Doctoral Dissertation
A study on the environmental fate of organic pollutants adsorbed onto plastics(プラスチックに吸着した有機汚染物質の環境運命に関する研究)

HASSAN MD NAZMUL
ハッサン エムデイナズムル
D174330

Department of Chemical Engineering
Graduate School of Engineering, Hiroshima University(広島大学大学院工学研究科 化学工学専攻)

A study on the environmental fate of organic pollutants adsorbed onto plastics(プラスチックに吸着した有機汚染物質の環境運命に関する研究)

Written by
HASSAN MD NAZMULSupervisor
Professor SATOSHI NAKAI

Polymer Technology Laboratory
Department of Chemical Engineering
Graduate School of Engineering, Hiroshima University(広島大学大学院工学研究科 化学工学専攻)

November 2021

Abstract
Plastic debris has become a serious issue because of its ubiquitous distribution and adverse effects onorganisms via pollution adsorption and ingestion. In the aquatic environment, floating plastic debris is
exposed to sunlight, and the adsorbed pollutants are photodegraded. Although plastic debris may age,plastic texture may affect the photodegradation behaviors of pollutants. Furthermore, if the density ofplastic debris exceeds that of seawater, it will settle and become buried in sediment. To consider the risk
derived from the adsorbed pollutants, whether the pollutants are photodegraded and whether plastics affectdegradation behaviors should be understood. In addition, possible pollutant migration from plastic debris
to sediment should be addressed because sediment contamination may cause harmful effects on benthic
organisms. This study photodegraded 2,3,4,5-tetrachlorophenol (TeCP) and pentachlorophenols (PCP) on
virgin and artificially aged polyethylene (PE), polyvinylchloride (PVC), and polyethylene terephthalate(PET) sheets to investigate effects of plastic textures on chlorophenol photodegradation behavior. On the
PE sheet, the photodegradation of TeCP and PCP was 1.50 times higher than that on the glass plate,thereby confirming the acceleration by PE. In contrast, the photodegradation rate was slower on PVC andPET than glass. These results confirmed that plastic textiles affect pollutant photodegradation behaviors.
Furthermore, PE accelerated the photodegradation of a polyaromatic hydrocarbon (PAH), naphthacene(NAP), and oxygenated NAP. Naphthacenequinone was detected at an abundant level on the PE sheetwith other oxygenated intermediates. Owing to the possibly higher toxicity of oxygenated PAHs than the
precursor PAHs, the results confirmed that the existence of pollutants together with degradationintermediates should be considered for the risk assessment. When the artificially aged PE, PVC, and PETplastics were used, photodegradation of the adsorbed 2,4,5-trichlorophenol (TriCP) was suppressed,
contributing to changes in the plastic textures such as the surface shapes and transparency. Finally, themigration of the aliphatic and aromatic organochlorines, TriCP, and 1-chlorooctadecane from the PE,
PVC, and PET sheets to sediment was confirmed. Dissolution followed by adsorption and directtranslation and direct translocation to sediment may be possible as migration mechanisms. These results
provide evidence that plastic debris can act as a carrier of pollutants to benthic ecosystems.

Contents
Page

Chapter 1. Introduction ………………………………………………………………. 1
1.1 Plastic pollution ………………………………………………………... 1-3
1.2 Production and distribution of microplastics …………………………... 3
1.3 Emergence of microplastics in creatures ………………………………. 6
1.4 Potential of microplastics as a carrier of pollutant ……………………... 7
1.5 Importance of understanding the fate of organic pollutants on plastics ... 10
1.6 Objectives and outlines of this study …………………………………… 11
1.7 Literature cited …………………………………………………………. 14

Chapter 2. Effect of Plastics on the Photodegradation Behavior of CPs …………….

24

2.1 Introduction …………………………………………………………….

24

2.2 Experimental …………………………………………………………..

26

2.2.1 Reagents and solvents …………………………………………..

26

2.2.2 Photodegradation ………………………………………………..

26

2.2.3 Analysis of the CPs and photodegradation products …………...

27

2.2.4 Characterization of the plastics ………………………………….

28

2.2.5 Statistical analysis ……………………………………………….

28

2.3.1 Photodegradation of the CPs ……………………………………

28

2.3.2 Factors that affect photodegradation ……………………………

32

2.3.2.1 PE ………………………………………………………

33

2.3.3 Possible effect of additives ……………………………………..

33

2.4 Conclusion ……………………………………………………………...

36

2.5 Literature cited …………………………………………………………

37

1

Chapter 3. Effect of PE on Photodegradation Behaviors of PAHs …………………..

41

3.1 Introduction …………………………………………………………….

41

3.2 Experimental …………………………………………………………...

42

3.2.1 Materials ………………………………………………………….

42

3.2.2 Photodegradation test of NAP and NAPQ ……………………….

42

3.2.3 Analyses ………………………………………………………….

45

3.3 Results And Discussion ………………………………………………...

45

3.3.1 Photodegradation of naphthacene (NAP) ……………………….

45

3.3.2 photodegradation intermediates produced from NAP …………..

48

3.3.3 Photodegradation pathways of NAP …………………………….

51

3.4 Conclusion ……………………………………………………………...

54

3.5 Literature cited ………………………………………………………..

55

Chapter 4. Effects of the Aging of Plastics on the Photodegradation of Adsorbed
Organic Pollutants ……………………………………………………………………

59

4.1 Introduction …………………………………………………………... 59
4.2 Experimental …………………………………………………………

61

4.2.1 Materials and chemicals ………………………………………..

61

4.2.2 Plastic sample preparation and additive extraction …………….

61

4.2.3 Photodegradation ……………………………………………….

62

4.2.4 Characterization of plastic sheets ………………………………

62

4.2.5 Analyses ………………………………………………………..

64

4.2.6 Statistical analysis ……………………………………………...

64

4.3 Results And Discussion ………………………………………………

64

4.3.1 Comparison of photodegradation rates of TriCP on virgin andaged plastics …………………………………………………...

642

4.3.2 Changes in surface textures …………………………………….

66

4.3.3 Suppression of photodegradation by the methanol extract …….

69

4.3.4 Changes in transparency ………………………………………

70

4.3.5 FT-IR analysis of aged plastic sheets …………………………..

71

4.4 Conclusion ……………………………………………………………

73

4.5 Literature cited ………………………………………………………

74

Chapter 5. Migration of adsorbed organochlorine from plastics to sediment ……….

78

5.1 Introduction …………………………………………………………..

78

5.2 Experimental …………………………………………………………

79

5.2.1 Materials ………………………………………………………

79

5.2.2 Sediment preparation …………………………………………

79

5.2.3 Migration of adsorbed organochlorine from plastic to sediment 80
5.2.4 Mechanism of the migration process ………………………….

81

5.2.5 Analysis of TriCP and ClOD ………………………………...

81

5.2.6 Data analysis …………………………………………………..

82

5.3 Results and Discussion ……………………………………………….

83

5.3.1 Migration of organochlorine …………………………………...

83

5.3.2 Effect of DOC on the migration of organochlorine ……………

85

5.3.3 Migration pathway ……………………………………………..

86

5.4 Conclusion ……………………………………………………………

87

5.5 Literature cited ………………………………………………………

88

Chapter 6. Conclusions ……………………………………………………………...

91

Acknowledgment …………………………………………………………… ………

95

List of publications ……………………………………………………………….......

973

Chapter 1

Introduction
1.1 Plastic pollution
Plastics and their products are widespread throughout the Earth. Chemically, plastics aresynthetic organic polymers with enormous social benefits and are used in health, safety,
energy-saving, material conservation, and a wide variety of sectors (Figure 1.1) [1], [2]. The
global demand for plastics has also increased, with an annual growth rate of around 9% [3].
Among the various types of plastics, the demand for polyethylene (PE), polypropylene (PP),polyvinyl chloride (PVC), polyurethane, and polyethylene terephthalate (PET) are the top five
single-use plastics (Figure 1.2) [2]. In addition, over 1/3 of the manufactured plastics in boththe United States and Europe are used to make disposable products, such as packaging
materials and trash bags, which are designed to be discarded within three years after theirproduction [4]. Due to their numerous uses and convenient disposal, plastics have gained
popularity, and every year a substantial proportion of plastics enter and persist in the marine
environment [5], [6]. Among the littered plastics, the most common identified plastics are PE,
PP, PVC, and PET [7]–[9]. Unwisely discarded plastic materials pose numerous threats to themarine environment. Due to the mismanagement of plastics, around 4.80 to 12.70 Mt of plastic
debris enters the oceans annually [5], [10], [11], [12]. The surging trend of plastics in theenvironment especially, the aquatic environment, is increasing problems and creating a
toxicological milieu for marine organisms.

1

others
12%

Packaging36%

Textiles14%
Industrial machinery1%
Consumer and institutional
products
10%Transportation
7%

Building andconstruction
16%

Electrical/Electronic
4%

Figure 1.1 Global plastic production by industrial sectors, 2015 [2].

PolyethyleneLow density(PE-LD)
19%

Others19%

Polystyrene (PS)6%

Polyethylene -Highdensity/Medium
density (PE-HD/MD)
12%

Polyethylene
terephthalate (PET)8%

Polypropylene (PP)19%

Polyurathane (PUR)
8%
Polyvinyl
chloride(PVC)
10%

Figure 1.2 Global plastic demand according to types [3].
2

At present, oceans are unexpectedly the reservoir of plastic debris on a global scale.
Approximately 60–80% of the world’s litter entering the ocean is in the form of plastics [13],and almost 10% of the world's annually produced plastics end up in the ocean [2]. Prominentroutes of plastics into the oceans are beaches and land-based sources like rivers, stormwaterrunoff, wastewater discharges, or the land litter transport by the wind [5], [10], [11]. The
Foresight Future of the Sea report determined that plastic pollution, set to increase threefold
between 2015 and 2025 without intervention, has a physical presence in the oceans and canaccumulate on the coasts or in particular areas of the sea [14]. Ellen MacArthur Foundation
reported that plastic in the oceans will outweigh fish by 2050 [15]. Although the negative
effects of plastics on the ecosystems are insufficiently understood, the health damage to sea
creatures, restriction on their movement [16], and pollution of beaches with accumulatinghydrophobic organic pollutants [17]–[20] have been extensively studied.

1.2 Production and distribution of microplastics
Plastic debris are defined based on their sizes and broken-down pieces (Figure 1.3). Andrady
et al. [21] defined microplastics as the barely visible particles that pass through a 500-μm sieve
and retained by a 67-μm sieve (0.06–0.5 mm in diameter); particles larger than this were called
mesoplastics. Betts et al. [22], Eriksen et al. [11], and Fendall and Sewel [22] defined
microparticles as being in a size range < 5 mm (recognizing 333 μm as a practical lower limit
when neuston nets are used for sampling). However, plastics debris having dimensions ranging
from a few μm to 500 μm (5 mm) are commonly found in seawater [4], [21].

3MacroMeso≥ 25 mm< 25 – 5 mm

Microplastics (MPs)Mini-micro< 5 – 1 mm

< 1 mm – 1 𝜇m

Figure 1.3 Size-based definition of plastics. (Image source:(https://www.naturalblaze.com/?s=microplastics), access on October 7, 2021).
Microplastics are also categorized based on manufacturing sources, such as i) primary
microplastics and ii) secondary microplastics. Primary microplastics are purposely
manufactured for various applications, such as exfoliants (microbeads) in personal careproducts [24]. Secondary microplastics are fragments of larger plastic debris, including plastic
containers, nets, line fibers, films, and tires [5]. The generation of plastic waste is increasing
continuously (Figure 1.4) [2], [11], with an estimation reported that more than 5.25 trillion
pieces of plastics weighing over 268,940 tons are floating on the ocean surfaces [11]. These4

plastics are continuously degraded on ocean surfaces and at shorelines [25]. The degradationof plastics occurs through chemical changes, resulting in reduced average molecular polymer
weight. As the mechanical integrity of plastics invariably depends on their average molecularweight, any significant degradation extent inevitably weakens them [21]. For example,
exposure to ultraviolet (UV) radiation catalyzes the photo-oxidation of plastics, making themmore brittle. In addition, they undergo degradation by chemical and physical processes in theenvironment, i.e., by the action of wind and abrasion by waves [16]. Plastics become brittle
enough to fall apart into powdery fragments, often invisible to the naked eye. However, atpresent, nano plastics are also increasingly being manufactured and being used in paints,
adhesives, drug delivery vehicles, electronics, and other product types [26]. The size reduction
due to uses and environmental degradation may induce unique particle characteristics andinfluence their potential toxicity.
Microplastics are driven by ocean winds, currents, river outflow, and drifting plastic debris toreach the remote islands and the benthic and pelagic habitats of oceans such as the Pacific, theArctic, and the deep oceans [4]. For example, the estimated concentrations of microplastics atthe surface of seas were 4.38×104–1.46×106 pieces/km2 in the Arabian Bay [27], 640–42000
pieces/km2 in the northwestern Pacific Ocean [28], 1051 pieces/m 3 in South Korea coastal
waters (the sampling areas are Busan, Ulsan, Yeongil, Gwangyang, Deungnyang city, and threerural areas) [29], and 0–1.31 and 0–11.5 pieces/m3 on the Norwegian coast (NIWR, 2017) andsubsurface of artic water [30]. In addition, microplastics have been found in sediment samples
such as in the Yellow Sea, China at depths of < 20 m, 20–40 m, 40–60 m, and 60–80 m, withmicroplastic concentrations of 1765, 2135, 2346.7, and 2771.3 pieces/kg (dry weight),
respectively [19].

5Figure 1.4 Surging trend of plastic waste generation [2].

1.3 Emergence of microplastics in creatures
Ingestion of microplastics by wildlife is a frequently noted problem, with plastic debris foundin the guts of seabirds [20], [31], fishes [32], [33], shrimp [34], turtles, and whales [35]. Plastic
debris causes the death of marine organisms in different ways, such as entanglement andblockage of the organism’s digestive system [31], [36], [37]. Ingestion of microplastics has
also been documented in deposit- and suspension-feeding sea cucumbers [5], deposit-feeding
lugworms [38], detritivores amphipods [39], and even zooplankton [40] (Figure 1.5).

6

Figure 1.5 Overall focused point where microplastics gained concern [41].

1.4 Potential of microplastics as a carrier of pollutants.
Besides the main polymers, most plastics contain several types of additives to improve plastic
features like ductility, hardness, durability, or resistance to weathering. Among these, someplastic additives, including plasticizers, are suspected as endocrine disruptors [42]–[44].
Various kinds of plasticizers are used to modify the properties of plastics, reaching up to 10–
70% of the plastic material weight. Plasticizers are mostly phthalates, terephthalates, epoxies,
aliphatic, trimellitates, and citrates [7], [45]. These are not chemically bonded with the structureof the main plastics and can, therefore, leach out into the environment [45]. As plastics areaccumulated and fragmented in oceans, the leaching of plastic additives also represents anothertoxicological risk in the marine environment.
Microplastics transport hydrophobic organic chemicals (Figure 1.5) (HOCs) [41], [46], [47].
For example, polycyclic aromatic hydrocarbons (PAHs) [48], organochlorine pesticides [49],7[50], polychlorinated biphenyls (PCBs) [51], [52], and chlorophenols (CPs) [49] are adsorbed
and concentrated on microplastic surfaces [53], [54]. In addition, they also accumulate heavy
metals such as cadmium, zinc, nickel, and lead [55]. A summary of the adsorption of HOCs on
plastics is shown in Table 1.1.
Table 1.1: Adsorption of pollutants on MPs.
Pollutants namePlastics

Experimental

Adsorption

Reference

-

[24]
[33]

Sample/ Medium
Phenanthrene and DDTPE

Artificial seawater

PAHs

PE, PP, Environmental

0.063–0.208

PS

µg/g

sample(San Diego Harbor)

PCBs

LDPE

Environmental

3.4-35 ng/g

[33]

Aqeuous solutions

-

[48]

River water

126-144 μg/g

[49]sample(Oceansurface)
PAHs

LDPE
HDPE

2,4,6-trichlorophenol

PP(TriCP) andPE

53.7-99.5 μg/g

pentachlorophenol (PCP)
Phenanthrene and DDTPE,

Seawater-[50]

Seawater

1.15–15.5 µg/g

[50]Environmental

4-117 ng/g

[53]

biphenyls (PCBs)

samples0.16-3.1 ng/g

Dichlorodiphenyl(Collectedtrichloroethane (DDT)Japanese coasts)

PVC,
uPVC
PAHs

PE
PVC

PolychlorinatedPP

Nonylphenols

8from 0.13-16 μg/g

Pollutants namePlastics

Experimental

Adsorption

Reference-47 ng/g[56]

Sample/ Medium
PCBs

PS, PS- Environmental

PAH

foam

DDT

samples (San Diego 30-1900 ng/g
beaches, California)

PAH

76 ng/g300-1900 ng/g

PAH

PS-foam

Seawater

240-1700 ng/g

[56]

Phenanthrene

PE, PP

Seawater

0.6-6.1 µg/L

[57]

Artifical seawater

303.03–714.29

[58]

PVC
PAHs

PE,
PVC,

µg/g

PS
Fuel aromatics

PE

Artificial seawater

1-10 µg/g[59]

Fresh and seawater

-

[60]

-

[61]

PVC
PS
17 PCBs

PE,PS

Hexachloro

PE, PP, Seawater

cychlohexanes

and PS

chlorinated benzenes

Apart from being adsorbed microplastics, HOCs are further ingested by animals, resulting intheir transmigration to the food chains (Figure 1.5) [41], [62]. A recent research paper reported
that the coexistence of polystyrene and nickel exposed toxicity on Daphnia magna, losing
mobility [18], [63]; moreover, polystyrene altered the bioavailability [18], [64] andphenanthrene transfer into Daphnia magna biota cells. In addition, microplastics have been
detected in a range of marine biota, including seabirds, crustaceans, and fish [32], [65]. Ma etal. [18] reported that the significant bioaccumulation of phenanthrene occurred by feeding
phenanthrene-adsorbed polystyrene to Daphnia magna. These results are evidence thatmicroplastics can act HOC carriers to organisms. In addition, the harmful effects of HOCs are9also well known. For example, cancer, allergies and hypersensitivity, damage to the central andperipheral nervous systems, reproductive disorders, and disruption of the immune system havebeen reported [66], though the endpoints are dependent on the species of HOCs. In addition,some HOCs are also considered to be endocrine disrupters, causing health problems including
hormonal systems, dysfunctional immune and reproductive systems, greater susceptibility todisease, and the diminished intelligence of exposed individuals and their offspring [67].
Therefore, the behaviors of HOCs adsorbed onto microplastics and the negative impacts onorganisms should be understood for environmental management.

1.5 Importance of understanding fates of organic pollutants on plastics
The fate of HOCs and their distribution among different environmental sectors (Figure 1.5)
are determined via various physical, chemical, and biological processes [68]. HOC degradation
occurs by chemical, biological, and/or their combined mechanisms [69], [70]; the former
occurs via photochemical, redox, and hydrolysis reactions [69]. Although photodegradation
can occur only in the photic zone, this is an essential HOC degradation mechanism during thetransport of microplastics.
On the other hand, plastics may potentially alter the photodegradation behaviors of pollutants
adsorbed onto microplastics. A previous study showed that the photodegradation ofpolystyrene films containing brominated flame retardants, such as decabromodiphenyl ether;tetra-bromophenol A produced 14 brominated photodegradation products via the brominationand oxidation of polystyrene [71]. This indicates the necessity of investigating the possible
effects of plastics on HOC photodegradation behaviors from the viewpoint of risk assessment.
Although many studies have investigated the photodegradation behaviors of HOCs [69], [72];10however, no information is available about the effect of plastic on the photodegradation of theadsorbates.
All plastics age in the environment, changing the plastic’s properties and adsorption capacityfor HOCs [73], [74]. Aging increases the crystallinity of plastics; however, crystallinity is nota controlling factor of HOC adsorption on plastics [75]. Along with changes in the plastic
property, the aging of plastic generates radicals that may also change the fate of pollutants [71].
This implies that aged plastics might show different effects on the photodegradation of HOCs
compared to virgin ones. Therefore, the effects of plastic aging on photodegradation behaviorsof adsorbed HOCs need to be addressed because of the possible production of more toxic
pollutants than the precursors.
Furthermore, once the density of plastic debris reaches that of seawater by the formation of
biofilms, these settle in the sediment (Figure 1.5) [76]. Once buried, photodegradation does
not occur. However, HOCs adsorbed onto microplastics might be distributed to sedimentbecause various types of HOCs have been detected in the marine sediment [52], [77], [78].
Although microplastics are potential HOC carriers to the benthic ecosystem, no study has
investigated the migration of HOCs from microplastics.

1.6 Objectives and outlines of this study
For understanding the fates of HOCs adsorbed onto the surface of microplastics in the oceans,the effects of plastics and their aging on the fates of HOCs should take priority to be elucidated
for risk assessment. During floatation, microplastics are exposed to sunlight. As plastics canalter the photodegradation pathways of HOCs, the alteration effects of plastics should beunderstood. In addition, microplastics are expected to be a HOC carrier to the benthic
ecosystem. Therefore, this research focused on the photodegradation processes of11

organochlorines adsorbed onto plastics and migration phenomena from plastics to sediment.
The objectives of this research were as follows:
1. To reveal the effects of plastic textures on photodegradation behaviors of HOCs.
2. To investigate how the aging of plastics changes the photodegradation behaviors ofadsorbed HOCs.
3. To determine the migration behaviors of adsorbed organic pollutants from plastic
surfaces to marine sediment.

This thesis consists of 6 chapters, as follows
In Chapter 1, the background and negative consequences of plastics and the importance of
research about the fate of adsorbed organic pollutants onto plastics are described. The
description of pollutants adsorption onto microplastics, microplastic numeration, concentrationin the environment, and their toxic effect are presented as in summarized form.
In Chapter 2, the photodegradation behaviors of organochlorine TeCP and PCP on the plastics
and glass were investigated, and the plastic effects on the photodegradation of organochlorines
are discussed.
In Chapter 3, the photodegradation behaviors of NAP as a model PAHs on PE surface and glass
were compared to investigate whether PE with hydrogen donation potential accelerates thephotodegradation of NAP and or generates new persistence toxic intermediates.
In Chapter 4, the photodegradation behaviors of TriCP on artificially aged plastics were
investigated to confirm whether the aging of plastics affects the photodegradation behaviors ofadsorbed TriCP and to find causes for the aging effects.

12

In Chapter 5, migration of organochlorine from plastics to sediment was examined by burying
organochlorine-adsorbed plastic sheets into wet sediments, and migration behaviors of organic
pollutants from the plastics to sediments were discussed.
In Chapter 6, the general conclusions of all chapters are presented, with recommendations for
future study.

Chapter 1: Introduction

Chapter 3: Effect of polyethylene on
photodegradation behaviors of polycyclicaromatic hydrocarbon

Chapter 2: Effect of plastics on thephotodegradation
behaviorofchlorophenols

Chapter 4: Effects of the aging of plastics on the photodegradation of a chlorophenol

Chapter 5: Migration of adsorbed organochlorines from plastics to sediment

Chapter 6: Conclusions

Figure 1.6 The outline of this research

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14

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23

Chapter 2

Effect of plastics on the photodegradation behavior of chlorophenols

2.1 Introduction
The marine environment is sensitive to pollution because of its vulnerable ecology. Plastic
debris is a new environmental threat to the marine environment because of its adsorption
behaviors [1], [2], and littering sizes, including macro-to-micro and nanosized particles [3].
However, these fragmented plastics are enumerated because of their ubiquitous abundance andfloatation on ocean surfaces [4]–[6], they have negative effects on the marine environment andbiota [7]–[9]. The presence of microplastic and nanoparticles in the environment may haveeffects on the fate of HOCs, with HOCs adsorbed on their surfaces. The most commonly
detected pollutants on plastic litters are PAHs [2], PCBs, hexachlorobenzene (HCB),
dichlorodiphenyltrichloroethane (DDT) [10], and CPs [11]. As HOCs, such as CPs and PCBs,
are persistent in the environment and act as endocrine disruptors or mutants [12], the fates ofhydrophobic pollutants adsorbed on microplastics should be clarified from a risk assessment
viewpoint.
The photodegradation of any substance in the natural environment is a combination of reactionsof all types of photochemical processes depending on the absorption level, radiation
wavelength, quantum yield, radiation exposure duration, and physical state of the transforming
compound [13]. For persistent halogenated organic pollutants, photodegradation could be amajor cause of transformation on the sea surface. The shortest cutoff wavelength of sunlighthitting the earth’s surface (~290 nm) can cause the dechlorination and degradation of PCP and
PCBs [14].
Many studies have investigated the photo degradability of organohalogen pollutants in various
media, such as water [13] and organic solvents [14]–[16], using UV lamps as the light source.
24Interestingly, HCB photodegradation experiments in n-hexane, isopropanol, and methanol
showed the photodegradation rate in n-hexane was the highest [16], while the photosensitizeddechlorination of HCB was observed with the coexistence of a polymeric carbazole
chromophore [17]. Moreover, Minghui et al. reported chlorination by a photolysis process in
CCl4 by confirming the production of Hexa-, hepta-, and octachlorobenzo-p-dioxin (HxOCDDs) from 1,2,3,7,8-pentachlorodibenzo-p-dioxin (PeCDD) [15]. These results confirmed
the matrix effect in the photodegradation process, indicating the need to understand how
plastics affect the transformation of organohalogen pollutants.
Most studies on the photodegradation of organohalogen pollutants have revealed the
production of dehalogenated intermediates. For example, PCP photodegradation provided
mono, di-, tri-, and tetra-CPs as intermediates [13], [16]. Moreover, in the processing of plastics,
manufacturers use halogenated compounds as additives. A recent study showed thatpolystyrene impregnated with brominated flame retardants, such as decabromodiphenyl ether
and tetrabromobisphenol, produced 14 new brominated products by oxidation and brominationof polystyrene under UV irradiation [18]. Considering the above-mentioned matrix effect on
the photodegradation of organohalogen pollutants, it can be surmised that the characteristics ofplastics may also affect the photodegradation behavior of organohalogen pollutants. However,the photodegradation of organohalogen pollutants sorbed on plastics has not been investigated.
In this study, the photodegradation behavior of organochlorine pollutants adsorbed on threecommonly used plastics, PE, PVC, and PET was compared to investigate the matrix effect ofthe plastic. Furthermore, the mechanisms of the matrix effect are discussed based on the plastic
characteristics and additives.

25

2.2 Experimental
2.2.1 Materials
Virgin sheets of PE, PVC, and PET with a thickness of 1.0 mm were purchased from Sanplatec
Corp. (Tokyo, Japan). Before the photodegradation experiments, each plastic sheet was cut intoa circle shape with a 26-mm diameter. As organochlorine pollutants, PCP and TeCP were tested.
As the pure PCP reagent was not available, a mixture of 10% PCP and 90% TeCP, TriCP (>95% purity), and 2,5-dichlorophenol (DCP, > 98% purity) was obtained from Tokyo Chemical
Industries (Tokyo, Japan). Sodium dihydrogen phosphate (NaH2PO4, 99% purity) andphosphoric acid (H3PO4, 85% purity) were purchased from Wako Chemical (Japan). Methanol(99% purity) and n-hexane (97% purity) were purchased from Nacalai Tesque Inc. (Japan).

2.2.2 Photodegradation
The circle-shaped plastic sheets were placed at the bottom of quartz glass Petri dishes (diameter
26 mm), and an n-hexane solution containing 10 mg/L PCP and 90 mg/L TeCP was poured onthem (Figure 2.1). After allowing 60 min for the evaporation of n-hexane at 25 °C, the Petridishes were covered with quartz glass lids and then placed in a Navis Schreiner photochemical
reactor (DM-05, Daishin Kogyo Co. Ltd., Hiroshima, Japan) equipped with a 10 W lowpressure UV lamp emitting 254 nm wavelength UV light at 32 µW/cm2 (Figure 2.1). During
UV irradiation for 4 h, the photochemical reactor was kept at 25 °C, and three samples wererandomly sacrificed to analyze the remaining PCP, TeCP, and the photodegradation products.
A control experiment was performed under the shading of UV radiation by wrapping the Petridishes with aluminum foil.

26

UV light source

Evaporation of solvent

Cover with a quartz glass lid

UV irradiation

Figure 2.1 Experimental setup for photodegradation of chlorophenols on the plastic surfaces.

2.2.3 Analysis of the CPs and photodegradation products
To measure the remaining concentrations of the CPs and photodegradation products, methanol
was poured into the Petri dish for extraction. The methanol extract was mixed with a 20 mM
phosphate buffer (pH 2.50) solution at a volume ratio of 4:1. A sample of the mixture (100 μL)
was then analyzed by a liquid chromatograph (HPLC) on an octadecyl silica column (Cosmosil
5C18-AR-II, 4.6 mm × 150 mm, Nacalai Tesque Inc., Japan) at a flow rate of 1.0 mL min −1,
using a mixture of methanol and phosphate buffer (4:1 volume ratio) as the mobile phase. The
CPs were detected by a UV–visible detector at 254 nm (SPD-10A, Shimadzu, Japan).
The photodegradation products were analyzed by liquid chromatography-mass spectrometry(LTQ ORBITRAP XL-20930, Fisher Thermo Scientific, USA) using a high-resolution orbitrap
mass spectrometer. The same Cosmosil 5C18-AR-II column was used, while elution was
performed with a binary solvent in an isocratic system consisting of methanol and acidified
water (0.1% v/v HCOOH) in a 90:10 volume ratio. The flow rate was set at 0.50 mL min −1,
and the injection volume was 10 μL. The scan parameters were from m/z = 90 to 800 amu(atomic mass unit) in full-scan negative electrospray ionization mode. For identification, themass spectrum patterns were compared with those of the authentic TriCP and DCP samples.
27

2.2.4 Characterization of the plastics
To analyze the functional groups of the plastic surfaces, the UV-irradiated samples wereanalyzed by a Fourier transform infrared (FTIR) spectrometer (IRPrestige-21, Shimadzu,
Japan) equipped with a tungsten-iodide light source. The spectra were recorded from 400 to
4000 cm−1 (nominal resolution of 4 cm−1, 32 scan summation).

2.2.5 Statistical analysis
One-way analysis of variance (ANOVA) followed by the Tukey test was performed for theremaining PCP and TeCP concentrations to determine the statistical significance as expressedby a p-value for photodegradation enhancement or suppression on the assumptions of a normal
probability distribution and homogeneity of variance.

2.3. Results and Discussion
2.3.1 Photodegradation of the CPs
The change of the PCP concentration on the PE sheet under UV irradiation is shown in Figure
2.2. The PCP concentration decreased from 1.75 to 0.04 mg/cm 2 after 240 min of UV
irradiation. As the control experiment did not show a change in the PCP concentration, this98% decrease in the PCP concentration was attributed to photodegradation. Although TeCP
could be produced by PCP photodegradation [13], only 4.12% of the initially adsorbed TeCP
remained after UV irradiation, whereas 98% of TeCP remained after 240 min in the control
experiment (data not shown). This confirmed that PCP and TeCP adsorbed on the PE sheet
were photodegraded by UV irradiation. This provided the first evidence for the occurrence ofphotodegradation of PCP and TeCP on the plastic sheet.
To investigate whether the plastics affected the photodegradation of PCP and TeCP, thephotodegradation rate constants were determined based on pseudo-first-order reaction kinetics

28

(Figure 2.3(a) and 2.3(b)). Except for PVC, the photodegradation rate constants were
estimated at a coefficient of determination (R2) > 0.85. From 30 to 120 min, the PCP
concentration on PE was lower than that on glass (p < 0.0013, Figure 2.3(a)). After 240 min,a significant difference was not observed for this dataset because of the large variance, but the
average value on PE was lower than that on the glass. The photodegradation rate constant of
PCP on the PE sheet was 1.80  10−2 min−1, 1.50 times higher than that on glass (1.22  10−2
min−1).
Although the concentration of TeCP might include the fraction produced fromphotodegradation of PCP by dechlorination [13], the remaining TeCP concentration on PE waslower than that on the glass after 120 (p = 0.0764) and 240 min (p = 0.0409), and thephotodegradation rate constant of TeCP on PE was higher than that on glass (Figure 2.3(b)).
These results confirmed the acceleration of photodegradation of PCP and TeCP by the PE sheet.
The degradation products TriCP, DCP, and CPs were detected, as shown in Figure 2.4. As
these products can be produced by reductive dechlorination [14], [16], PE may act as ahydrogen donor (Figure 2.6).

2.5PEPE (UV)PCP [mg/cm2]

2.0
1.5
1.0
0.5
0.0
0

30

60

90
120
150
180
UV irradiation time [min]

210

240

Figure 2.2 Degradation of PCP adsorbed on the PE sheet with and without UV irradiation. The
bars indicate the standard deviation (n = 3).
29

The coefficients of determination for PCP and TeCP estimated using second-order reaction
kinetics were 0.95 and 0.97, respectively. These were high compared to that of PCP (0.86) and
TeCP (0.93), also estimated using first-order reaction kinetics (Figures 2.3(a), 2.3(b));
however, we used the photodegradation constants of PCP and TeCP estimated using first-order
reaction kinetics for a comparison between PET and glass. The PCP and TeCP concentrations
on PET were significantly higher than those on glass after 60 min (p < 0.0221, Figure 2.3(a);p < 0.0138, Figure 2.3(b)), except for TeCP after 120 min (p = 0.133). The photodegradation
rate constants on the PET sheet were 8.10  10−3 min−1 for PCP and 7.00  10−3 min−1 for TeCP,
lower than those on glass (1.22  10−2 min−1 for PCP and 1.09  10−2 min−1 for TeCP). This
showed a photodegradation suppression by the PET sheet. A possible reason for this is adecrease of UV radiation reaching PCP and TeCP owing to sorption by the PET sheet and/or
additives, as discussed in Section 2.2.3.
Surprisingly, PCP on the PVC sheet showed a different response under UV irradiation, where
the photodegradation rate slowed and then stopped after 60 min (Figure 2.3(a)). A possible
reason for the PCP behavior on PVC is the rechlorination of the intermediates by chlorineradicals produced from PVC. If rechlorination occurs, TeCP might transform into PCP and
TriCP. A previous study confirmed the production of Hx-OCDDs from PeCDD by photolysis
in CCl4 [15], indicating the occurrence of chlorination. Additionally, the production of chlorineradicals during the deterioration of PVC under UV irradiation has been reported [19].
As for TeCP, the concentration on PVC was lower than on the other plastics and glass during
the first 120 min (p < 0.0285, Figure 2.3(b)), and this did not contradict the possible
occurrence of rechlorination. However, the degradation behavior did follow second-order
reaction kinetics (R2=0.97) better rather than first-order reaction kinetics (Figure 2.3(b)). The
possible reasons for this difference in suitable reaction order kinetics might be different
photodegradation mechanisms.
30

Ln(C/C0) [-]0

kPVC= 0.0046
R² = 0.86-1

kPET = 0.0081
R² = 0.86-2kGL = 0.0122
R² = 0.94

-3

Glass

PE

PET

PVC

kPE = 0.018
R² = 0.85-40kPET = 0.007
R² = 0.93

Ln (C/C0) [-]-1-2

kGL = 0.0109
R² = 0.96-3kPVC = 0.0153
R² = 0.42

Glass
PET

PE
PVC

kPE = 0.0156
R² = 0.96

-4
0

30

60

90
120
150
180
UV irradiation time [min]

210

240

270

Figure 2.3 Photodegradation of (a) PCP and (b) TeCP on PE, PET, and PVC. Bars indicate thestandard deviation (n = 3).

31

For example, photodegradation and rechlorination might consume TeCP on PVC if
rechlorination occurs.
These findings confirmed that the photodegradation behavior of organochlorine pollutants is
affected by plastics, although the effect might depend on the plastic characteristics andadditives. To investigate the different effects of plastics, the changes in the plastic molecular
structures by UV irradiation were analyzed.

0.00 - 29.98
6.64

100
80

0.60

1.84

6.80

5.49

3.67

8.92 10.13

14.03

10.83

22.70

25.68 26.02 26.65

29.13

22.23 23.07

25.66 26.65

29.04

15.81 17.18 17.43 19.84 20.54

60
40
20
0
100

9.35

PCP

80
60

Relative abundance

40
20

6.55 7.30

0.92 2.11 3.01 4.16

0
100

8.96

9.68 10.76

12.97

14.53 15.63 16.89

18.96 20.23

6.62

80

TeCP

60
40
20

0.90 1.93

0
100

6.87

3.26 4.66 5.47
5.47

80

7.32 9.03 10.81

12.45 13.54 14.55 15.95

18.47 19.51

22.10

23.81

25.52 26.58

29.44

TriCP

60
40
20

0.76

0
100

2.68

5.06

4.84

5.72 6.17 8.40

10.00 10.67

12.77

14.66 16.08

17.75

20.23

22.07

24.69 25.39 27.35 28.68

4.52

80
60

DCP

40
1.66 2.20 4.39

20

5.04

5.58

8.92

10.76

12.41 13.72 15.61 17.07 17.86

25.66
23.81 25.05
25.99 28.29

20.09 20.83

0
0

2

4

6

8

10

12

14
16
Time (min)

18

20

22

24

26

28

Figure 2.4 Total ion chromatograms of the CPs formed by UV irradiation of PCP and TeCP
on the PE sheet.

2.3.2 Factors that affect photodegradation
Photolysis and photo-oxidation are mechanisms of polymer photodegradation. Polymer
radicals are formed by UV irradiation and readily react with oxygen molecules to form peroxy
radicals [20]. Peroxy radicals abstract hydrogen from polymers to form hydroperoxide groups32

(R–O–O–H), dissociated by UV irradiation or heat [21]. Finally, carbonyl groups are formed
and give different dissociative intermediates and products according to the Norrish I and II
mechanisms [22]. The Norrish I mechanism involves radical scission of the bond adjacent to
the excited carbonyl group with the formation of two radicals leading to terminal alcohol andcarboxyl acid groups. The Norrish II mechanism involves intramolecular hydrogen abstraction
from the -carbon atom with chain scission without producing radicals to give an unsaturated
polymer chain end and a polymer chain with an end carbonyl group [22], [23].
The roughness of the plastic surface was expected to change the photodegradation rate of CPs
at some degrees. As discussed here, the deterioration of the PP, PET, and PVC surfaces wasobserved in FT-IR analyses, suggesting the change of surface roughness during thephotodegradation. As the surface roughness of plastic samples was not measured, this paper
addressed the effects of the chemical characteristics of the plastics on the photodegradation ofCPs.

2.3.2.1 PE
Photodegradation of PCP and TeCP was accelerated on the PE sheet (Figure 2.3(a) and 2.3(b)).
Owing to photodegradation by reductive dechlorination, the possible acceleration mechanism
might be hydrogen donation from the PE sheet (Figure 2.6).
Hydrogen and alkyl radicals and carbonyl groups are formed on PE by UV irradiation owing
to structural defects or chromophore impurities. These radicals are formed by C–C bondscission, hydrogen abstraction, double bond formation, and crosslinking of polymer radicals[20]. In terms of the Norrish I mechanism, the photolysis of the carbonyl group in PE involves
the scission of α and β C–C bonds after electronic excitation of the C=O group to formhydrogen and alkyl radicals and ketones. In terms of the Norrish II mechanism, the crosslinking

33

of macroradicals removes hydrogen atoms from the PE sheet following -/-hydrogen
abstraction [24].
The FT-IR spectra of the virgin PE sheet and the PE sheet irradiated by UV light for 240 min
are shown in Figure 2.5. After UV irradiation, there were increases in the bands representing(i) C=C disubstituted (cis) unsaturated bonds at 730 cm−1 [25], (ii) CH2 asymmetric stretchingat 1440 cm−1, and (iii) CH2 symmetric band stretching or CH2 end groups at 1360–1300 cm−1[23], [26]. Moreover, the bands for oxygen-containing groups at 3200–3600 cm −1 and carbonylgroups at 1728 cm−1 revealed that oxidation occurred, and the broad weak band at 2685 cm −1for olefinic CH2 stretching supported hydrogen abstraction from the carbon backbone orcrosslinking of macromolecules [24].

5
PE virgin

PE UV Experimental

4

C=C

Absorbance

CH2

CH2

3
2

OH
> CO

1
0
4000

3600

3200

2800

2400

2000

1600

1200

800

400

Wavenumber [cm-1]
Figure 2.5 FT-IR spectra of the virgin PE sheet and the PE sheet irradiated by UV light for
240 min.

34

O

hv

2

n
+

hv

H

OH +

Figure 2.6 Photodegradation mechanism of PE [27].

2.3.3 Possible effect of additives
Generally, chemical additives, such as plasticizers, antioxidants, and softeners, are added toplastic products to extend their lifetime, prevent oxidation, and improve their stability [27]. For
example, UV absorbers, such as hindered amine light stabilizers, are often used to obtain
resistance to weathering. Antioxidants inhibit the propagation phase of photo-oxidation [28]and impede the reaction. Based on these properties, the photodegradation process of CPs canbe affected by plastic characteristics and additives. On PVC, the lowest TeCP concentration
was observed among the four media during the first 60 min; however, PVC might include
additives at the highest level (Figure 2.7). A plausible explanation for these results might be
the inclusion of additives inert to photodegradation in addition to rechlorination; however, theadditives have not yet been identified. Future studies should investigate the effect of additives
on the photodegradation of organic pollutants and rehalogenation to clarify the fate and risk
assessment of organic pollutants in microplastics.

35

Integraded area [105]335.19
350
300
250
200
150

0.69

100

23.73

50
0
PE

PET

PVC

Figure 2.7 Total areas of the unknown peaks surmised as additives for the virgin plastics
detected by HPLC for the plastic matrices.

2.4 Conclusions
The photodegradation of PCP and TeCP by dehalogenation of sorbed CPs on PE, PET, and
PVC sheets was confirmed. Photodegradation of the CPs was accelerated by PE because of
possible hydrogen donation, whereas PET suppressed photodegradation. A possible reason is
UV absorption by the surrounding PET additives. For PVC, the photodegradation reaction was
suppressed owing to possible rechlorination because chlorine atom migration was observed in
the PVC sheet. This study suggests that plastic characteristics should be considered to assess
the fate of organic pollutants sorbed on plastics through photo transformation.

36

2.5 Literature cited[1]

Guo X. and J. Wang, “The chemical behaviors of microplastics in marine environment:
A review,” Mar. Pollut. Bull., vol. 142, pp. 1–14, May 2019.

[2]

Wang J., X. Liu, and G. Liu, “Sorption behaviors of phenanthrene, nitrobenzene, andnaphthalene on mesoplastics and microplastics,” Environ. Sci. Pollut. Res., vol. 26, no.
12, pp. 12563–12573, Apr. 2019.

[3]

Alimi O. S., J. Farner Budarz, L. M. Hernandez, and N. Tufenkji, “Microplastics andNanoplastics in Aquatic Environments: Aggregation, Deposition, and Enhanced
Contaminant Transport,” Environ. Sci. Technol., vol. 52, no. 4, pp. 1704–1724, Feb. 2018.

[4]

Eriksen M. et al., “Plastic Pollution in the World’s Oceans: More than 5 Trillion Plastic
Pieces Weighing over 250,000 Tons Afloat at Sea,” PLoS ONE, vol. 9, no. 12, p. e111913,
Dec. 2014.

[5]

Sebille van E. et al., “A global inventory of small floating plastic debris,” Environ. Res.
Lett., vol. 10, no. 12, p. 124006, Dec. 2015.

[6]

Wilcox C., B. D. Hardesty, and K. L. Law, “Abundance of Floating Plastic Particles Is
Increasing in the Western North Atlantic Ocean,” Environ. Sci. Technol., vol. 54, no. 2,
pp. 790–796, Jan. 2020.

[7]Davidson K. Davidson and S. E. Dudas, “Microplastic Ingestion by Wild and Cultured
Manila Clams (Venerupis philippinarum) from Baynes Sound, British Columbia,” Arch.
Environ. Contam. Toxicol., vol. 71, no. 2, pp. 147–156, Aug. 2016.

[8]

Mato Y. Mato, T. Isobe, H. Takada, H. Kanehiro, C. Ohtake, and T. Kaminuma, “Plastic
Resin Pellets as a Transport Medium for Toxic Chemicals in the Marine Environment,”
Environ. Sci. Technol., vol. 35, no. 2, pp. 318–324, Jan. 2001.

37

[9]

Ziajahromi S., A. Kumar, P. A. Neale, and F. D. L. Leusch, “Environmentally relevant
concentrations of polyethylene microplastics negatively impact the survival, growth andemergence of sediment-dwelling invertebrates,” Environ. Pollut., vol. 236, pp. 425–431,
May 2018.

[10] Bakir A., I. A. O’Connor, S. J. Rowland, A. J. Hendriks, and R. C. Thompson, “Relative
importance of microplastics as a pathway for the transfer of hydrophobic organic
chemicals to marine life,” Environ. Pollut., vol. 219, pp. 56–65, Dec. 2016.
[11] Tubić A. et al., “Significance of Chlorinated Phenols Adsorption on Plastics andBioplastics during Water Treatment,” Water, vol. 11, no. 11, p. 2358, Nov. 2019.
[12] Sun W. et al., “Combined effects of pentachlorophenol and its byproducthexachlorobenzene on endocrine and reproduction in zebrafish,” Chemosphere, vol. 220,
pp. 216–226, Apr. 2019.
[13] Czaplicka M., “Photo-degradation of chlorophenols in the aqueous solution,” J. Hazard.
Mater., vol. 134, no. 1–3, pp. 45–59, Jun. 2006.
[14] Zhou Z., Y. Zhang, H. Wang, T. Chen, and W. Lu, “The Comparative Photodegradation
Activities of Pentachlorophenol (PCP) and Polychlorinated Biphenyls (PCBs) Using UVAlone and TiO2-Derived Photocatalysts in Methanol Soil Washing Solution,” PLoS ONE,vol. 9, no. 9, p. e108765, Sep. 2014.
[15] Minghui Z., B. Zhicheng, X. Xiaobai, and W. Keou, “Mechanism of photodegradation ofpolychlorinated dibenzo-p-dioxins in carbon tetrachloride,” Chemosphere, vol. 32, no. 4,
pp. 603–607, Feb. 1996.

38

[16] Yamada S., Y. Naito, M. Takada, S. Nakai, and M. Hosomi, “Photodegradation ofhexachlorobenzene and theoretical prediction of its degradation pathways using quantum
chemical calculation,” Chemosphere, vol. 70, no. 4, pp. 731–736, Jan. 2008.
[17] Nowakowska M. and K. Szczubiałka, “Photosensitized dechlorination of polychlorinatedbenzenes. 1. Carbazole-photosensitizeddechlorinationofhexachlorobenzene,”

Chemosphere, vol. 39, no. 1, pp. 71–80, Jul. 1999.
[18] Khaled A., A. Rivaton, C. Richard, F. Jaber, and M. Sleiman, “Phototransformation of
Plastic Containing Brominated Flame Retardants: Enhanced Fragmentation and Releaseof Photoproducts to Water and Air,” Environ. Sci. Technol., vol. 52, no. 19, pp. 11123–
11131, Oct. 2018.
[19] Rånby B., J. F. Rabek, and G. Canbäck, “Investigation of PVC Degradation Mechanism
by ESR,” J. Macromol. Sci. Part - Chem., vol. 12, no. 4, pp. 587–608, May 1978.
[20] Kuzina S. I. and A. I. Mikhailov, “Chain and photochain mechanisms of photooxidationof polymers,” High Energy Chem., vol. 44, no. 1, pp. 37–51, Jan. 2010.
[21] Hurley C. R. and G. J. Leggett, “Quantitative Investigation of the Photodegradation ofPolyethylene Terephthalate Film by Friction Force Microscopy, Contact-Angle
Goniometry, and X-ray Photoelectron Spectroscopy,” ACS Appl. Mater. Interfaces, vol.
1, no. 8, pp. 1688–1697, Aug. 2009.
[22] RABEK J. F., Polymer Photodegradation: Mechanisms and experimental methods.
Springer Science & Business Media, 1994.
[23] Benítez A., J. J. Sánchez, M. L. Arnal, A. J. Müller, O. Rodríguez, and G. Morales,“Abiotic degradation of LDPE and LLDPE formulated with a pro-oxidant additive,”
Polym. Degrad. Stab., vol. 98, no. 2, pp. 490–501, Feb. 2013.

39

[24] Gardette M. et al., “Photo- and thermal-oxidation of polyethylene: Comparison of
mechanisms and influence of unsaturation content,” Polym. Degrad. Stab., vol. 98, no.
11, pp. 2383–2390, Nov. 2013.
[25] Martínez-Romo A., R. G. Mota, J. J. S. Bernal, C. F. Reyes, and I. R. Candelas, “Effectof ultraviolet radiation in the photo-oxidation of High Density Polyethylene andBiodegradable Polyethylene films,” J. Phys. Conf. Ser., vol. 582, p. 012026, Jan. 2015.
[26] Jung M. R. et al., “Validation of ATR FT-IR to identify polymers of plastic marine debris,
including those ingested by marine organisms,” Mar. Pollut. Bull., vol. 127, pp. 704–716,
Feb. 2018.
[27] Gewert B., M. M. Plassmann, and M. MacLeod, “Pathways for degradation of plastic
polymers floating in the marine environment,” Environ. Sci. Process. Impacts, vol. 17,
no. 9, pp. 1513–1521, 2015.
[28] Fotopoulou K. N. and H. K. Karapanagioti, “Degradation of Various Plastics in theEnvironment,” in Hazardous Chemicals Associated with Plastics in the Marine
Environment, vol. 78, H. Takada and H. K. Karapanagioti, Eds. Cham: Springer
International Publishing, 2017, pp. 71–92.

40

Chapter 3

Effect of polyethylene on photodegradation behaviors of a polycyclicaromatic hydrocarbon

3.1 IntroductionMore than five trillion pieces of plastics weighing equal to 25000 Mt are floating on the ocean
surface [1]. During floatation, plastics are broken down into smaller pieces, microplastics (< 5mm) and nanoplastics (< 0.1 mm) by light, waves, and currents [2], [3]. Microplastics arewidely spread into different environmental sections such as freshwater [4], open ocean [2], [5],coastal ecosystems [6], sediment [7], and arctic beaches [8], adsorbing HOCs because of astrong affinity due to hydrophobic interactions [9]–[11].
PAHs are a type of HOCs widely distributed in the atmosphere, water, soil, and sediment [12]–[15]. PAHs are easily absorbed at higher amounts on plastics due to - interactions [16], [17].
In addition, planar PAHs molecules can demonstrate higher sorption efficiencies than
nonplanar HOCs with the same hydrophobicity [18]. Several studies have reported theadsorption of PAHs on plastics in laboratory-scale experiments [9], [19]. In addition, PAHs
have been detected in the plastic debris samples collected from the aquatic environment [19]–[21].
In a previous study, the photodegradation of PAHs such as anthracene and benzo(a) pyrene insoil under sunlight was examined [22]; although these were photodegradable, more harmfulintermediates such as 1(2H)-acenaphthylenone and 9,10-phenanthrenedione were produced[22]. In addition, another study reported the production of 1-hydroxy-anthraquinone byphotodegradation of anthracene [23]. Chapter 2 reported the accelerated photodegradation of
PCP and TeCP on the PE sheet and suppressed photodegradation on the suppressed one on the
PET and PVC sheets. As the possible cause for this acceleration by PE, hydrogen donation was
41

expected. Owing to the previously reported oxidative photodegradation pathways of PAHs [22],[23], it was reasonable to expect that PE might affect the photodegradation of PAHs.
Although PAHs were detected in the microplastics [19]–[21], the effects of plastics on thephotodegradation of PAHs have not been addressed. Therefore, in this chapter, thephotodegradation of a PAH on PE was examined to investigate whether photodegradation
behaviors of the PAH were affected by PE.

3.2 Experimental
3.2.1 Materials
As a model PAH, NAP (> 97.0% purity) (Figure 3.1) and oxy-PAH, naphthacenequinone(NAPQ) (> 98% purity), were purchased from Tokyo Chemical Industry, Tokyo, Japan. PE 1.0
mm in thickness was chosen as a model plastic (Sanplatec, Osaka, Japan). Before use, PE sheets
were washed three times with distilled water and then dried at 50 °C. A mixed solvent of 90%(v/v) toluene (> 99.5% purity, Sigma-Aldrich, Mo, US) and 10% (v/v) cyclohexane (> 99.5%
purity, TCI) was used to extract NAP and photodegradation intermediates.

3.2.2 Photodegradation test of NAP and NAPQ
The PE sheet was cut to 2.2 mm × 7.0 mm and then put on the bottom of the quartz glass Petridishes (diameter 26 mm). Then, 2.0 mL of NAP solution (40.0 mg/L) was added dropwise on
the plastic sheets and allowed to stand in a dark place under laminar airflow for 2 hours tovaporize the mixed solvent (Figure 3.1). As mentioned in 3.3.2, NAPQ was produced byphotodegradation of NAP, photo degradability of NAPQ was investigated using the NAPQadsorbed PE prepared in the same manner of NAP photodegradation method. The NAPabsorbed and NAPQ-adsorbed PE sheets were respectively transferred into the quartz cuvettes
and capped with a rubber stopper and then irradiated under UV rays of 32 µW cm -2 for 2 hours

42

(λ=254 nm, DM-5, Daishinkogyo, Osaka, Japan). The cuvettes were kept at 15 C by a coolingplate (CP-120; Scinics Cool Plate, Tokyo, Japan) during UV irradiation time. Three cuvettes
were randomly taken for analysis of NAP and NAPQ at each sampling time. For extraction,
2.0 mL of the mixed solvent was added dropwise to the cuvettes. The resulting solution was
analyzed by GC-quadrupole mass spectrometry (GC/MS) (GC-8860 and MS-5977B, Agilent
Technologies, California, US) under the conditions described in Table 3.1. The mass spectrum
patterns generated by GC/MS were compared with NIST library data (NIST MS Search) to
search for photodegradation products. In the control experiment, a glass plate was used.
Experiments without UV irradiation were also carried out.

FTIR

Figure 3.1 Photodegradation of NAP on the PE sheet and measuring technique.

43

Table 3.1 Analytical condition of GC/MS
Gass chromatograph

GC 8860(Agilent Technologies, California, US)

Mass spectrometry detector

MSD 5977B(Agilent Technologies, California, US)

ColumnHP-5ms (30 m×0.32 mm, 0.25 µm, (Agilent
Technology, US)

Inject volume

1 µl (Split ratio15:1）

Inlet temp.

270℃

Pressure

97.3 kPa

Temperature profile

100℃（hold 3min）
15℃/min to 270℃ (hold 4 min)
20℃/min to 300℃ (hold 5min）

CarrierHe, 1.3 ml/min

Ionization

Electron Impact (70 eV)

Scan range

15.0 to 500.0 m/z

Hold time

5.0 min44

3.2.3 Analyses
To verify the consecutive degradation pathway of NAP via NAPQ production on PE, theconcentration of NAPQ during UV irradiation time was calculated from the degradation of
NAP assuming a sequential reaction of NAP to NAPQ.

k1

CNAP

k2

CNAPQ

Further photodegradation intermediates(1)

where k1 and k2 are the photodegradation rate constants for NAP and NAPQ [min-1],
respectively, and CNAP and CNAPQ are the NAP and NAPQ concentrations [mg/L],
respectively. As the ratio of k2 to k1 (0.0111/0.0144) does not equal 1, equation (4) can bederived from the equations (2) and (3):dCNAP/dt = - k1CNAP(2)

dCNAPQ/dt = k1CNAP-k2CNAPQ

(3)

∗

=

(𝑒

−𝑒

)+

∗
∗

(4)

𝑒

where CNAP* and CNAPQ* are the initial NAP and NAPQ concentrations [mg/L], respectively.

3.3 Results and Discussion
3.3.1 Photodegradation of NAP
The NAP concentration changes on the glass and PE sheets under UV irradiation are shown in
Figure 3.2 (a) and (b), where NAP on the glass and PE sheets decreased along with theirradiation time. However, without UV irradiation, the NAP concentration on the glass and PE
sheet did not significantly decrease. These results confirmed that NAP was photodegradable45

on the glass and PE sheet. Although NAP was degraded on glass, the photodegradation rate on
PE was higher than that of glass. The photodegradation of NAP on the glass and PE sheet
followed pseudo-first-order reaction kinetics (Figure 3.2 (c)), similar to the photodegradationof anthracene in soil [22]. The photodegradation rate constants on glass (kglass) and PE (kPE)
were 0.0101 and 0.0144 min-1, respectively (Figure 3.2 (c)); these values showed an
acceleration of NAP photodegradation by PE.
Accelerated photodegradation of PCP and TeCP was observed for PE, as described in Chapter
2. PE produces polymer alkyl radicals by removing hydrogen atoms by UV irradiation [24],[25]. Possible mechanisms on the accelerated photodegradation of NAP on the PE sheet mightbe (i) hydrogen donation by PE and/or (ii) production of alkyl radicals and hydroxyl radicals
produced by the reaction of the alkyl radical and oxygen (Figure 2.6).

46

NAP [mg/cm2]

0.025
Glass (UV)

Glass (w/o UV)(a)

PE (UV)PE (w/o UV)(b)kPE (w/o UV) = 0.0003
R² = 0.40

(c)

0.02
0.015
0.01
0.005
0
0.025

NAP [mg/cm2]

0.02
0.015
0.01
0.005
0
0.5

LnC/C0) [-]0
kGlass (w/o UV) = 7×10-05
R² = 0.47-0.5

kGlass = 0.0101
R² = 0.99-1

Glass (w/o UV)Glass (UV)
PE (w/o UV)
PE (UV)

-1.5
-2
0

30

kPE = 0.0144
R² = 0.98

6090
UV irradition time [min]120

Figure 3.2 Change of remaining NAP concentration on (a) the glass and (b) PE sheet
and (c) the result of photodegradation behavior analysis by the first-order reaction
kinetics. Bars indicate standard deviation (n=3).
47

3.3.2 Photodegradation intermediates produced from NAP
To investigate the photodegradation pathways of NAP, its photodegradation intermediates
were analyzed. Figure 3.3 shows the total ion chromatogram of NAP and its photodegradationintermediates obtained from the PE sheet after 120 min of irradiation. There were the two major
peaks at A and C. Based on the mass spectra of peak A, this was expected to be NAPQ (Figure3.4). As its authentic sample showed the same retention time and mass spectra, peak A was
identified as NAPQ. In addition, the mass spectra of peaks B and D were respectively identicalfor 1-phenyldibenzofuran and 2,5-diphenyl-1,4-benzoquinone. These results showed that NAP
was degraded via the production of the oxygenated NAP, with NAPQ as an intermediate. This
result agreed with the previous studies reporting the production of oxygenated PAHs
containing quinone structure by photodegradation of the precursor PAHs [26].
Although NAPQ was detected on the glass plate, its abundance was negligible (Figure 3.3).
This may be because NAPQ can easily evaporate from the glass plate, as described in 3.3.3. In
addition, it has been reported that oxygenated PAHss are more carcinogenic than the precursor
PAHs [27], [28]. The carcinogenicity of NAPQ is unknown; however, these results provided
evidence that plastics may hold photodegradation intermediates of PAHs, although some ofthem are highly toxic. As for the possible reason for the retainment of photodegradationintermediates, this might be their strong interaction with PE via π-π interaction, hydrophobic
interaction, and hydrogen bonding [11], [30]. The nascent hydrogen radical generated fromplastics might react with peroxy radical or alkyl peroxide radical, altering NAP degradation
pathways producing oxygenated intermediates.

48

200
Glass

Intensity [-]

D

PE

150

C

B

A

100

50

0
8

10

12
14
Retention time [min]16

18

Figure 3.3 Total ion chromatogram of the NAP and its photodegradation intermediates on
PE after 120 min of UV irradiation.

A (5,12-naphthacenedione)

Retention time = 17.00 min

49B (1-phenyldibenzofuran)

Retention time = 16.48 min

C (Naphthacene)

Retention time = 15.40 min

D (2,5-diphenyl-1,4-benzoquinone)

Retention time = 14.80 min

Figure 3.4 Mass spectra of naphthacene and the 3 photodegradation intermediates.
50

3.3.3 Photodegradation pathways of NAP
As NAPQ was detected on the glass plate and PE sheet, it was surmised that photodegradation
NAP produces NAPQ; then, NAPQ is further degraded (Reaction 1).

NAP → NAPQ → Further photodegradation intermediates such as
2,5-dipheny-l-1,4-benzoquinone (Reaction 1)

In this study, a difference in the photodegradation pathways of NAP was not observed between
the glass and PE sheet; however, it was expected that the photodegradation of NAPQ mightalso be accelerated by the PE sheet. Therefore, photodegradation of NAPQ was examined on
the glass plate and PE sheet.
Owing to the vaporization of NAPQ from the glass, the photodegradation rate constant (kGlass)may include NAPQ loss, and therefore, a comparison of photodegradation rate constants was
not possible (Figure 3.5). However, considering the vaporization rate constants (kGlass w/o UV),
the photodegradation rate constants of NAPQ on the glass plate and PE sheet seem to be thesame.
In the photodegradation experiment of NAPQ on the PE sheet, the intermediate compounds 1phenyldibenzofuran,and2,5-dipheny-l-1,4-benzoquinone

were

detectedonthephotodegradation of NAP were not observed. However, the concentration of NAPQ was
decreased compared to samples without UV irradiation.

51

NAPQ [mg/cm2]

0.025
0.02

Glass

Glass (w/o UV)PE

PE (w/o UV)

(a)

0.015
0.01
0.005
0

NAPQ [mg/cm2]

0.025
0.02

(b)

0.015
0.01
0.005
0

kPE (w/o UV) = 0.0017

0R² = 0.94 (c)kGlass (w/o UV) = 0.0035
R² = 0.95
kPE(UV) = 0.0111
R² = 0.98

Ln (C/C0) [-]

-0.5

-1

-1.5
Glass (w/o UV)Glass (UV)
PE (w/o UV)
PE (UV)-2

kGlass = 0.0126
R² = 0.96

-2.5
0

30
60
90
120
UV irradiation time [min]Figure 3.5 Change of remaining NAPQ concentration on (a) the glass and (b) PE sheet and(c) the result of photodegradation behavior analysis by the first-order reaction kinetics. Bars
indicate standard deviation (n=3).
52

Next, to verify the hypothesized photodegradation pathway of NAP (Reaction 1), a sequential
reaction analysis was carried out to estimate a NAPQ concentration using the photodegradation
rate constants for NAP and NAPQ. Because the detection level of NAPQ was negligible on theglass plate, here the degradation behaviors on the PE sheet were analyzed. As shown in Figure
3.6, the estimated NAPQ concentration was higher than that of the actual NAPQ concentration,
thereby suggesting the existence of other photodegradation pathways of NAP. As described in
Figure 2.6, PE can provide alkyl, hydrogen, and hydroxyl radicals, and this might result inother degradation pathways from NAP, though evidential degradation intermediates are still inthe quest.

0.0175
NAP (UV on PE)
NAPQ (Produced from NAP degradation on PE)

0.015
Comncentration (mg/cm2)

NAP (Expected)
NAPQ (Expected)

0.01250.01
0.00750.005
0.00250060
UV irradiation time [min]120

Figure 3.6 Simulation of NAP photodegradation pathway and production NAPQ on PE
following first-order reaction. Bars indicate standard deviation (n=3).

53

3.4 Conclusions
The photodegradation of NAP on the glass plate and PE sheet was confirmed. The
photodegradation rate constant of NAP on the PE sheet was higher than that on glass, likely
due to the radicals produced by the photolytic degradation of the PE structure. The
photodegradation of NAP produced the oxygenated NAP and NAPQ. Surprisingly NAPQ was
remarkably abundant on the PE sheet but not the glass plate. As oxygenated PAHs may be
more toxic than the precursor PAHs, these results confirm that attention should be paid toretaining organic pollutants and their photodegradation intermediates on plastics for risk
management. Photodegradation pathways of NAP via production of NAPQ were checked by
the sequential reaction analysis; however, unknown degradation pathways adding to the
production of NAPQ might exist.

543.5 Literature cited[1] Eriksen M. et al., “Plastic Pollution in the World’s Oceans: More than 5 Trillion Plastic
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Dec. 2014.
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Lin W. et al., “Sorption properties of hydrophobic organic chemicals to micro-sized

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pollution increases gene exchange in aquatic ecosystems,” Environ. Pollut., vol. 237, pp.
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from two beaches of the Portuguese coast,” Mar. Pollut. Bull., vol. 60, no. 11, pp. 1988–
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and PCBs in the sediment pore water of the German Bight and Wadden Sea,” Mar. Pollut.
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58

Chapter 4

Effects of the aging of plastics on the photodegradation of a chlorophenol

4.1 Introduction
The global estimated mass of the floating plastic particles indicates that the amount of plastic
waste entering the ocean from land ranged from 4.8 to 12.7 Mt in 2010 [1]. Floating plastics
are continuously fragmented into smaller plastic pellets, such as microplastics, through aging
processes such as physical abrasion [2], [3], ocean turbulence [4], UV photo-oxidation [3], andbiodegradation [5]. Thus, the proportion of such fragments in total marine debris is growing[3], [6]. Fragmented plastics adsorb hydrophobic organic pollutants such as PAH [7], PCBs [8],and DDT [9] in seawater with a 105–106 concentration factor [10]. Microplastics can enter the
food chain through ingestion by aquatic organisms such as deposit and suspension feeders [11],acting as organic pollutant carriers. Therefore, the interaction of organic pollutants andmicroplastics must be studied.
Microplastics floating in seawater are exposed to sunlight, and the photodegradation ofadsorbed pollutants may occur. In addition, microbial degradation may occur; however,hydrophobic organic pollutants, such as PAH and PCBs, are generally recalcitrant to microbial
degradation. Furthermore, chemical degradation caused by natural oxidants, such as hydroxyl
radicals, may occur; however, photodegradation is generally dominant in the environment [12].
Therefore, this research focused on photodegradation. In chapter 2, plastics effect on thephotodegradation of tetra- and PCP adsorbed on plastic sheets showed that the degradation wasaffected by the texture of the plastic sheets. The photodegradation of CPs was accelerated on
PE sheets but suppressed on PVC and PET sheets. Hydrogen donation is a possible cause of
the acceleration of the photodegradation on the PE sheet. Another study investigated thephotodegradation of polystyrene films containing brominated flame retardants, such as
59

decabromodiphenyl ether and tetrabromobisphenol A, observing 14 brominated products,
including tetrabromoacetophenone and bromobenzoic acid, via bromination and oxidation ofpolystyrene [13].
Aging processes can change the chemical and physical characteristics of plastics [14], causing
plastic fragmentation in marine environments. Aging increases the specific surface area andoxygen content of plastics [15], changing the color and transparency of the plastic. A previous
study indicated that UV radiation increased the number of carbonyl groups in PP andstrengthened its yellowness index [15], [16]. In addition, plastic additives such as di-isobutyl
phthalate, dimethyl phthalate, and diethyl phthalate can leach into aquatic environments [18].
These changes suggest that aging affects the photodegradation of organic pollutants adsorbed
on plastics. Ultimately, all plastics are aged in marine environments. Despite the importance ofunderstanding the effects of the aging of plastics on the photodegradation of organic pollutants
adsorbed on plastics, previous studies [13],[14] have not addressed this issue.
In this study, the photodegradation of an organochlorine adsorbed on virgin and aged plastic
sheets was investigated to determine the effects of plastic aging on organochlorine
photodegradation. Chemical and physical properties of plastics, such as functional groups,
surface roughness, and transparency caused by aging, were investigated to establish the aging
effects of plastics. Owing to the possible elution of the plastic additives in the marine
environment, the effects of plastic additives were also investigated using methanol-extracted
plastics.

60

4.2 Experimental
4.2.1 Materials
For the materials, 1.0-mm-thick Virgin PE, PVC, and PET sheets were purchased from the
Sanplatec Corp., Japan. This thickness was chosen as the thickness of microplastics is generally
lower than 5 mm. Each sheet was washed thrice with distilled water before use and dried at
50°C. TRiCP (Tokyo Chemical Industries, Tokyo, Japan) was used as the modelorganochlorine pollutant.

4.2.2 Plastic sample preparation and additive extraction
Plastics drift in oceans and are distributed at shorelines and beaches by tides [18]–[22]. Aging
treatments need to reproduce both dry and wet conditions; however, our aging test chamber
could not reproduce wet conditions. Therefore, artificially aged plastics were prepared under
dry conditions in this study. Aged plastic samples were prepared by irradiating PE, PVC, andPET sheets at an intensity of 75 mW cm-2 in a Q-SUN Xenon test chamber (Q-SUN1000, Qpanel, China) at 40C and 50% humidity for 24 h. The effects of additives in the PE, PVC, andPET sheets were examined using additives extracted from the plastic samples. Briefly, eachplastic sheet (approximately 5 g) was immersed in 50 mL of methanol at 25 C for 72 h. Next,
2-mL samples of the methanol extracts containing plastic additives equivalent to 0.20 g per
plastic sheet were decanted onto the 0.50 g virgin plastic sheets. Additive-spiked plastic
samples were obtained by evaporating the methanol at 25 °C for 60 min in a laminar airflow
chamber. Additionally, the plastic samples from which additives had been removed bymethanol extraction were tested after methanol evaporation.

61

4.2.3 Photodegradation
All plastic sheet samples were cut into circles (diameter, 26 mm). The circular plastic sheets
were placed at the bottom of quartz glass Petri dishes (diameter, 26 mm), and TriCP was thenadsorbed onto the plastic sheets by decanting 2-mL samples of an n-hexane solution containing
100 mg L-1 TriCP onto the plastic sheets. After n-hexane evaporation in a laminar airflow
chamber at 25°C, the Petri dishes were covered with quartz glass lids and placed in a Navis
Schreiner photochemical reactor (Model DM-05, Daishin Kogyo Co., Ltd., Japan) (Figure 4.1)
equipped with a 10 W low-pressure UV lamp (Model: NEC GL-10; UV-C) emitting 254 nm
UV light at 32 µW/cm2. Although the tested TriCP concentration and the light intensity utilized
may be higher than that in the actual environment, we carried out photodegradation
experiments under this condition because our objective was to determine the effects of plastic
aging on the organochlorine photodegradation rate rather than to precisely measure thephotodegradation rate.
UV light source

Evaporation of solvent

Cover with a quartz glass lid

UV irradiation

Figure 4.1 Experimental setup for photodegradation of TriCP.

4.2.4 Characterization of plastic sheets
Surface texture changes of the plastic sheets were investigated. The plastic sheet samples wereplaced in a carbon coater in a vacuum and coated with a carbon layer (thickness, approximately
2 nm) (Meiwa Forsis CADE, MeiwaForsis Co., Ltd., Tokyo, Japan). The samples were62

examined using field-emission scanning electron microscopy after carbon coating (FE-SEM;
High-Tech S-5200, Hitachi, Tokyo, Japan). Images were acquired at 3.0 kV to avoid charging
effects during image acquisition. FE-SEM images of each sample were used to estimate thechanges in each plastic surface. The images were analyzed by the local surface roughness
method [23] using the ImageJ 1.53a SurfCharJ plugin (http://imagej.nih.gov/ij). The arithmetic
mean deviation (Ra) and skewness (Rsk) were estimated using the following equations [24]:

Ra =

∑

Rsk =

∑

𝑍

∑

∑

(1)
(2)

𝑍
/

Rq =

∑

∑

(3)

𝑍

where Nx is the arithmetic mean deviation for the x-axis, Ny is the arithmetic mean deviation
for the y-axis, Zij represents the assessed surface, and Rq represents the kurtosis.

Although the TriCP distribution in the plastic sheets was not studied, it is plausible that a
proportion of the added TriCP could enter the sheets by penetrating the TriCP-containing nhexane. The transparency of the plastic sheet was therefore examined by measuring theintensity of the transmitted light with a UV meter (CUSTOM, UV-340c, Taiwan). The UV
meter was placed 10 cm below a UV light source. The plastic sheet samples were placed on
the UV sensor, and UV attenuation was determined. An FT-IR spectrometer (IRPrestige-21,
Shimadzu, Japan) equipped with a tungsten iodide light source was used to investigate thechanges in functional groups on the surfaces of the plastic sheet samples. The FT-IR spectra
were recorded from 400 to 4000 cm−1 at a nominal resolution of 4 cm−1 for a total of 32 scans.

634.2.5 Analyses
Changes in the TriCP concentration caused by UV irradiation were determined as follows.
Extraction was performed by pouring methanol into a Petri dish containing a plastic sample.
The methanol extract was mixed with 0.1% formic acid (pH 2.70) solution at a volume ratio of
4:1. Next, 100 μL of the sample mixture was subjected to chromatography on an octadecyl
silica column (Cosmosil 5C18-AR-II, 4.6 × 150 mm, Nacalai Tesque Inc., Japan) at a flow rateof 1.0 mL min-1, with a mixture of methanol and formic acid (4:1 volume ratio) as the mobile
phase. A UV-visible detector with a wavelength of 254 nm (SPD-10A, Shimadzu, Japan) wasused for detection.

4.2.6 Statistical analysis
ANOVA and subsequent Tukey tests were performed to determine the statistical significanceof the photodegradation rate constant. Normal probability distribution and homogeneity ofvariance were assumed.

4.3 Results and Discussion
4.3.1 Comparison of photodegradation rates of TriCP on virgin and aged plastics
The TriCP photodegradation on virgin and aged plastic sheets followed pseudo-first-order
reaction kinetics is shown in Figures 4.2(a) and (b). The rate constant for TriCP
photodegradation on glass was 1.6 × 10−2 min−1. The recorded rates of degradation on virginplastic sheets were 2.8 × 10−2 min−1, 8.9 × 10−3 min−1, and 1.2 × 10−2 min−1 on the PE, PET,and PVC sheets, respectively (Table 4.1). The photodegradation constants indicate thatphotodegradation was accelerated on the virgin PE sheet but suppressed on the PET and PVC
sheets. This is consistent with the results of a previous study on the photodegradation of tetra
and PCP on PE, PET, and PVC sheets; the acceleration on PE might be attributed to hydrogen640kControl (Glass) = 0
R² = 1.0
kPET = 0.0089
R² = 0.92-1

Ln (C/C0) [-]

-2
-3
-4

kPVC = 0.0124
R² = 0.98
kGL = 0.0163
R² = 0.99

Glass
Virgin PE
Virgin PVC
Virgin PET
Control (Glass)-5-6

kPE = 0.0282
R² = 0.99-70kControl (Glass) = 0
R² =1.0kPET = 0.0034
R² = 0.91
kPVC = 0.0075
R² = 0.99
kPE = 0.0127
R² = 0.98-1

Ln (C/C0) [-]

-2
-3
-4

kGL = 0.0163
R² = 0.99

Glass
Aged PE
Aged PVC
Aged PET
Control (Glass)

-5
-6
-7
0

60

120
180
UV irradiation time [min]

240

300

Figure 4.2 Photodegradation of TriCP on PE, PET, and PVC sheets: (a) virgin and (b) aged
plastics. Bars indicate standard deviations (n = 3) except the aged PVC and PET (n=2). Bars
for the aged PVC and PET indicate the differences between the averaged and measured values.

donation (Figure 2.6). The reason for the large error bar for the virgin PE after 120 min is
unknown; however, it can be attributed to independent samples. The photodegradation rate65

constants on aged PE, PET, and PVC sheets were 1.2 × 10−2 min−1, 3.4 × 10−3 min−1, and 7.5
× 10−3 min−1, respectively; these rates are lower than those of the corresponding virgin plastic
samples. Although photodegradation of TriCP was accelerated on virgin PE (p = 1.9 × 10−6),
aging significantly suppressed photodegradation on PE (p = 1.9 × 10−6). These results confirm
that organochlorine photodegradation is suppressed on aged plastics.
Such degradation suppression can be attributed to the surface textural changes, causing UVshading and interception due to changes in the functional groups in the plastic sheets. Therefore,
the changes in surface roughness, skewness, transparency, and functional groups in the aging
plastics were investigated.

Table 4.1 TriCP Photodegradation rate constants [min−1] on plastic surfaces
Plastic name

VirginAged

Methanol-treated

Extracted
additive-spikedPE

2.8×10-2

1.2×10-2

2.2×10-2

2.3×10-2

PVC

1.2×10-2

7.5×10-3

7.2×10-3

2.6×10-3

PET

8.9×10-3

3.4×10-3

8.7×10-3

5.2×10-3

4.3.2 Changes in surface textures
Figure 4.3 indicates that aging changed the surface textures of all the plastic samples. Aging
treatment resulted in valley formation (Rsk). Valleys can shade light; therefore, surface changes
could contribute to the decreased photodegradation rates on the aged plastics (Figure 4.2).
Aging increases the plastic’s surface roughness (Ra) by causing cracks, crevices, andmorphological changes [25]. Analysis of the changes in surface roughness showed that the Ravalues decreased (p < 3.3 × 10−5) (Figure 4.4(a)). This might be because of differences between
our artificial and natural aging processes. In aquatic environments, plastic additives can leach
during aging [17]. The plastic samples did not come into contact with water in artificial aging
66

with a Q-SUN Xenon test chamber, indicating that the treatment might simulate aging on lands
under fine weather rather than that in aquatic environments. However, the aging treatment led
to increases in the Rsk values at the surfaces of all the plastic sheets (p < 0.047) (Figure 4.4(b)),
evidential for the observed valleys (Figure 4.3).

(f)

(e)

Figure 4.3 Three-dimensional SEM images of (a) virgin PE, (b) aged PE, (c) virgin PVC, (d)aged PVC, (e) virgin PET, and (f) aged PET sheet surfaces.

67

20
PE

PVC

(a)

PET

Ra (μm)

16
12
8
4
0
VirginAged
Plastic types

Methanol treated

1.2
1

PE

PVC

(b)

PET

Rsk (-)

0.8
0.6
0.4
0.2
0
VirginAged
Plastic types

Methanol treatedFigure 4.4 Changes in the plastics (a) surface roughness (Ra) and (b) skewness (Rsk) caused byaging and methanol extraction. Bars indicate standard deviation (n = 3).

The methanol extraction process did not alter the Rsk value of PE but increased that of PET and
PVC (p < 0.049). The extraction treatment increased the Ra value of PE (p = 3.0 × 10−5) but notthat of PET and PVC. A comparison of the TriCP photodegradation rate constants on the virgin
68and methanol-extracted plastic samples indicated that the latter samples produced lower
photodegradation rate constants (Table 4.1). We do not have any evidence as to whether Ra or
Rsk significantly reduce photodegradation rates. However, the photodegradation rate constants
decreased when the Ra or Rsk values increased due to aging or methanol extraction (Figure
4.4(a) and (b)).
A comparison of the TriCP photodegradation rate constants on the methanol-extracted plastic
samples and aged plastic samples indicates that the rate constants of the methanol-extracted
plastic samples were higher than those of the aged plastic samples (Table 4.1). However, themethanol extraction process significantly increased the Ra value for PE (p = 3.0 × 10−5) and Rsk
values for PET and PVC (p < 0.049) (Figure 4.4(a) and (b)). This inconsistency can be
attributed to the fact that certain plastic additives that suppress the TriCP photodegradation canbe removed by methanol.

4.3.3 Suppression of photodegradation by the methanol extract
The TriCP photodegradation was suppressed by spiking the plastic samples with the methanol
extract (Table 4.1). In the manufacturing of plastics, anti-UV agents, free-radical scavengers,andhydroperoxide

decomposers,

such

as

hydroxybenzophenones,

hydroxyphenylbenzotriazoles, and derivatives of phenyl salicylates are added as radical
scavengers to hinder plastic photodegradation [14], [26]. These scavengers penetrate thematerial, and energy is rapidly dissipated because of the fall in UV intensity; therefore,
degradation is concentrated near the surface. Although we did not analyze the methanol
extracts of each plastic sample, it is reasonable to expect that commercially obtained virginplastic sheets contain such agents.

69

4.3.4 Changes in transparency
Aging changes color [14] and opacity of plastic surfaces [16]. All plastic sheets possessed
yellowish discoloration in this study. As the transparency of the PVC sheet to UV could not bemeasured using our method, only the virgin and aged PE and PET sheets were tested. We
established that the aging treatment reduced the transparency of the PE and PET sheets byapproximately 4 μW m-2; however, methanol extraction did not affect the transparency (Figure4.5). These results indicate that a decrease in transparency can suppress the TriCP
photodegradation rate in aged plastics. The photodegradation rate constants on the PE and PET
sheets were reduced by 46% and 56%, respectively (Table 4.1) by the aging treatment.
Although the reduced transparencies of the PE and PET sheets remained the same atapproximately 4 μW m-2, the reduction percentages were different (70% for PE and 0% forPET) (Figure 4.5). This suggests that the transparency reduction did not primarily suppress
photodegradation.

18
VirginTransmited light (μW/cm2)

16

Methanol treated

14

Aging

12
10
8
6
4
2
0
PE

PET

Figure 4.5 Changes in UV transparency of PE and PET sheets caused by aging. Bars indicate
standard deviation (n = 3).

70

4.3.5 FT-IR analysis of aged plastic sheets
Polymer degradation is initiated by C–H bond breakage. During propagation, carbonyl groups
are formed during the aging of plastics through Norrish Type I and II reactions [14]. This leads
to plastic fragmentation [3] via termination [14], [27]. UV irradiation increases the plastic
fragmentation rate because of the production of intramolecular polymers, hydroperoxy, andhydroxyl radicals [14], [26], slowing hydrogen abstraction [28]. The FT-IR spectra of themodel plastic samples indicated that molecular changes were caused by plastic aging (Figure
4.6(a)–(c)).
In the spectrum of the aged PE sheet, peaks corresponding to alkene stretching vibrations
appeared at 990 cm−1 and 890 cm−1 (Figure 4.6(a)). These peaks provide evidence of hydrogen
abstraction caused by aging. As dechlorination occurs during the photodegradation of CPs on
PE [29], the suppression of TriCP photodegradation on an aged PE surface could be caused bydecreased hydrogen donation. In addition, a new peak was observed at 1120 cm−1,
corresponding to –C=O aldehyde stretching or carboxylic acid ester bending. Magnification
peaked at 2660 cm−1, corresponding to –C–H aldehyde stretching or carboxylic acid ester
bending. These peaks normally appear in the spectra of aged plastics [30], possibly increasing
and decreasing the surface skewness and transparency, respectively.
The PVC aging weakened the C–H bending peaks at 638 cm −1 and 705 cm−1 in the PVC
spectrum (Figure 4.6(b)). In addition, the signals at 1400 cm−1 from C–H bending, 1580 cm−1
from cyclic alkenes, and 1716 cm−1 from C–O bending were weakened. This suggests thataging causes unsaturation of PVC. Dehydrochlorination and C=C formation in PVC have beenpreviously reported [31]; our desaturated data agree with these findings. However, the C=C
peak did not appear in the aged PVC spectrum. This might be attributed to crosslinking [31]rather than alkene formation. The FT-IR spectra provide evidence of molecular changes caused

71

(a)

(b)

(c)

Figure 4.6 FT-IR spectra of the virgin and aged (a) PE, (b) PVC, and (c) PET sheets.
72

by PVC aging. These changes might contribute to surface textural changes, such as increased
skewness.
PET photodegradation is caused by chain scission, leading to a decrease in the molecularweight of the polymer and generating carboxyl end groups [32]. We found that the aging of
PET weakened the alkene bending peaks at 3030 cm−1 and 1942 cm−1, and the C–H bending
peak at 745 cm−1 (Figure 4.6(c)). These changes indicate that the aging of virgin PET changed
the surface structure of PET, altering PET skewness and transparency.
The plastic samples did not come into contact with water in the aging treatment; however, it isplausible that aging processes might be affected by water in aquatic environments. Future
research needs to analyze plastic samples aged in contact with water, to understand the changes
in the chemical structures of plastics due to aging in the aquatic environment.

4.4 Conclusions
TriCP photodegradation on plastic surfaces was suppressed by plastic aging. This could becaused by changes in the plastic surface due to aging, such as valley formation in the surface
texture. In addition, the decrease in plastic transparency caused by plastic aging can also
suppress photodegradation. Furthermore, the TriCP photodegradation was suppressed by theaddition of methanol extracts of the plastics. These results indicate that the aging of plastics
can affect the photodegradation of contaminants through changes in the surface texture andbehavior of plastic additives.

734.5 Literature cited[1] Wilcox C., B. D. Hardesty, and K. L. Law, “Abundance of Floating Plastic Particles Is
Increasing in the Western North Atlantic Ocean,” Environ. Sci. Technol., vol. 54, no. 2,
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[2] Koelmans A. A., E. Besseling, and E. M. Foekema, “Leaching of plastic additives tomarine organisms,” Environ. Pollut., vol. 187, pp. 49–54, Apr. 2014.
[3] Song Y. K., S. H. Hong, M. Jang, G. M. Han, S. W. Jung, and W. J. Shim, “Combined
Effects of UV Exposure Duration and Mechanical Abrasion on Microplastic
Fragmentation by Polymer Type,” Environ. Sci. Technol., vol. 51, no. 8, pp. 4368–4376,
Apr. 2017.
[4] Suhrhoff T. J. and B. M. Scholz-Böttcher, “Qualitative impact of salinity, UV radiationand turbulence on leaching of organic plastic additives from four common plastics — A
lab experiment,” Mar. Pollut. Bull., vol. 102, no. 1, pp. 84–94, Jan. 2016.
[5] Yang Y., J. Yang, and L. Jiang, “Comment on ‘A bacterium that degrades and assimilatespoly(ethylene terephthalate),’” Science, vol. 353, no. 6301, pp. 759–759, Aug. 2016.
[6] Andrady A. L., “Microplastics in the marine environment,” Mar. Pollut. Bull., vol. 62, no.
8, pp. 1596–1605, Aug. 2011.
[7] Fries E. and C. Zarfl, “Sorption of polycyclic aromatic hydrocarbons (PAHs) to low andhigh density polyethylene (PE),” Environ. Sci. Pollut. Res., vol. 19, no. 4, pp. 1296–1304,
May 2012.
[8] Devriese L. I., B. De Witte, A. D. Vethaak, K. Hostens, and H. A. Leslie,“Bioaccumulation of PCBs from microplastics in Norway lobster (Nephrops norvegicus):
An experimental study,” Chemosphere, vol. 186, pp. 10–16, Nov. 2017.

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[9] Bakir A., I. A. O’Connor, S. J. Rowland, A. J. Hendriks, and R. C. Thompson, “Relative
importance of microplastics as a pathway for the transfer of hydrophobic organic
chemicals to marine life,” Environ. Pollut., vol. 219, pp. 56–65, Dec. 2016.
[10] Mato Y., T. Isobe, H. Takada, H. Kanehiro, C. Ohtake, and T. Kaminuma, “Plastic Resin
Pellets as a Transport Medium for Toxic Chemicals in the Marine Environment,” Environ.
Sci. Technol., vol. 35, no. 2, pp. 318–324, Jan. 2001.
[11] Thompson R. C. et al., “Lost at Sea: Where Is All the Plastic?,” Science, vol. 304, no.
5672, pp. 838–838, May 2004.
[12] Czaplicka M., “Photo-degradation of chlorophenols in the aqueous solution,” J. Hazard.
Mater., vol. 134, no. 1–3, pp. 45–59, Jun. 2006.
[13] Khaled A., A. Rivaton, C. Richard, F. Jaber, and M. Sleiman, “Phototransformation of
Plastic Containing Brominated Flame Retardants: Enhanced Fragmentation and Releaseof Photoproducts to Water and Air,” Environ. Sci. Technol., vol. 52, no. 19, pp. 11123–
11131, Oct. 2018.
[14] Gewert B., M. M. Plassmann, and M. MacLeod, “Pathways for degradation of plastic
polymers floating in the marine environment,” Environ. Sci. Process. Impacts, vol. 17, no.
9, pp. 1513–1521, 2015.
[15] Rouillon C. et al., “Is carbonyl index a quantitative probe to monitor polypropylene
photodegradation?,” Polym. Degrad. Stab., vol. 128, pp. 200–208, Jun. 2016.
[16] Baltscheit J., N. Schmidt, F. Schröder, and J. Meyer, “Investigations on the aging
behavior of transparent bioplastics for optical applications,” InfoMat, vol. 2, no. 2, pp.
424–433, Mar. 2020.

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[17] Paluselli A., V. Fauvelle, F. Galgani, and R. Sempéré, “Phthalate Release from Plastic
Fragments and Degradation in Seawater,” Environ. Sci. Technol., vol. 53, no. 1, pp. 166–
175, Jan. 2019.
[18] Kooi M., E. H. van Nes, M. Scheffer, and A. A. Koelmans, “Ups and Downs in the Ocean:
Effects of Biofouling on Vertical Transport of Microplastics,” Environ. Sci. Technol., vol.
51, no. 14, pp. 7963–7971, Jul. 2017.
[19] Song Y. K. et al., “Horizontal and Vertical Distribution of Microplastics in Korean
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[20] Khan F. R., K. Syberg, Y. Shashoua, and N. R. Bury, “Influence of polyethylene
microplastic beads on the uptake and localization of silver in zebrafish (Danio rerio),”
Environ. Pollut., vol. 206, pp. 73–79, Nov. 2015.
[21] Camacho M. et al., “Organic pollutants in marine plastic debris from Canary Islands
beaches,” Sci. Total Environ., vol. 662, pp. 22–31, Apr. 2019.
[22] Bergmann M., B. Lutz, M. B. Tekman, and L. Gutow, “Citizen scientists reveal: Marine
litter pollutes Arctic beaches and affects wild life,” Mar. Pollut. Bull., vol. 125, no. 1–2,
pp. 535–540, Dec. 2017.
[23] Vasanth Kumar R. and B.M. Preetham “Study of Machining on Surface Roughness ofEngineering Plastics,” Int. J. Eng. Res., vol. 3, no. 5, p. 5, 2014.
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2007.

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[25] Liu P., X. Zhan, X. Wu, J. Li, H. Wang, and S. Gao, “Effect of weathering on
environmental behavior of microplastics: Properties, sorption and potential risks,”
Chemosphere, vol. 242, p. 125193, Mar. 2020.
[27] White J. R. and A. Turnbull, “Weathering of polymers: mechanisms of degradation andstabilization, testing strategies and modelling,” J. Mater. Sci., vol. 29, no. 3, pp. 584–613,
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[30] Fernández-González V., J. M. Andrade-Garda, P. López-Mahía, and S. MuniateguiLorenzo, “Impact of weathering on the chemical identification of microplastics from usual
packaging polymers in the marine environment,” Anal. Chim. Acta, vol. 1142, pp. 179–
188, Jan. 2021.
[31] Rånby B., J. F. Rabek, and G. Canbäck, “Investigation of PVC Degradation Mechanism
by ESR,” J. Macromol. Sci. Part - Chem., vol. 12, no. 4, pp. 587–608, May 1978.
[32] Hurley C. R. and G. J. Leggett, “Quantitative Investigation of the Photodegradation ofPolyethylene Terephthalate Film by Friction Force Microscopy, Contact-Angle
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1, no. 8, pp. 1688–1697, Aug. 2009.

77

Chapter 5

Migration of adsorbed organochlorines from plastics to sediment

5.1 Introduction
Microplastics are now omnipresent in the environment and have been detected in significant
amounts in different environments such as freshwater [1], soil [2], beaches, and sediment [1],[5]–[7]. It was reported that microplastics adsorb HOCs at a concentration factor of 10 5-6 [1]greater than the surrounding ocean surface. Thus, it is apparent that microplastics can act as acarrier of HOCs for translocation, magnification, and accumulation into different
environmental sections, including living biota [2]–[4].
The density of buoyant microplastic debris may increase during their residence in the ocean
environment by biofouling. Microplastics start sinking upon reaching seawater density [5].
Microplastics have been detected in the different sediment zones of the seas, including deepsea shorelines [6], [7]. As such, sediment is the final destination of microplastics in the marine
environment [4], [8]. Microplastics cause negative impacts on the survival, growth of benthic
invertebrates [9]. As microplastics are apparent on the benthic zone sediment [4], [7] and apotential carrier of HOCs [1], [10], [11], the migration behaviors of HOCs-adsorbed onto
plastics to sediment needs to be evolved, but no study has been focused on this issue.
In this study, the migration of TriCP and 1-chlorooctadecane (ClOD) from PE, PVC, and PET
to sediment was examined in sediment under control conditions.

78

5.2 Experimental
5.2.1 Materials
Virgin sheets of PE, PVC, and PET with a thickness of 1.0 mm were purchased from Sanplatec
Corp. (Tokyo, Japan). Before use, each pristine sheet was washed with distilled water threetimes and then dried in the oven at 50 °C to remove potential pollutant interference. TriCP and
ClOD were obtained from Tokyo Chemical Industries (Tokyo, Japan). Methanol (99% purity),
phosphoric acid (H3PO4), potassium hydroxide (KOH) (99% purity), acetone (98% purity), andhexane (98% purity) were purchased from Nacalai Tesque Inc. (Japan).

5.2.2 Sediment preparation
Sediment was collected at the mouth of Enkougawa River, Hiroshima, Japan, in October 2019,at a habitat of many sediment-dwelling invertebrates. Sediment samples were wet sieved
through a 60-mesh stainless steel sieve to remove larger particles and debris, and then thesieved sediment was weighed and storaged in glass jars. To obtain microplastic-free sediment,
the sediment was washed with milli-Q water with passing air for floating light debris adhered
to the sediment [12]. After removing the excess water by decantation, the sediment sample wasobtained. The characteristics of the sediment samples are summarized in Table 5.1. To avoid
microbial interference and biological degradation, the sediment sample was autoclaved at
120C for 30 min before its use.

79

Table 5.1 Characteristics of the sediment sample.
Water content

31%

Ignition loss (IC)

5.2%

pH

7.9

Dissolve organic carbon (DOC) ofinterstitial water
Salinity of interstitial water

50.54 mg/L
0.60 psu

5.2.3 Migration of adsorbed organochlorine from plastic to sediment
Approximately 5.0 g of wet sediment was kept in a round cap glass bottle and autoclaved.
Before use, the glass bottles were washed with methanol and distilled water, respectively. For
preparing the organochlorine-adsorbed plastic samples, 2.0 ml of 100 mg/L containing TriCP
and 2 mL of 100 mg/L containing ClOD were respectively put dropwise on 5.30 cm 2 of PE,
PET, and PVC sheets, and the solvent was removed by evaporation for two hours. Then, about
0.5 g of each plastic sheet onto which TriCP and ClOD were respectively adsorbed at amounts
7.53 and 1.88 mg/cm2 were buried in the sterile sediment sample separately at around 2–3 mm
depth. The glass bottles were sealed with a rubber cap and incubated in a dark room at 25 C
for 72 hours. To measure the migrated amount of TriCP and ClOD to the sediment, the plastic
samples were picked up from the sediment every 24 hours, and TriCP and ClOD wererecovered from plastics by methanol extraction. Migration of TriCP and ClOD into the artificial
interstitial water was estimated using equation (1).

5.2.4 Mechanism of the migration process
HOC migration may occur (i) via dissolution followed by adsorption onto sediment and (ii)direct translocation by contact with sediment particles. To verify these hypotheses, HOC
dissolution into the interstitial water was checked. Briefly, formic (HCOOH) acid was added80

to Milli-Q water to attain the same DOC (50 mg/L) as that of the interstitial water and adjusted
to pH 8.0 using H3PO4 and 0.1 M KOH solutions according to the ratio as determined by theHenderson-Hasselbalch equation. Then the organochlorine adsorbed plastic sheets were put
into this artificial interstitial water and maintained at 25 °C in the darkroom. To estimate thedissolved amount of TriCP and ClOD, the remaining TriCP and ClOD were extracted from theplastics using methanol and acetone respectively. The migration of TriCP and ClOD into theartificial interstitial water was estimated using equation (1).

5.2.5 Analysis of TriCP and ClOD
The methanol extracted TriCP was analyzed using an HPLC-UV detector; a 20-mM phosphatebuffer (pH 2.50) solution was mixed at a volume ratio of 4:1 and used as mobile phase. A
mixture sample (100 μL) was then analyzed by chromatography on an octadecyl silica column(Cosmosil 5C18-AR-II, 4.6 mm × 150 mm, Nacalai Tesque Inc., Japan) at a flow rate of 1.0
mL min−1. The TriCP was detected by a UV–visible detector at 254 nm (SPD-10A, Shimadzu,
Japan).
The acetone solution was analyzed for ClOD by GC/MS (GC-8860 and MS-5977B, Agilent
Technologies, CA, US) under the conditions described in Table 5.2.

81

Table 5.2 Operating condition of GC/MS for analysis of ClOD.
Gass chromatograph

Mass spectrometry detector

GC 8860(Agilent Technologies, California)
MSD 5977B(Agilent Technologies, California)

ColumnHP-5ms (30 m×0.25 mm, 0.25 µm)

Inject volume

1 µl (Split ratio 15:1）

Inlet temp.

270℃

Pressure

97.3 kPa
100℃（hold 3 min）

Temperature profile

15℃/min to 270℃ (hold 4 min)
20℃/min to 300℃ (hold 5min）

CarrierHe, 1.3 ml/min

Ionization

Electron Impact (70 eV)

Scan range

15.0 m/z to 500.0 m/z

Hold time

5.0 min

5.2.6 Data analysis
A migration ratio of the organochlorines from the plastic sheets to the sediment and artificial
interstitial water was estimated as the basis of the initial and remaining concentration oforganochlorines on the plastics using the following equation:
Migration (%) =Remaining concentration of organochlorine on plastics
Initial concentration of organochlorine on plastics

82

×100 ……………….. (1)5.3 Result and Discussion
5.3.1 Migration of organochlorine
As shown in Figure 5.1(a), more than 90% of the adsorbed TriCP migrated from PE to
sediment after 24 hours of incubation; however, the PET sheet showed the highest migration
ratio of ClOD (Figure 5.1(b)). These results confirmed that the organochlorine adsorbed onto
the plastic sheets can migrate into the sediment; however, the migration behaviors weredependent on the properties of organochlorine and plastics. An aliphatic organochlorine, ClOD,
showed the highest migration ratio for PET; however, its migration from PVC and PE seems
to be the same. On the other hand, an aromatic organochlorine, TriCP, showed the lowest
migration ratio for PET. These might be due to their different adsorption mechanisms.
HOCs are adsorbed on plastics via hydrophobic interaction, - interaction, electrostaticinteraction, and hydrogen bonding [13]–[15]. In contrast, HOC desorption from plastics toanimals in gut conditions such as pH and temperature is enhanced; however, migration into
sediment is not elucidated [16]. The migration probability of HOCs was higher in sedimentfrom plastic surfaces because of the leaching of additives and interaction with interstitial waters.
The migration of TriCP from plastics to sediment was in the order of PE > PVC > PET (Figure
5.1 (a)) and PET > PE > PVC (Figure 5.1 (b)) for C1OD. These results provide evidence thatthat migration of organochlorine from plastic to sediment occurred, and that migration
behaviors depend on plastics and HOCs interaction.

83

100
PE
PET
PVC

Migration (%)

80
60

(a)

40
20
0
0

24
48
Incubation time [hour]

72

100
PE
80
Migration (%)

(b)

PET
PVC

60
40
20
0
0

24
Incubation time [hour]

48

72

Figure 5.1 Migration of (a) TriCP and (b) ClOD from plastics to sterile sediment (at pH 7.9
and DOC 50.54 mg/L). Bars indicate standard deviation (n=3).

84

5.3.2 Effect of DOC on the migration of organochlorine

100
80
Migration (%)

(a)

PE
PET
PVC

60
40
20
0
0

24

48
Incubation time [hour]

100

Migration (%)

(b)

PE
PET
PVC

80

72

60
40
20
0
0

24

48
Incubation time [hour]

72

Figure 5.2 Migration as dissolution of (a) TriCP and (b) ClOD from plastics to artificial
interstitial water (pH 8.0 and DOC 50 mg/L). The values were observed as the average (n=2).
As shown in Figure 5.2 (a), migration of TriCP and ClOD from all plastic sheets to the artificial
interstitial water was confirmed. When Milli-Q water was used, no migration was observed.

85

However, the migration behaviors were different from that than sediment. PET showed a higher
migration ratio for both TriCP and ClOD, to which the additives (Figure 2.7) might contribute.
In contrast, PVC showed different trends for TriCP and ClOD, reasonably indicating that the
migration of aromatic TriCP was greater than aliphatic ClOD from PVC. Although thedissolution trends could not be explained at this time, the results confirmed the dissolution oforganochlorine from plastic surfaces.

5.3.3 Migration pathways
In this experiment using sediment, PET showed 60% migration (Figure 5.1(a)). As for PE,
more than 90% migration was observed for TriCP in the sediment (Figure 5.1(a)); however,in the artificial interstitial water, the migration ratio of PE was less than 10% (Figure 5.2(a)).
This suggests that direct translocation by contacting with sediment particles may occur too. As
for ClOD on PE and PVC, direct translation may be negligible because of the similar migration
ratios in the sediment and artificial interstitial water (Figure 5.2 (a) and (b)).

MPs

DOC medium

MPs(Interstitial water)
Pollutant adsorbed onplastics

Biofilm formation

Adsorption

MPs

Sediment
Density changes

MPs

MPs

MPs
Figure 5.3 Plausible migration pathways of adsorbed organochlorines from plastics tosediment.
865.4 Conclusions
This study provided the first vital evidence that aliphatic and aromatic organochlorines, TriCP,
and ClOD can migrate from plastics to sediment, indicating plastics’ role in sedimentcontamination. As for migration mechanisms, dissolution followed by adsorption, directtranslation, and direct translocation to sediment may be possible. However, the proof of
migration of adsorbed HOCs from plastics to sediment via dissolution in interstitial water
followed by adsorption on sediment and or direct migration of pollutants from plastics tosediment remains primitive. In the future, further focus is needed to elucidate the realistic
migration of HOCs via plastic debris.

87

5.5 Literature cited[1] Mato Y., T. Isobe, H. Takada, H. Kanehiro, C. Ohtake, and T. Kaminuma, “Plastic Resin
Pellets as a Transport Medium for Toxic Chemicals in the Marine Environment,” Environ.
Sci. Technol., vol. 35, no. 2, pp. 318–324, Jan. 2001.
[2] Bergmann M., B. Lutz, M. B. Tekman, and L. Gutow, “Citizen scientists reveal: Marine
litter pollutes Arctic beaches and affects wild life,” Mar. Pollut. Bull., vol. 125, no. 1–2,
pp. 535–540, Dec. 2017.
[3] Coffin S., G.-Y. Huang, I. Lee, and D. Schlenk, “Fish and Seabird Gut Conditions Enhance
Desorption of Estrogenic Chemicals from Commonly-Ingested Plastic Items,” Environ.
Sci. Technol., vol. 53, no. 8, pp. 4588–4599, Apr. 2019.
[4] Graca B., K. Szewc, D. Zakrzewska, A. Dołęga, and M. Szczerbowska-Boruchowska,“Sources and fate of microplastics in marine and beach sediments of the Southern Baltic
Sea—a preliminary study,” Environ. Sci. Pollut. Res., vol. 24, no. 8, pp. 7650–7661, Mar.
2017.
[5] Cozar A. et al., “Plastic debris in the open ocean,” Proc. Natl. Acad. Sci., vol. 111, no. 28,
pp. 10239–10244, Jul. 2014.
[6] Chiba S. et al., “Human footprint in the abyss: 30 year records of deep-sea plastic debris,”Mar. Policy, vol. 96, pp. 204–212, Oct. 2018.
[7] Kanhai L. D. K., C. Johansson, J. P. G. L. Frias, K. Gardfeldt, R. C. Thompson, and I.
O’Connor, “Deep sea sediments of the Arctic Central Basin: A potential sink for
microplastics,” Deep Sea Res. Part Oceanogr. Res. Pap., vol. 145, pp. 137–142, Mar.
2019.

88

[8] Woodall L. C. et al., “The deep sea is a major sink for microplastic debris,” R. Soc. Open
Sci., vol. 1, no. 4, p. 140317, Dec. 2014.
[9] Ziajahromi S., A. Kumar, P. A. Neale, and F. D. L. Leusch, “Environmentally relevant
concentrations of polyethylene microplastics negatively impact the survival, growth andemergence of sediment-dwelling invertebrates,” Environ. Pollut., vol. 236, pp. 425–431,
May 2018.
[10] Teuten E. L., S. J. Rowland, T. S. Galloway, and R. C. Thompson, “Potential for Plastics
to Transport Hydrophobic Contaminants,” Environ. Sci. Technol., vol. 41, no. 22, pp.
7759–7764, Nov. 2007.
[11] Bakir A., I. A. O’Connor, S. J. Rowland, A. J. Hendriks, and R. C. Thompson, “Relative
importance of microplastics as a pathway for the transfer of hydrophobic organic
chemicals to marine life,” Environ. Pollut., vol. 219, pp. 56–65, Dec. 2016.
[12] Qiu Q., Z. Tan, J. Wang, J. Peng, M. Li, and Z. Zhan, “Extraction, enumeration andidentification methods for monitoring microplastics in the environment,” Estuar. Coast.
Shelf Sci., vol. 176, pp. 102–109, Jul. 2016.
[13] Liu P., X. Zhan, X. Wu, J. Li, H. Wang, and S. Gao, “Effect of weathering on
environmental behavior of microplastics: Properties, sorption and potential risks,”
Chemosphere, vol. 242, p. 125193, Mar. 2020.
[14] Hahladakis J. N., C. A. Velis, R. Weber, E. Iacovidou, and P. Purnell, “An overview ofchemical additives present in plastics: Migration, release, fate and environmental impact
during their use, disposal and recycling,” J. Hazard. Mater., vol. 344, pp. 179–199, Feb.
2018.

89

[15] Paluselli A., V. Fauvelle, F. Galgani, and R. Sempéré, “Phthalate Release from Plastic
Fragments and Degradation in Seawater,” Environ. Sci. Technol., vol. 53, no. 1, pp. 166–
175, Jan. 2019.
[16] Bakir A., S. J. Rowland, and R. C. Thompson, “Enhanced desorption of persistent organic
pollutants from microplastics under simulated physiological conditions,” Environ. Pollut.,
vol. 185, pp. 16–23, Feb. 2014.

90

Chapter 6

Conclusions
As a proof-of-concept, this study focused on the fate of HOCs adsorbed on plastic surfaces in
the ocean, investigating possible effects on the photodegradation behaviors of HOCs during
their floatation and migration from plastics to sediment after settlement. In thephotodegradation experiments, virgin and artificially aged PE, PET, and PVC sheets were
tested, according to their extensive use in the world, whereas PCP, TeCP, and TriCP CPs and
NAP were used as model HOCs. In the migration experiment, TriCP, together with an aliphaticorganochlorine, ClOD, were tested to find possible structure-specific effects. The highlights ofthis study are shown in Figure 6.1 as follows.
1) The virgin PE accelerated the photodegradation of PCP and TeCP; the photodegradation
rate constants on the PE surface were approximately 1.50 times higher than that on a glass
surface. In contrast, the photodegradation of these CPs on PET was slower than on the glass
surface. Furthermore, the photodegradation of PCP on the PVC surface was suspended after
approximately 60 min of ultraviolet irradiation. These results confirmed that plastics affect thephotodegradation behavior of HOCs adsorbed on their surfaces. FTIR analysis showed that
C=C bonds were produced on PE after UV irradiation, suggesting that PE might act as ahydrogen donor to accelerate the photodegradation of PCP and TeCP. The findings suggestthat the characteristics of plastics should be considered to understand the fate of HOCs
adsorbed on plastics through phototransformation.
2) NAP was photodegraded on the glass and PE; however, PE accelerated photodegradation of
NAP due to the possible hydrogen donation. As the photodegradation intermediates, oxy-PAHs,
such as NAPQ, were found on both glass and PE; however, the abundance of NAPQ on the PE
was much higher. As oxy-PAHs demonstrate higher toxicity than their precursors, these results

91

show that the remaining intermediates and precursor HOCs must be considered for risk
assessment.
3) On the virgin PE sheet, the accelerated photodegradation of TriCP was also confirmed;
however, on the aged PE, its photodegradation was suppressed rather than accelerated. In
addition, the PET and PVC sheets also showed further suppression of TriCP photodegradation
by aging, thereby confirming that plastic aging affects the photodegradation behaviors of theadsorbed TriCP. By aging, valley production and reduced plastic transparency to UV wereconfirmed, possibly contributing to the suppressed photodegradation on the aged plastic sheets.
On the other hand, the possible release of additives was expected to be positive for thephotodegradation of HOCs.
4) The migration of TriCP and ClOD to sediment was confirmed by burying the TriCP- andClOD-adsorbed PE, PET, and PVC sheets. The degree of migration of the aromaticorganochlorine, TriCP was in the order of the PE, PVC, and PET sheets, whereas the PET sheet
released an aliphatic organochlorine, ClOD, at the most abundant level. These different trendsof organochlorines in migration might be correlated with the plastic’s properties andelectrostatic interaction, such as π-π between plastics and pollutants. These results provided
reasonable evidence that microplastics may act as HOC carriers from seawater to the benthic
ecosystem to cause HOC sediment contamination.

Recommendations
Although this study revealed that plastics affect the photodegradation behaviors of HOCs andact as HOC carriers, understanding the possible effects of plastics on the fate of adsorbed
pollutants is still in a primitive stage. Comprehensive studies are needed to elucidate the fateof pollutants associated with plastics.

921) As the reported data are laboratory-based using UV, further studies should use actual
sunlight to estimate photopollutant photodegradation.
2) The effect of plastics additives and inbound plasticizers, stabilizer effects on thephotodegradation of adsorbed pollutants should be considered as these are not chemicallybonded with plastics and may hinder the pollutant photodegradation process.
3) As plastics are sunken and reached in sediment, anaerobic degradation of plastics andadsorbed pollutants need to be understood.
4) Heavy metals are also adsorbed on plastics along with HOCs. Due to a possible action as acatalyst, studies should reveal the effects of heavy metals on the photodegradation of HOCs.

93

(3) Aging of plastics changed photodegradation behaviorsof the adsorbed CPs. Change of the physical characteristics
such as surface shapes and transparency to UV might
contribute to the suppression.

MPs

UV

MPs

MPs

Photodegradation

MPs

MPs
Aging of plastics

MPs(1) PE, PVC and PET affected the photodegradation behaviors of adsorbed CPs. PE
accelerated photodegradation of CPs, but PET and PVC did suppress it.

Biofilm formation(2) PE accelerated photodegradation of NAP;

however, naphthacenequinone,oxygenated NAP was found on PE, indicating that existence of precursor HOCs together
with degradation intermediates should be considered for risk assessment.
Density changes

MPs(4) The organochlorines migrated from the plastics sheetMPs

to the sediment, thereby confirming that microplastics
can act as a carrier of HOCs to the benthic ecosystem.

MPs
Figure 6.1 Highlights of this study.
94

Acknowledgment
First and foremost, I would like to be grateful to the most generous, merciful, and creator of
the earth and heavens Allah, who gives me passion and patience to carry out my doctoral
research.
Second, I would like to express cordial thanks and humble gratitude to Prof. Dr. Satoshi Nakai
for his supervision, guidance, and right hand to carry out my doctoral research and writing
dissertation under his supervision.
I would like to express my special thanks to Prof. Dr. Wataru Nishijima for his sincere andvaluable comments which were necessary for me to conduct my research success.
Next but not least, I would like to be grateful to Dr. Takehiko Gotoh and Dr. Toshikazu Suenaga
along with all the members of Polymer Technology Laboratory who continuously supports andhelp me in the analysis and operation of Japanese instruments. During the last four years, thepolymer research group of Professor Satoshi Nakai and Waratu Nishijima was provided
continuous supports and a friendly milieu that gives me the spirit to carry out my research. I
tried to enjoy and learn new things every day in the past four years of studying with them.
I am particularly thankful to the Ministry of Education, Culture, Sports, Science, andTechnology (MEXT) for honoring me as a MEXT scholar to pursue my research for three years
and providing living expenses.
I would like to say thank you to my wife Kana Khatun and our son Imad al-Rafi because of
their strong motivation which was the fuel to finish my research.
I would like to convey a special thanks to my research mate Mr. Atsushi Kuzukami and all of
my friends who supported me in the experiment, writing and incentivized me to strive towards
my goal. Not but least I am grateful to all members of the polymer technology laboratory for95their cordial cooperation and help to execute the analysis and operation of Japanese analytical
instrument software.
Finally, I would like to thank my family for their unconditional support. Words cannot express
how grateful I am but I know that without their help I would not have been able to achieve
what I have achieved.

96List of publications
1) Hassan M. N., Kuzukami A., Nakai S., Nishijima W., Gotoh T. “Effect of Plastics on
the Photodegradation Behavior of Chlorophenols.” J. Chem. Eng. Japan. 2020, 53 (10),660–666. (Related to Chapter 2)2) Hassan M.N., Nakai S., Nishiima W., Gotoh T., Suenaga T. “Effects of the Aging of
Plastics on the Photodegradation of Adsorbed Organic Pollutants.” 環境科学会誌,
Japan, 2021, 34 (6), 1‒11. (Related to Chapter 4)International Conference
1) Hassan M. N., Nakai S., Nishijima W., Gotoh T., “Effects of aged plastics on
photodegradation behavior of sorbed organic pollutants.” Society of Environmental
Science, Japan, The 9th symposium, September 10th – 11th, 2021
2) Hassan M. N., Nakai S., Kuzukami A., Nishijima W., Gotoh T., Suenaga T.
“Photodegradation of Polycyclic Aromatic Hydrocarbons (PAHs) on Polyethylene.”
Water and Environmental Technology Conference, WET2021-online, 10 th -12th
August 2021 (Related to Chapter 3).
3) Hassan M.N., Nakai S., Kuzukami A., Nishijima W., Gotoh T. “Partitioning of organic
pollutants sorbed onto plastics to sediments.” In: The 6th International Conference on
Water Resource and Environment (WRE2020), August 23rd – 26th, 2020, Online
Conference. (Related to Chapter 5).

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