymer’s crystalline domains, adsorption and pore–filling better

243

describe the interaction.57 The total sorption is the sum of the two processes, partition and

244

adsorption. So

245

QT = QA + QP

246

where QT represents the total sorption capacity, QA is the adsorption capacity and QP is the

(1)

10


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247 248


partition capacity.58, 59 The sorption isotherms of BPA on various MPs are well fitted using Freundlich isotherm

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model, thus the total sorption can be described as

250

QT = a Cen

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where a and n are constants. In this dual-mode model of the interaction, the contribution of

252

adsorption is nonlinear and can reach saturation rapidly depending on the surface area of the

253

MPs, while the contribution of partition can increase linearly with time depending on the

254

BPA concentration and the type of MPs. The isotherms should be linear at high

255

concentrations, and the total sorption amount QT′ can then be described as

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QT′= MCe + N

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where M and N are constants. MCe is the sorption contribution through partition (QP) and N

258

represents the maximum adsorption capacity (QAmax). Therefore, the adsorption capacity (QA)

259

can be calculated as:

260

QA = aCen- MCe

(2)

(3)

(4)

261

Equations 1–4 thus allow quantifying the isotherm parameters QT, QP, and QA (Table S6).

262

The contributions of QP and QA can be quantified by fitting a dual-mode sorption model,

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which depends heavily on the BPA concentrations and the structural properties of MPs

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involved (Figure S4). Within the range of BPA concentrations studied here, the sorption by

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rubbery MDPE, HDPE and PP and by glassy PET was dominated by adsorption at low BPA

266

concentrations (< 200 μg L-1) but then by partition at higher BPA concentrations. For the

267

glassy PS, PVC and PMMA polymers and rubbery LLDPE, adsorption was more important

268

than partition at all of the concentrations studied. Given a high enough BPA concentration,

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the sorption of BPA on MPs will eventually be partition-dominated due to hydrophobic

270

interaction.60 For glassy PA and PU, the hydrophobic, partition-dominated sorption process is 11



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remarkably strengthened by hydrogen-bond interactions. The hydrogen bond energy values

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(2–40 kJ mol-1) involved are much higher than those of hydrophobic bonds (5 kJ mol-1).61

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The Effects of Water Chemistry on BPA Sorption.

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Polymer particles in aqueous environments can be carried out to sea through river

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networks or be transported back into estuarine habitats from the sea via tidal flow.62 Complex

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interactions between such particles and any BPA present may vary in different water

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environments. The particles’ sorption capacity for BPA with tend to decrease as the pH

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increases from 3.0 to 11.0 (Figure 1b), probably due to BPA’s greater solubility and

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hydrophilicity at higher pH values. In addition, BPA will exist partially as an anion (pKa=

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9.6–10.2) and the surfaces of MPs tend to be negatively charged at high pH values (For

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example, the pHPZC of PE, PP, and PS are at pH 4.30, 4.26 and 3.96, respectively).52 That

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might increase the electrostatic repulsion between BPA and the MPs, decreasing the BPA

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sorption capacity. In natural surface water (6.5< pH< 8.5), MPs will exhibit relatively stable

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sorption of BPA and can be regarded as potential BPA sinks. Sorption of estrone,

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17β-estradiol and 17α-ethinylestradiol on PA 6 has also been shown to decline drastically

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with increasing pH (from 2 to 12), especially at pH> 10.5 (the pKa of many

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endocrine-disrupting compounds).63 However, Wang et al. revealed that the sorption of PFOS

288

on PE and PS increased with a decrease in pH (from 7.0 to 3.0), but that pH had no effect on

289

the sorption of PFOS on PE or PS.10 Interestingly, Xu et al. have reported that the sorption of

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tetracycline on PE, PP and PS initially increases then decreases as the pH increases from 2.0

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to 12.0 due to the fact that tetracycline is a zwitterionic compound.64

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Usually, high salinity would increase the availability of BPA for sorption onto MPs owing

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to salting-out. In this study, when salinity was increased from 12 ‰ to 35 ‰, it indeed had a

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slightly negative effect on the sorption capacity of the MPs, except for those of PA and PU

295

(Figure 1c). Higher salinity has also been shown to inhibit the sorption of tetracycline and 12


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musk on MPs.64, 65 However, increased salinity showed no obvious effect on the sorption of

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phenanthrene by PE or PVC. It did, though, decrease the sorption of DDT.13 Increased

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salinity enhanced the sorption of triclosan but showed only a minor influence on

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4-methylbenzylidene camphor, carbamazepine and 17α-ethinylestradiol sorption on PE.43

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The effects of salinity on sorption may diverge depending on the nature of the adsorbate. In

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this work, the influence of salinity on BPA sorption on MPs can be neglected.

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The BPA sorption capacity was in the sequence of ultrapure water> Yangtze River water>

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seawater (Figure 1d). The presence of dissolved organic matter (DOM) appears to be the

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main factor affecting the sorption capacities in different water matrices. The DOM content in

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the ultrapure water, Yangtze River water and seawater was negatively correlated with the

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BPA sorption capacity, which might be attributed to enhanced dissolution of BPA through

307

the formation of BPA/DOM complexes. Previous studies have shown that there is negligible

308

interaction between MPs and DOM,43, 64 so it is probably the combination of BPA with DOM

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that reduces the BPA sorption capacity.

310

In water environment, HOCs are likely to be present and may compete for MPs’ sorption

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sites. The BPA sorption capacity of the various MPs did not seem to change much at E2

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concentrations ranging from 10 to 100 μg L-1 (Figure 1e). Moreover, the sorption capacity of

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E2 was basically unaffected in binary systems compared with that in single solute systems

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(Figure 1f). That is to say, E2 will basically not influence the sorption of BPA on MPs except

315

LDPE and PC, and vice versa. However, the presence of E2 inhibits the leaking of BPA from

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LDPE and PC with different degree. It has been reported that DDT appears to have an

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antagonistic effect on the sorption of phenanthrene on MPs, because DDT (log Kow of 6.79) is

318

more hydrophobic than phenanthrene (log Kow of 4.6).37, 66 The similar hydrophobicity of E2

319

(log Kow of 3.94) and BPA (log Kow of 3.32) may also explain why E2 and BPA did not show

320

any synergistic or antagonistic effect on each other’s sorption. 13



(a)

(b)

20

LDPE HDPE PC PMMA

15

LLDPE PP PVC PET

MDPE PS PA PU

4 2

10

0

5

-2

0

-4 -12

-18 0

200

400 600 -1 Ce g L 

(d) 0

0

-1

-1 -1 Ultrapure water 12 ‰ salinity 24 ‰ salinity 35 ‰ salinity

-2 -12 -18

(e)

4

-2

Ultrapure water Yangtze river water Real seawater

-12 -18 (f) 3

BPA

2 0 -2 -12 -18

-1

10g L BPA -1 -1 10 g L BPA-50 g L E2 -1 50 g L BPA -1 -1 50 g L BPA-50 g L E2 -1 100 g L BPA -1 -1 100 g L BPA-50 g L E2

Qe of E2 g g-1)

Qe of BPA g g )

pH= 2.85 pH= 6.90 pH= 7.60 pH= 9.10 pH=11.10

800

(c)

321 322

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-1

50 g L E2 -1 -1 10 g L BPA + 50 g L E2 -1 -1 50 g L BPA + 50 g L E2 -1 -1 100 g L BPA + 50 g L E2

E2

2

1

0

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Figure 1. (a) The sorption isotherms of BPA (100 μg L-1) on MPs, (b) the effect of pH, (c)

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salinity (10 μg L-1), (d) water chemistry (10 μg L-1), (e) competition of E2, and (f) the effect

325

of competition from BPA on E2 sorption (The order of MPs from left to right is as follows:

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LDPE, LLDPE, MDPE, HDPE, PP, PS, PC, PVC, PA, PMMA, PET, and PU).

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MPs as Source for BPA.

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Leaching of BPA from Virgin MPs.

329

How likely is it that virgin LDPE and PC might release BPA into an aqueous environment? 14


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The sorption isotherms in Figure 1a suggest that a certain amount of the BPA present in

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LDPE and PC might be leach out, depending on the BPA concentration. BPA may be present

332

as an additive or as unreacted monomer in plastics manufacture.35 The release of BPA from

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LDPE and PC in ultrapure water was faster initially and then decreased with time to reach

334

equilibrium within 3 days (Figure 2a). The BPA amount released reached up to 2.68 μg g-1

335

from LDPE and 14.45 μg g-1 from the PC tested. While, the leaching of BPA, phthalates, and

336

citrates from 5×5 mm squares of PE, PS, PET, and PVC to reach equilibrium required 80 days

337

in simulated seawater.67 The smaller plastic particles tested here presumably shortened the

338

diffusion length from the inner part to the outer surface of the MPs in these experiments.

339

LDPE and PC can be persistent sources demonstrated through three consecutive releasing

340

cycles (inset of Figure 2a). The release of BPA into water from plastics bottles has only been

341

observed at temperatures of 80–100 °C.29,

342

sustained release of BPA from LDPE and PC at room temperature, suggesting a potential

343

environmental risk of the release of plastic additives BPA under indoor conditions.

68

But this study has demonstrated a rapid and

344

The amount of BPA released from LDPE increased slightly with increasing pH from 3.0 to

345

11.0, while BPA release from PC increased more dramatically (Figure 2b). Theoretically,

346

alkaline conditions will promote the ionization of BPA molecules and catalyze the hydrolysis

347

of PC, thus increasing solubility of BPA in water leading to its release.35 This has been

348

confirmed in research on the release of BPA from PC baby bottles into water at 80 ℃ .

349

Released BPA attained concentrations reaching up to 20 μg L-1 at pH 6 and 1010 μg L-1 at pH

350

12.30 Higher salinity, by contrast, would be expected to inhibit BPA release through

351

salting-out effect. As the water’s salinity increased from 12 ‰ to 35 ‰, the BPA released

352

capacity decreased from 2.66 to 2.19 μg L-1 on LDPE or from 14.52 to 3.15 μg L-1 on PC

353

(Figure 2c), respectively. Obviously, the inhibiting effect of salinity on BPA release is more

354

significant with PC particles (decreased by 78.2 %) than with LDPE (decreased by 17.5 %). 15



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Figure 2d shows the BPA released from LDPE or PC in different water matrices. The

356

sorption amounts are in the order Yangtze River water> ultrapure water> seawater>

357

simulated seawater> simulated physiological fluid. Considering the water matrices with the

358

same salinity (ultrapure water vs. Yangtze River water; simulated seawater vs. real seawater

359

or simulated physiological fluid), significantly more BPA was released in the real

360

environmental water, presumably due to its higher pH and the presence of DOM.64 Release of

361

BPA from both LDPE and PC was dramatically less in the simulated body fluid. Previous

362

studies reported that the presence of sodium taurocholate in simulated body fluid led to

363

greater desorption of phenanthrene, DDT, and DEHP from PE and PVC, and as much as a

364

20-fold increase in the desorption rate compared with that in seawater.8, 19 This divergence

365

may be ascribed to the fact that phenanthrene, DDT, and DEHP are not pH-sensitive

366

chemicals, whereas BPA will ionize at higher pH. Consequently, low pH (4.0) and high

367

salinity (3.5 ‰) should be responsible for the smaller release of BPA from LDPE and PC in

368

simulated physical conditions. Other factors that maybe lead to inhibit desorption of BPA

369

from MPs in simulated body fluid still need to be explored. 20

(a)

PC

40

LDPE

15 10

Released BPA (g g-1)

10

20

0

5

1st

10

2st 3st

0 0 20

0

100 200 Time (hours)

(c)

15

2 20

PC LDPE

15

10

10

5

5

0 0

370

PC LDPE

(b)

30

10 20 30 Salinity (‰)

40

4

6 pH

8

10

(d)

Ultrapure water Yangtze River water Simulated seawater Real seawater Simulated physical condition

0 LDPE

PC

371

Figure 2. BPA release from LDPE and PC as a function of (a) shaking time (with an inset

372

showing three consecutive releases from the same particles), (b) solution pH, (c) salinity, and 16


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(d) water matrices.

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Desorption of BPA from Adsorbed MPs.

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MPs have been regarded as transporting vectors for HOCs in marine environments.13

376

Desorption of BPA from BPA/MPs combinations during transport would be expected,

377

depending on the environmental conditions.50 In this study, the BPA desorption from the

378

rubbery PEs and PP was generally greater than from the glassy PVC, PS, PMMA, PET and

379

PU (Figure 3a). It is worth mentioning that no BPA desorption from PA was observed even

380

after 120h in ultrapure water, indicating the irreversible sorption process of BPA due to

381

strong hydrogen bonding. BPA was detected in three consecutive desorption experiments

382

whether in ultrapure water or seawater (Figures 3b and 3c). A large portion of the sorbed

383

BPA (70–90 %) was desorbed from the PEs and PP over 3 cycles of desorption in ultrapure,

384

but only smaller fractions were desorbed from PS, PMMA, and PET particles (0–45 %). BPA

385

desorption was significantly lower in seawater than in ultrapure water (Figure 3c), indicating

386

an obvious salting-out effect. The desorption of BPA from PEs, PP and PVC is therefore to

387

some extent reversible, but less so from PS, PMMA, and PET. Desorption of BPA from MPs

388

in simulated physiological conditions was below the study’s detection limit (0.3 ng L-1).

389

However, a previous study has shown that the desorption of phenanthrene and DDT from PE

390

and PVC under gut conditions could be up to 30 times greater than in seawater.19 This

391

discrepancy may be because BPA is pH-sensitive compared with phenanthrene and DDT,

392

which inhibits its desorption in simulated physiological conditions.

393

To verify these results, sorption-desorption isotherms were plotted and an index of

394

hysteresis was calculated. The index results were in the order PMMA (2.69–3.81)≫ PET

395

(0.64–1.27)> PS (0.32–0.82)> PVC (0.27–0.29)> PEs and PP (0.04–0.20) (Figure S5). That

396

there was no significant desorption hysteresis with the PEs and PP suggests that the

397

desorption processes were essentially reversible. This result further indicates that the sorption 17



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of BPA in flexible, rubbery subdomains of PEs and PP is a reversible process.24 In contrast,

399

the hysteresis loops in the sorption-desorption isotherms of the glassy PVC, PS, PMMA, and

400

PET suggest that desorption was only partially reversible process due to the polymers’ rigid

401

or condensed inner structure.24 That would inhibit desorption of BPA. Irreversible molecular

402

interactions such as π-π bonding will cause such desorption hysteresis.69 Liu, however, found

403

that desorption of BPA from PS nanoparticles was essentially reversible.21 This divergence

404

can presumably be ascribed to the difference in particle sizes, but it still needs to be explored.

40 20 0 50

100

150

Time (h)

Ultrapure water

60

(b)

20

(c)

Seawater

15

40

10

20

5

0

200

25

0

st

1 desorption

LLDPE MDPE HDPE PP PS PVC PA PMMA PET PU

60

0

80


(a)


Desorption (%)

80

nd

2 desorption

rd

3 desorption

405 406

Figure 3. (a) Desorption kinetics of BPA from MPs, (b) 3-cycle consecutive desorption of

407

BPA from MPs in ultrapure water, (c) in seawater.

408

These kinetics, recycling and marine water experiments have demonstrated LDPE and PC

409

(Figure 2) and BPA-sorbed MPs (Figure 3) to be persistent BPA release sources in natural

410

aquatic environments. Further, BPA released from LDPE and PC under physiological

411

conditions (Figure 2d) might be an ecological risk for marine organisms. The four typical

412

marine organisms were exposed to BPA released from LDPE, PC or BPA-sorbed MPs using

413

standard exposure doses and procedures using the data provided by literature studies, which

414

did not show noticeable toxicity (Table S7), however.70, 71 With only one exception, the BPA

415

concentrations released were well below the lowest observed effect concentrations (100 ng

416

L-1) found using Japanese medaka fish.72 Overall, the long term risks arising from the release

417

of chemicals from plastic particles are complex and deserve further study.

418

Environmental Implications 18


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Due to gradual increase in the quantity of plastic debris in marine environments,73 plastic

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debris may influence the fate and transport of HOCs and could pose ecological problems. In

421

this study, MPs were verified to be potential sinks and transport vectors for BPA in marine

422

environments. Buoyant plastic pellets easily sorb HOCs and enriched them in a surface water

423

microlayer.1 They may then transport them over long distances and affect the environmental

424

and biological systems. This study has shown that BPA molecules are then continuously

425

released into the surrounding environment. The dual-mode model of sorption nicely explains

426

BPA’s sorption by various polymer particles. It may be extended to predicting the sorption

427

and environmental risks of other related HOCs. These results may therefore facilitate

428

building an improved universal sorption model for HOC sorption which allows better

429

prediction of their transport and fate and any risk they pose to aquatic environments. These

430

results may also provide a foundation for evaluating the transport, fate and potential risk of

431

other plastic additives such as bisphenol sulfonyl and bisphenol fluorine.74 In real

432

environment, the effect of bacteria,22 weathering and fragmentation of MPs, as well as the

433

formation of biofilm on MPs certainly play vital role on the BPA sorption/desorption/release

434

on/from MPs. Therefore, it is important to explore further the mechanisms of BPA

435

sorption/desorption/release on/from MPs under conditions that are as close to the real

436

environment as possible.

437

Graphic abstract

438 439



ASSOCIATED CONTENT 19



Page 20 of 26

440

Supporting Information

441

The supporting information is available free of charge via the Internet at http://pubs.acs.org.

442

BPA detection methods, calculation of hysteresis index (HI), Temporal Patterns of BPA

443

Sorption on MPs, Table S1-Table S7, and Figure S1-Figure S5.

444

■ AUTHOR INFORMATION

445

Corresponding Authors

446

*E-mails:[email protected] and [email protected]

447

Note

448

The authors declare that they have no competing financial interest.

449

■ ACKNOWLEDGMENTS

450

This work was supported from Shanghai Natural Science Foundation (no. 19ZR1414900) and

451

China’s National Key Research and Development Programs (no. 2016YFC1402204). D. D.

452

Dionysiou also acknowledges support from the University of Cincinnati through a UNESCO

453

co-Chair Professor Position on “Water Access and Sustainability” and the Herman Schneider

454

Professorship in the College of Engineering and Applied Sciences. We kindly acknowledge

455

the reviewers and the editor for their comments that helped improve the quality of the MS.

456

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