Microplastic Retention by Type in Several Species of Fish from the Great Lakes

by

Keenan Emily Munno

A thesis submitted in conformity with the requirements for the degree of Masters of Science

Ecology and Evolutionary Biology University of Toronto

© Copyright by Keenan Emily Munno 2017

ii

Microplastic Retention by Type in Several Species of Fish from

the Great Lakes

Keenan Emily Munno

Masters of Science

Ecology and Evolutionary Biology University of Toronto

2017

Abstract

Microplastics are plastic particles <5 mm in size. There are several types of microplastics. One

microplastic type, microbeads, were lost as a result of chemical digestion of fish tissues. An

assessment of chemical digestion methods found that temperature >60 °C melted some types of

microplastics. A room-temperature basic reaction was selected for use in feeding experiments.

Six microplastic types were fed to three species of freshwater fish in a laboratory. Some

microplastic types may have greater potential for retention in fish digestive tracts. Differences in

retention among microplastic types were significant for Rainbow Trout (Oncorhynchus mykiss)

but not for White Sucker (Catostomus commersonii) or Fathead Minnow (Pimephales promelas).

Polystyrene foam beads and spherical microbeads were observed in the esophagus and gills of

Rainbow Trout respectively, demonstrating potential for accumulation and blockage. Retention

of microplastics is a concern for potential physical effects in individuals and ecological effects

for impacted fish communities.

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Acknowledgments

I would like to thank my supervisor, Dr. Donald A. Jackson, for this incredibly rewarding

experience. His support, knowledge and experience have been instrumental in completing my

Masters thesis. I would also like to thank Dr. Paul Helm for his support, advice and expertise

throughout this process, and my committee member, Dr. C. Ken Minns, for his continued input

and support. I owe a great deal of gratitude to Chelsea Rochman for all of her advice.

Thank you to all of the Jackson Lab for all of their input and assistance throughout this process.

Thanks to Dave Poirier, Richard Chong-Kitt, Satyendra Bhavsar and Kathleen Stevack for their

invaluable help and assistance throughout this entire project. I would also like to thank Alina

Sims for interpreting the FT-IR spectra, Garret Zimmer for recording FT-IR spectra and assisting

with microplastic particle preparation, and Giuseppe Gigliotti for conducting additional boiling

tests on microbeads from a variety of commercially-available personal care products, and for

assisting with sample processing in the fish feeding studies.

I would like to thank the Ontario Ministry of the Environment and Climate Change (MOECC),

the Toronto Region Conservation Authority and the University of Toronto for funding,

assistance and resources used in this project.

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Table of Contents Acknowledgments   iii  

Table  of  Contents   iv  

List  of  Tables   vi  

List  of  Figures   vii  

List  of  Appendices   viii  

Chapter  1   1  

  Literature  Review   1  1

Chapter  2   12  

  Assessing  Chemical  Digestion  Methods  for  the  Recovery  of  Microplastics   12  2

2.1   Introduction   12  

2.2   Materials  and  Methods   16  

2.3   Results   20  

2.4   Discussion   26  

2.5   Conclusions   29  

Chapter  3   31  

  Microplastic  Ingestion  and  Retention  by  Type  in  Three  Species  of  Fish  from  Lake  Ontario   31  3

3.1   Introduction   31  

3.2   Materials  and  Methods   39  

3.3   Results   45  

3.4   Discussion   56  

3.5   Conclusions   67  

Chapter  4   68  

  Conclusions  and  Future  Directions   68  4

4.1   Conclusions   68  

4.2   Changes  to  Experimental  Design   69  

4.3   Future  Directions   69  

Literature  Cited   72  

v

Appendices   81  

vi

List of Tables Table  1.  Percent  recovery  across  four  different  chemical  digestion  methods   22  Table  2.  Sum  of  microplastic  counts  for  Rainbow  Trout   49  Table  3.  Sum  of  microplastic  counts  for  White  Sucker   53  Table  4.  Sum  of  microplastic  counts  for  Fathead  Minnow   56  

vii

List of Figures Figure  1.  Mean  %  recovery  for  chemical  digestion  methods   21  Figure  2.  FT-­‐IR  spectra  for  SB3  and  SB1   25  Figure  3.  Summary  of  the  mean  number  of  microplastics  of  each  type  observed  in  water  after  microplastic  

exposure  for  Rainbow  Trout   47  Figure  4.  Mean  mean  number  of  microplastics  of  each  type  observed  in  water  and  digestive  tracts  of  Rainbow  

Trout   48  Figure  5.  Polystyrene  foam  beads  in  one  Rainbow  Trout   49  Figure  6.  Mean  mean  number  of  microplastics  of  each  type  observed  in  water  digestive  tracts  of  White  Sucker   51  Figure  7.  Summary  of  the  mean  number  of  microplastics  of  each  type  observed  in  water  after  microplastic  

exposure  for  White  Sucker   52  Figure  8.  Summary  of  the  mean  number  of  microplastics  of  each  type  observed  in  water  after  microplastic  

exposure  for  Fathead  Minnow   54  Figure  9.  Mean  number  of  microplastics  of  each  type  observed  in  water  and  digestive  tracts  of  Fathead  Minnow   55  Figure  10.  Polystyrene  foam  beads  accumulated  at  the  top  of  the  esophagus  in  one  Rainbow  Trout   58  

viii

List of Appendices A.  1.  Body  size  of  fish  used  in  feeding  experiments   81  A.  2.  G-­‐test  results  of  feeding  experiments   81  A.  3.  The  mean  number  of  microplastic  particles  in  each  phase  for  Rainbow  Trout   82  A.  4.  The  mean  number  of  microplastic  particles  in  each  phase  for  White  Sucker   83  A.  5.  The  mean  number  of  microplastic  particles  in  each  phase  for  Fathead  Minnow   84  A.  6.  Examples  of  microplastic  types  used  in  the  assessment  of  chemical  digestion  methods  in  Chapter  2   85  A.  7.  Experimental  fibre  next  to  a  contamination  fibre   86  

1

Chapter 1

Literature Review 1With increasing population size and changes in life style, humans produce large amounts of

waste, including plastic waste. Just 192 coastal countries generated approximately 275 million

tonnes of plastic waste in 2010 (Jambeck et al. 2015). By 2025, this amount is expected to

increase by an order of magnitude (Jambeck et al. 2015). In litter, a variety of different sizes and

shapes of plastic debris have been identified. Macroplastics (>200 mm) are easily visible to the

human eye, including large items like discarded water bottles, plastic bags and food containers

(Eriksen al. 2013). Mesoplastics are slightly smaller in size (5-200 mm), though typically still

visible with relative ease (Eriksen et al. 2013). Microplastics are defined as plastic particles less

than five millimeters in size in their largest dimension (Arthur et al. 2009). The lower bound is

defined often by the method of collection; many field collections use a mesh or sieve size of 0.33

mm (Arthur et al. 2009).

It is estimated that roughly 5.25 trillion floating particles of microplastics were present on the

ocean surfaces globally in 2013 (Eriksen et al. 2014). A number of polymer types have been

identified in microplastics found in the environment including polyethylene, polypropylene,

polystyrene, polyurethane, polyvinylchloride (PVC), and polyethylene terephthalate (PET)

(Moore et al. 2001; Thompson et al. 2004; Rios et al. 2007; Castaneda et al. 2014; Yonkos et al.

2014; Corcoran et al. 2015; Ballent et al. 2016; Mason et al. 2016b; Sutton et al. 2016).

Typically, polyethylene is the most abundant polymer type observed in the environment

(Castaneda et al. 2014; Yonkos et al. 2014; Corcoran et al. 2015; Ballent et al. 2016; Mason et

2

al. 2016b). This is likely because polyethylene is one of the most commonly produced polymer

types, accounting for 53% of plastic polymer production in the United States (Jambeck et al.

2015). Five common categories used when identifying different microplastic types are beads,

fragments, foam, fibre/line and pellets (Eriksen et al. 2013; Ballent et al. 2016; Mason et al.

2016b; Sutton et al. 2016). Primary microplastics are intentionally produced in their observed

form, often used in cleansers, exfoliants and other consumer products, or as pre-production

pellets and industrial abrasives (Arthur et al. 2009). Secondary microplastics result from the

breakdown of larger plastics through weathering, ultraviolet radiation and other energetic forces

(Arthur et al. 2009).

Primary microplastics come from a range of consumer products and industrial sources.

Polyethylene and polypropylene pellets are most commonly used in the plastic industry for the

production of larger plastic products (Eerkes-Medrano et al. 2015). A number of microplastic

types are classified as fragments, including shavings and irregularly shaped particles found in

cleansers. Shavings appear to be the result of removal of extraneous material from the seams and

edges of solidified plastic products, typically made of polyethylene and polypropylene (Ballent

et al. 2016). The microplastics isolated from several cleansers were present in a number of

shapes including ellipses, ribbons, threads and irregular fragments and microbeads, with

polyethylene being the dominant polymer (Napper et al. 2015). Microplastics in rinse-off

personal care products are referred to as microbeads (Canada 2016), and are often in the shape of

small, spherical beads and irregularly shaped beads resembling an agglomeration of smaller

spherical beads. Microbeads are also used in sandblasting (Eriksen et al. 2013). The dominant

polymer in microbeads in St. Lawrence River sediments was polyethylene (Castaneda et al.

3

2014); however, some polystyrene sulfonate (PSS) beads used in ion exchange for water

purification and softening were found in Lake Ontario sediments (Ballent et al. 2016).

Microbeads have been the subject of controversy in recent years and, as a result, those

microbeads used in rinse-off personal care products will be banned in the United States and in

Canada beginning in 2018 (United States Congress 2015; Canada 2016). Like microbeads that

are rinsed down the drain, many microplastics may be introduced to aquatic ecosystems through

wastewater treatment plants. Mason et al. (2016a) estimated that 13 billion microbeads are

released into United States waterways each day from these facilities. In Lake Ontario,

microbeads made up approximately 30% of microplastics in wastewater treatment plant effluent,

and 14% of microplastics in Lake Ontario surface water samples (Ontario 2016). While

microbeads are not often a major component of microplastics detected in water samples, they are

still released into aquatic ecosystems in high abundance. Napper et al. (2015) estimated that

4,594-94,500 microplastics could be rinsed down the drain and into wastewater treatment plants

with each use of a microplastic-containing cosmetic product, resulting in 16-86 tonnes of

polyethylene emitted per year for the entire United Kingdom population. Primary microplastics

can persist in their original form, or can break down into secondary microplastics.

Secondary microplastics originate from many potential sources and can account for a number of

the fragments, film and fibres observed in the environment. The polymer types of fibres are

challenging to detect because of their diameter and volume; however, those that have been

characterized are polyamide (nylon), polyethylene, polypropylene, polyester and rayon

(Thompson et al. 2004; Lusher et al. 2013; Ballent et al. 2016). Fibres are often one of the most

abundant microplastic types observed in aquatic environments (Thompson et al. 2004; Mason et

4

al. 2016b; Sutton et al. 2016). Several studies also note that fragments are the dominant

microplastic type observed in their respective study regions (Moore et al. 2001; Eriksen et al.

2014; Yonkos et al. 2014; Mason et al. 2016a). In some instances, fibres are more dominant in

one area of their study site, while fragments are more dominant in a different area (Ballent et al.

2016), which likely indicates varying relative contributions of potential microplastic sources in

the same ecosystem. Of the fragments characterized in the environment, polyethylene accounts

for the majority (Yonkos et al. 2014; Corcoran et al. 2015; Ballent et al. 2016; Mason et al.

2016b). Polyurethane, also a common fragment polymer, is used in production of foam for

construction of furniture and coating a number of products (Ballent et al. 2016). Broken-down

plastic bags are a likely source of many film-like particles classified as fragments (Eriksen et al.

2014). It has been estimated that one plastic object that is 200 mm in its largest dimension could

fragment into up to sixteen particles that are 50 mm in their largest dimension, and one 50 mm

particle could generate 625 large microplastics (2 mm in their largest dimension); this 2 mm

microplastic could fragment into approximately 6 smaller microplastic particles (Eriksen et al.

2014). Fragmentation of larger plastics reduces the particle size, which may reduce particles

below the size commonly detected using current methods (0.33 mm) (Eriksen et al. 2014).

Because of this sampling bias, there are likely far more fragmented microplastics present in the

environment than what current sampling practices commonly capture. Clearly, it is challenging

to quantify microplastics present in the environment.

Carpenter et al. (1972) were the first to attempt to quantify microplastics in water and they found

that microplastics occurred at an average of 3,500 particles per km2, with a range from 50-12,000

particles per km2 in surface waters near New York City. Plastics were ubiquitous in all 11

5

samples collected, with most described as cylindrical pellets, and the majority of others as

fragments (Carpenter et al. 1972). Plastics were collected in a similar manner as surface samples

are today, using net tows with a 0.33 mm mesh size (Carpenter & Smith 1972). Since then, a

number of studies have been conducted in an attempt to quantify microplastic concentrations in

various marine environments following essentially the same strategy. In areas where currents

converge, including the North Pacific Central Gyre (NPCG), debris accumulates and continues to

fragment into smaller pieces (Boerger et al. 2010). In the NPCG, a mean concentration of

334,271 particles per km2 was found, with a range of 31,982 particles per km2 to 969,777

particles per km2 (Moore et al. 2001). In the Mediterranean, up to 890,000 particles per km2 were

found (Eriksen et al. 2014). Five sub-tropical gyres (North Pacific, North Atlantic, South Pacific,

South Atlantic, Indian Ocean) were conservatively estimated for microplastic concentration,

leading to a global estimate of 5.25 trillion microplastic particles weighing 268,940 tonnes in

ocean surface waters (Eriksen et al. 2014). Of 680 tows, 70% had concentrations of

microplastics in the range of 1000-100,000 particles per km2, and approximately 92% of all tows

contained plastics (Eriksen et al. 2014). Evidently, microplastics are present in marine surface

waters globally, and ranges in concentrations vary over space and time.

More targeted studies have been conducted as a means of quantifying microplastics in areas

where microplastics are expected to be in high concentration, such as areas in close proximity to

major urban centres. Chesapeake Bay, the largest estuary in the United States, ranged from <1 g

microplastics per km2 to 563 g microplastics per km2 in surface waters, and microplastics were

found in 59 of 60 samples collected (Yonkos et al. 2014). Increased microplastic concentrations

were observed with increasing proximity to urban and suburban areas (Yonkos et al. 2014).

6

Three of four sites sampled displayed peaks in microplastic abundance following two major

storm events, suggesting that wind, precipitation and terrestrial runoff may add to, and resuspend

microplastics in aquatic systems (Yonkos et al. 2014), perhaps washing terrestrial microplastics

from surrounding urban centres into Chesapeake Bay leading to higher concentrations observed

following storm events. Sutton et al. (2016) characterized microplastics in treated wastewater

effluent from 8 facilities discharging into the San Francisco Bay and determined that they

discharged a total of 56 million microplastics per day for all facilities tested. The region of the

bay receiving higher volume of treated wastewater had higher concentrations of microplastics

(1,000,000 particles per km2) versus the central part of the bay (310,000 particles per km2)

(Sutton et al. 2016), indicating that areas receiving higher input from anthropogenic activities

lead to higher microplastic concentrations. Mason et al. (2016a) sampled 17 wastewater

treatment plants across the United States and determined that these plants release an average of

approximately 4,400,000 microplastics per day, with fibres being the most common (53%),

followed by fragments (33%). The types of microplastics present in wastewater treatment plant

effluent depend on human population size, land use, contribution of sewer systems, flow rate

through the plant, complexity of filtration systems and sources (Mason et al. 2016a). In

summary, studies examining a number of marine ecosystems around the world concluded that

ranges in concentration of microplastics vary from location to location, but are generally greater

with closer proximity to urban centres and anthropogenic sources (wastewater effluent, urban

runoff and industry).

While the majority of the focus has been on marine ecosystems, some studies have been

conducted in freshwater ecosystems. Microplastics have been observed in surface waters of the

7

Great Lakes. Lake Michigan had an average concentration of microplastics in the surface waters

of 17,276 particles per km2, with a range from 1,400-100,000 particles per km2 (Mason et al.

2016b). An average of 43,157 particles per km2 was observed in surface waters of the Laurentian

Great Lakes (Huron, Superior and Erie); however, the range went from 450 particles per km2 to

466,000 particles per km2 (Eriksen et al. 2013). There was a great deal of variation among the

lakes. Samples from Lake Erie were consistently the most concentrated, accounting for ~90% of

all plastic debris collected from surface waters (Eriksen et al. 2013). This observation is

consistent with the Lake Erie basin being the heavily populated relative to the other Laurentian

Great Lakes (Huron, Superior and Erie). Lake Ontario, which is the most urbanized of the

Canadian side of the Great Lakes overall, was found to have up to 6.7 million particles per km2,

with Humber Bay samples having the highest microplastic counts (Ontario 2016). The samples

taken from the Humber Bay region were in very close proximity to the City of Toronto (Ontario

2016). There appears to be a link between urbanization and microplastic concentration within the

Great Lakes surface waters.

Attempts are being made to quantify microplastic concentrations in water, but not all

microplastics are contained within the surface waters that can be sampled by trawling.

Microplastics may be lost from surface waters due to stranding on shores, ingestion by

organisms, biofouling, being trapped in detritus or degraded: ultimately, the fate of some

microplastics is not at the water surface (Eriksen et al. 2014). In the NPCG, tows conducted at a

10 m depth for subsurface concentration of microplastics found the concentration to be less than

half of surface trawls from the same regions (Moore et al. 2001). Surface trawls contained

mostly fragments, while the subsurface tows contained more monofilament line (Moore et al.

8

2001). This increase in fibrous microplastics may be related to decreased buoyancy of the fibrous

particles relative to other microplastic types. When more buoyant microplastics appear to sink, it

may be the result of a process referred to as biofouling. Carpenter & Smith (1972) observed

microbiotic communities of hydroids and diatoms on the surface of many plastics. Biofouling

can cause buoyant microplastics to sink below the surface, and potentially become buried in

sediment (Corcoran et al. 2015; Ballent et al. 2016), such as microbeads. Microbeads are

typically buoyant, but a mean of 52 microbeads per m2 were observed in St. Lawrence River

sediments (Castaneda et al. 2014). Buoyancy and biofouling are two factors that may contribute

to the accumulation of microplastics in sediments.

In Lake Ontario, polyethylene accounts for the majority of microplastic polymer types found in

sediment cores (74%), including open-lake, deep-water sedimentation zones (Corcoran et al.

2015). In nearshore areas of Lake Ontario, sediment cores of up to 15 cm in depth had

microplastics that were ubiquitous at all depths (Ballent et al. 2016). The portion of the Lake

Ontario watershed studied contained 20 major wastewater treatment plants (Ballent et al. 2016).

Five of the 66 watersheds studied along Lake Ontario tributaries (Etobicoke Creek, Mimico

Creek, Humber River and Don River) contained 40% of the population in that region and half of

the plastic manufacturing plants in the region (Ballent et al. 2016). Of the tributaries, Etobicoke

Creek had the greatest abundance of plastic-related industrial establishments and manufacturers

(Ballent et al. 2016). Sediments and beaches from Lake Ontario and tributaries had an average

concentration of 760 microplastics per kg of sediment, and ranged from 20-27,830 microplastics

per kg of sediment (Ballent et al. 2016). Nearshore sediment samples contained the greatest

average concentration of microplastics (980 microplastics per kg of sediment), and were most

9

concentrated in sediments collected from Humber Bay and Toronto Harbour (Ballent et al.

2016). Nearshore sediment samples contained microplastics mostly <2 mm in size, with fibres

and fragments being the dominant component (Ballent et al. 2016). This is consistent with sub-

tidal sediment samples from Plymouth, United Kingdom, where plastic polymers were identified

in 23 of 30 samples and the majority were brightly coloured fibres (Thompson et al. 2004). Lake

Ontario tributary samples (610 particles per kg of sediment) and beach samples (140 particles per

kg of sediment) were less concentrated on average than nearshore Lake Ontario samples, with

abundance in beach samples decreasing with increasing distance from Toronto (Ballent et al.

2016). The relatively high abundance of microplastics in sediments closest to Canada’s largest

urban center indicates that urbanization likely plays a major role in determining microplastic

output in sediment as well as surface waters of the Great Lakes.

The potential for microplastics to degrade once they have been buried in sediment is low

(Corcoran et al. 2015). According to sediment accumulation rates, microplastics first began to

appear in Lake Ontario either 18 or 38 years prior to 2015, with the greatest concentrations being

in the most-recent sediment increments (Corcoran et al. 2015). Because the rates of microplastic

accumulation appear to be increasing and degradation potential is low, microplastics may persist

in aquatic ecosystems for a considerable amount of time. This is particularly problematic for

aquatic ecosystems in close proximity to major urban centres, such as those surrounding the

Great Lakes, as the inputs appear to be high.

10

Microplastics are present in surface waters, in the water column and in sediments, and are readily

available for ingestion by aquatic organisms. Less is known about the actual impacts of

microplastics on aquatic organisms, and how that may affect aquatic communities. A wide

variety of organisms were observed to have ingested microplastics, including zooplankton (Cole

et al. 2013), bivalves (Van Cauwenberghe & Janssen 2014), and several species of fish (Boerger

et al. 2010; Fossi et al. 2012; Foekema et al. 2013; Lusher et al. 2013; Fossi et al. 2014; Avio et

al. 2015; Romeo et al. 2015). Observations of microplastic ingestion in the wild have prompted

studies addressing microplastic ingestion by fish in laboratory settings (Hoss & Settle 1990;

Rochman et al. 2014; Lönnstedt & Eklöv 2016; Grigorakis et al. 2017). While these lab studies

have demonstrated that fish ingest microplastics, less is known about differences among the wide

array of microplastic types that were observed in the environment in water, sediment and fish.

This thesis addresses two main objectives related to microplastic ingestion by freshwater fish.

Different types and sources of microplastics have varying shapes (or morphology), which may

behave in different ways in the environment and within organisms when ingested. First, it is

essential that chemical digestion methods used to isolate microplastics from biological tissues are

effective in ensuring microplastics are recovered from organisms that ingest them. Given the

diversity of microplastic sources and variety of microplastic types in freshwater ecosystems, it is

important to determine whether some microplastic types may be retained more or less in fish

digestive tracts. Microplastic types implicate different potential sources, so microplastics with

greater potential for retention in fish digestive tracts should be the most immediate targets for

management and policy development. These objectives are related to the broader goal of

determining whether there are any general patterns that make some individual fish more

11

susceptible to hazards of anthropogenic origins than others, and how humans are impacting

aquatic ecosystems.

12

Chapter 2

Assessing Chemical Digestion Methods for the 2Recovery of Microplastics

2.1 Introduction

The majority of the plastic debris present in aquatic environments is microplastic (Eriksen et al.

2014). With increasing concern regarding microplastics in the environment, questions have been

raised about the abundance of microplastics in several environmental matrices. Microplastic

contamination has been observed in marine and fresh water (Moore et al. 2001; Eriksen et al.

2013; Mason et al. 2016b), sediments (Castaneda et al. 2014; Corcoran et al. 2015; Ballent et al.

2016), fish (Boerger et al. 2010; Lusher et al. 2013; Foekema et al. 2013; Avio et al. 2015;) and

filter-feeding invertebrates (Cole et al. 2013; Van Cauwenberghe & Janssen 2014). The methods

used to quantify microplastics vary greatly from study to study raising concerns over the

accuracy of these methods, and whether the abundances estimated are under- or over-

representative of actual environmental contamination. A number of studies have investigated the

efficacy of several of these methods used to quantify and identify microplastics in these various

matrices (Cole et al. 2014; Collard et al. 2015; Catarino et al. 2016; Dehaut et al. 2016).

Methods vary in extraction techniques (e.g. manual sorting, density separation, chemical

digestion and enzyme digestion) and identification techniques (e.g. visual identification, Fourier

transform infrared spectroscopy (FT-IR), Raman Spectroscopy and Scanning Electron

Microscopy (SEM)). In some cases, extraction and identification techniques seem to vary

depending on access and availability of instrumentation.

13

Early methods for separating and identifying microplastics from organic matter involved manual

sorting and visual identification with or without microscopy (Hoss and Settle 1990; Carpenter et

al. 1972; Moore et al. 2001; Boerger et al. 2010). While these methods continue to be used,

small particles can be missed completely, or misidentified as plastic when they are actually not.

An alternative involves staining stomach contents with rose bengal which turns the majority of

organic matter pink, and plastics and other inorganic matter should remain unaffected (Davison

& Asch 2013); however, some plastics are pink in colour and may not be distinguishable from

dyed organic matter.

Density separations have also been proposed as an alternative for extracting microplastics in

sediments (Thompson et al. 2004). This involves one or more floatations of samples containing

microplastics in a hypersaline sodium chloride (NaCl) solution followed by decanting and

sieving (Thompson et al. 2004, Avio et al. 2015). Another frequently used alternative to NaCl in

density separations that typically does not damage polymers is sodium polytungstate (SPT)

(Corcoran et al. 2015; Ballent et al. 2016). Fouling by organic and inorganic materials can alter

the density of microplastics thus affecting their buoyancy, requiring subsequent manual sorting

of the non-buoyant fractions (Ballent et al. 2016).

Chemical digestion methods were proposed to solve some of these issues. They were developed

to digest natural organic matter and leave behind particles that are more easily identifiable as

microplastics, which are more resistant to the chemical digestion processes. Due to a need for

standardized methods for extracting microplastics, the NOAA (National Oceanic and

14

Atmospheric Administration) Marine Debris Program produced a laboratory methods guide for

the analysis of microplastics in water and sediment (Masura et al. 2015). Their methods are

based on a wet peroxide oxidation reaction to remove organic matter with 30% hydrogen

peroxide (H2O2) and an iron (Fe(II)) catalyst, applying heat during digestion and in drying

samples. Modified versions of this method have been used in a number of studies (Yonkos et al.

2014; McCormick et al. 2014; Mason et al. 2016b). The H2O2 digestion with heat has also been

applied to fish tissues, followed by a potassium hydroxide (KOH) digestion (S. Mason- SUNY

Fredonia, personal communication, Sept. 1, 2015). Foekema et al. (2013) used an alkaline 10%

KOH digestion over 2-3 weeks to break down fish digestive tracts. Sodium hydroxide (NaOH)

has been used in place of KOH as it has a similar base dissociation constant (pKb) (Cole et al.

2014; Catarino et al. 2016). Increasing temperature (60 oC) and molarity (2M) increased

efficiency; however, the solution degenerated aluminum used during the procedure leading to

possible contamination (Cole et al. 2014). Nitric acid (HNO3) has also been used to digest

mussel tissue (Claessens et al. 2013; Van Cauwenberghe & Janssen 2014). Claessens et al.

(2013) found a 22.5 M HNO3 solution with boiling water was effective. However, it has been

suggested that methods using HNO3 may be problematic, as some plastics are degraded

(polylauryllactam) and low-density polyethylene can yellow (Dehaut et al. 2016). Other strong

acids, such as hydrochloric acid (HCl), have been used in digestions as well, although HCl was

found to be the least effective method of those tested for biota-rich seawater samples, while

sodium hydroxide was slightly more effective (Cole et al. 2014). Collard et al. (2015) used a

combination of sodium hypochlorite (NaClO) and HNO3 rinses, followed by ultrasonification in

a methanol (CH3OH) solution to digest contents of fish stomachs, suggesting it to be effective,

although mass loss of 25% for PVC particles was observed. There remains no clear consensus on

the most effective chemical digestion method, and optimization of chemical extraction methods

15

to isolate microplastics from organic substrates in environmental samples remains an area of

active research.

Enzyme digestions have been proposed as an alternative to chemical digestions for extraction of

microplastics. Cole et al. (2014) used Proteinase K and Chitinase to digest >97% of organic

matter in copepod samples containing microplastics, with no observable effects on the

microplastics. Catarino et al. (2016) found industrial enzymes (protease Corolase 7089) to be

effective in digesting soft tissue mussel samples. However, the application of enzyme-based

digestions to a broader range of organisms and substrates needs further evaluation, as costs and

more complex tissues (crustaceans, larger fish) needing enhanced methods remain as

considerations (Catarino et al. 2016).

In the course of conducting preliminary recovery tests for a digestion method for fish digestive

tracts, microbeads isolated from a consumer product were lost during treatment. Due to the

potential for bias in reporting of abundances and categories of microplastics that could result

from loss during extractions, a more systematic evaluation of candidate digestion protocols was

undertaken. We evaluated recoveries of several types of microplastics using wet peroxide and/or

alkaline chemical digestion methods, as well as the influence of heat in simple boiling tests with

water. Microplastics from each test were analyzed using FT-IR to determine if the extraction

procedures impacted the ability to identify the polymers.

16

2.2 Materials and Methods

Microplastic Particles

Three types of spherical microbeads (SB1, SB2 and SB3), irregular-shaped microbeads and

fragments were isolated from different rinse-off personal care products readily available in

Ontario, Canada. They were isolated by rinsing through a 125 µm metal sieve using deionized

water. Shavings were mechanically generated from a polyethylene block using a pipe-threader in

a drill press. Polystyrene foam beads were taken from a sheet of polystyrene foam insulation

board. Synthetic fibres were cut from a nylon carpet sample. These microplastics (shapes and

composition) were chosen as they were being used in subsequent feeding experiments.

Digestion Treatments

Microplastic particles of each type were prepared (Ni=20) for each of three independent

replicates per treatment. The treatments consisted of a control containing the microplastics in

room temperature deionized water and no chemicals added. The control was also used as a

means of assessing potential contamination through the isolation, handling and counting stages.

The remaining treatments consisted of boiling deionized water, and four different chemical

digestion methods, selected based on common use and demonstrated efficacy and efficiency for

biotic samples. One chemical digestion method uses a digestion in 1N KOH at room temperature

for 14 days that was modified to exclude the addition of heat by Rochman et al. (2015),

originally adapted from Foekema et al. (2013). We further modified this method, using 4N KOH

at room temperature for 14 days. The wet peroxidation method adapted from the NOAA

protocols (Masura et al., 2015) was included, as was a digestion combining 4N KOH and H2O2

17

obtained from S. Mason, SUNY Fredonia (personal communication, Sept. 1, 2015) for digestion

of fish tissues. Each type of microplastic was transferred from a 125 µm metal sieve to plastic

vials or beakers with deionized water then treated as outlined below.

Control (Room Temperature). Microplastic particles transferred into polypropylene vials filled to

a minimum of 15 mL with deionized water, covered, and left standing in a fumehood at room

temperature for 14 days.

Boiling. Deionized water (30 mL) was added to the microplastics in 250 mL glass beakers,

heated on a hot plate to 100 °C and allowed to boil for 10 minutes. Contents were then sieved

while hot.

1N and 4N KOH. A minimum of 15 mL of either 4N KOH or 1N KOH, prepared from 85%

(w/w) KOH pellets (Sigma-Aldrich, Oakville, ON, Canada), was added to the microplastics in

polypropylene vials, covered, and left to stand for 14 days in a fumehood.

Fe2SO4 + H2O2. Under a fume hood, aliquots of 20 mL of a solution of Fe(II)SO4 (Sigma-

Aldrich, Oakville, ON, Canada) and 20 mL of 35% H2O2 (Sigma-Aldrich, Oakville, ON,

Canada) were added to the microplastics in 250 mL beakers. Once the reaction settled (no

boiling or bubbling), an additional 20 mL of 35% H2O2 was added, and this step was repeated

18

until a total of 5 aliquots had been added. Afterwards, contents of the beaker were rinsed into a

125 µm metal sieve. The contents of the sieve were rinsed back into the original beaker and

soaked in a 5:1 deionized water:Contrad 70® liquid detergent (Fisher Scientific, Ottawa, ON,

Canada) solution while covered for up to 24 hours, or until any crusted material from the

reaction had dissipated.

4N KOH + H2O2. Under a fume hood, a 30 mL aliquot of 4N KOH solution was added to the

microplastics in 600 mL beakers, which were placed on magnetic stir/hot plates and covered.

The samples were heated to 60 °C, stirred for one hour and then removed from the heat. Once the

reaction finished, a 5 mL aliquot of 35% H2O2 was added to each, and the solutions stirred for 15

minutes. The beakers were removed from the stir/hot plate and allowed to sit covered for two

hours.

Post-treatments, all samples were transferred to a 125 µm metal sieve, rinsed thoroughly with

deionized water, then transferred to glass petri or aluminum dishes and dried in an oven at 60 °C

for 12 hours, or captured on a 10 µm polycarbonate filter (Fisher Scientific, Ottawa, ON,

Canada). The remaining microplastic particles were counted under a dissecting microscope

(Leica S8 APO Stereozoom; Leica Microsystems Canada, Inc., Richmond Hill, ON, Canada) at

10-80 times magnification.

19

FT-IR Analysis

To indicate whether treatments impacted the polymer materials, selected microplastic particles

(n=3) of each type subjected to each treatment were analyzed by FT-IR using a VERTEX 70-

Platinum ATR Infrared spectrometer (Bruker Optics Ltd, Milton, ON, Canada) operating in

attenuated total reflectance mode.

Data Analysis

Percent recovery was determined by recording the number of microplastics remaining in each

sample (Nf) following exposure to each method using the following equation:

%  𝑅𝑒𝑐𝑜𝑣𝑒𝑟𝑦 =  𝑁!𝑁!

×  100%

Three independent replicates were used in each test. The mean percent recovery and standard

deviation were calculated for triplicates from each treatment.

A Kruskal-Wallis rank sum test, a non-parametric test similar to a one-way Analysis of Variance

(ANOVA), was performed in R to determine whether there were statistically significant

differences (α = 0.05) among groups of microplastic types for all tests.

20

2.3 Results

Differences between extraction methods

Recoveries among microplastic types were significantly different as confirmed by the Kruskal-

Wallis rank sum test (χ2 = 27.883, 7 d.f., α=0.05, p < 0.001). Differences were driven by SB1

and SB3 (Figure 1). Complete or nearly complete recoveries (95-100%) of particles were

observed across all shapes and material types for the negative control (room temperature water)

and the 4N KOH and 1N KOH treatments (Table 1). Recoveries exceeding 100% can likely be

attributed to breakage of some particles during processing, resulting in their being counted as

more than one particle. Losses may have occurred if broken pieces were small enough to pass

through the sieve, or during transfers. Some minor discoloration of particles, particularly SB1

and SB2, was noted during microscope examination.

21

Figure 1. Mean % recovery of microplastic types (n=3) from different consumer and industrial products across a range of chemical digestion methods, boiling water (~100 °C) and a room temperature water control. Microplastics tested include spherical microbeads (SB1, SB2, SB3),

22

irregular-shaped microbeads (IB), fragments (F), shavings (S), polystyrene foam beads (PSF) and synthetic fibres (SF).

Table 1. Percent recovery (mean ± standard deviation) for spherical microbeads (SB1, SB2, SB3) irregularly shaped microbeads (IB), fragments (F), shavings (S), polystyrene foam beads (PSF) and synthetic fibres (SF) across four different chemical digestion methods, boiling water and one control treatment (n = 3).

Control (Room Temperature)

Boiling 1N KOH 4N KOH Fe(II)SO4 + H2O2

4N KOH + H2O2

SB1 100.0 ± 0.0 0.0 ± 0.0 100.0 ± 0.0 95.0 ± 0.0 0.0 ± 0.0 3.3 ± 5.8

SB2 98.4 ± 2.7 96.7 ± 2.8 100.0 ± 0.0 100.0 ± 0.0 100.0 ± 0.0 98.3 ± 2.9

SB3 100.0 ± 0.0 11.7 ± 10.4

101.7 ± 2.9 105.0 ± 5.0 5.0 ± 5.0 100.0 ± 0.0

IB 100.0 ± 5.0 95.0 ± 5.0 105.0 ± 5.0 100.0 ± 5.0 100.0 ± 0.0 100.0 ± 0.0

F 98.3 ± 2.9 100.0 ± 0.0

98.3 ± 2.9 100.0 ± 0.0 98.3 ± 2.9 100.0 ± 0.0

S 98.3 ± 2.9 101.7 ± 2.9

100.0 ± 5.0 98.3 ± 2.9 98.3 ± 2.9 100.0 ± 0.0

PSF 101.7 ± 2.9 100.0 ± 0.0

98.3 ± 2.9 100.0 ± 0.0 98.3 ± 2.9 100.0 ± 0.0

SF 105.0 ± 5.0 103.3 ± 5.8

100.0 ± 10.0 100.0 ± 0.0 100.0 ± 0.0 100.0 ± 0.0

Control (Room Temperature). For almost all microplastic types, 100% recovery was observed

(Table 1). Where recovery was greater than or less than 100%, differences can likely be

attributed to procedural contamination or breakage of some particles during processing, resulting

in the situation that they may be counted as more than one particle or broken to a small enough

size that they pass through the sieve.

23

Boiling. During the boiling treatment, SB1 began to melt when the temperature reached 60 °C,

before boiling had initiated. After approximately two minutes, all SB1 beads were absent (0%

recovery) (Figure 1b). Melting and adhering to the stir bar was observed for SB2. The test was

repeated without the stir bar and no melting or adhering was observed. The boiling treatment for

SB3 resulted in complete loss in some replicates, with only a few spherical microbeads

recovered in others (~12%) (Figure 1b), but a small white, waxy mass was observed at the

bottom of the beaker in each replicate. For all other microplastic types, nearly 100% recovery

was observed (Figure 1b).

1N KOH and 4N KOH. Complete or nearly complete recovery (95-105%) was observed for all

microplastic types (Table 1). Minor discoloration of some microplastics, particularly SB1 and

SB2, was noted during examination using a microscope following 4N treatments.

Fe2SO4 + H2O2. For SB1, mean percent recovery was 0% (Figure 1e). Similar patterns were

observed for SB3 (5% recovery) (Figure 1e). For all other shapes, complete or nearly complete

recovery was observed (Figure 1e). After the first aliquot of H2O2 was added, the highest

recorded temperatures ranged from 72-89 °C, and typically increased for subsequent aliquots.

The peak temperature was recorded at 93 oC. The resulting lack of recovery for nearly all

treatments of SB1 and SB3 were consistent with other treatments where temperatures were

recorded at 60 °C or higher.

24

4N KOH + H2O2. For SB1, mean percent recovery of 3% was observed (Figure 1f). Loss

appeared to begin when temperatures were recorded between 63-68 °C, and complete loss

occurred rapidly if temperatures rose above 70 °C. Agglomeration of the beads occurred as

melting progressed. For all other microplastic types, complete or nearly complete recovery was

observed (Table 1, Figure 1f).

FT-IR Evaluation

Post-treatment FT-IR analysis of remaining microplastic particles confirmed that each digestion

method did not significantly impact the ability to confirm polymer type. For example, for SB3,

which was one of the materials subject to losses, each digestion type did not result in alterations

to the FT-IR spectra (Figure 2a). Where extraneous bands were observed in spectra, they were

attributed to residues from materials used (aluminum oxide from drying dishes, surfactant from

cleaning steps not sufficiently washed off).

25

Figure 2. FT-IR attenuated total reflectance spectra for (a) SB3 polyethylene microbeads subjected to different treatment conditions, and (b) SB1 (cera microcristallina) and irregularly shaped microbeads (IB) (polyethylene).

26

2.4 Discussion

Across the literature, microplastic extraction techniques vary and there have been several calls

for method standardization (GESAMP 2015). The chemical digestion method used to digest

organic matter will significantly impact the recovery of some types of microplastics. Two of

three types of spherical microbeads melted during treatments in this study and appear to be

considerably impacted by methods involving application of heat or reactions generating heat

above 60 °C. Our findings are consistent with previous studies noting melting and clumping of

polystyrene spheres using 22.5 M HNO3 at 100 oC (Claessens et al. 2013), and fusing of some

PET and high-density polyethylene (HDPE) under the similar conditions (Catarino et al. 2016).

Only ~4% recovery was observed following similar chemical digestion procedures with

polyethylene and polystyrene particles homogenized with 35 g of fish tissue (Avio et al. 2015).

While observing similar trends, these effects have been attributed to the strength of the acid

(Claessens et al. 2013; Avio et al. 2015; Catarino et al. 2016), and not to the heat applied or

created during the procedure. Nylon fibres and polystyrene were completely lost from samples

using strong acid digestion (HNO3) (Claessens et al. 2013; Catarino et al. 2016); however, nylon

fibres and polystyrene particles were not impacted by conditions used in our study. Loss is easily

identified in these types of studies where the initial number of microplastics entering the system

is known; however, loss in field studies will likely go undetected. Our results show that

application of, or uncontrolled exposure to, heat during the digestion process can lead to under-

representation of some microplastic particles and types of polymers in samples, and can lead to

an under-estimation of the total microplastics present in field-collected samples both in water

and in biological organisms where digestions are necessary to isolate microplastics. If it is

microbeads contained in personal care products that are more susceptible to such conditions, then

27

the relative contributions of this source of microplastics to the observed occurrence in the

environment and biota may be under-estimated.

Methods for processing marine and freshwater samples that use the wet peroxide oxidation, such

as the methods established by NOAA (Masura et al. 2015) using Fe(II) catalyst and wet peroxide

oxidation reaction, has demonstrated the potential to reach near-boiling temperatures. The

NOAA protocols note that a violent boil can occur at times, and recommends temperatures be

maintained at 75 °C and 95 °C respectively during different phases of processing (Masura et al.

2015; Mason et al. 2016b). Based on our results, these temperatures can result in the loss/melting

of some microbeads present in consumer products. Similarly, the use of wet peroxide oxidation

reaction following 4N KOH also uses or generates enough heat (S. Mason- SUNY Fredonia,

personal communication, Sept. 1, 2015) and we have shown the heat is sufficient enough to melt

some microbeads. Wet peroxide oxidation-based methods are quite effective in removing

organic material associated with aquatic organisms. Modifications of the wet peroxide oxidation

methods, or any other methods that include steps with elevated temperatures, to reduce

temperatures to 60 °C or below, both where heat is applied and where generated through the

oxidation reactions (e.g. using an ice bath), will ensure the protocols continue to be effective

when used to isolate microplastics from environmental matrices.

Based on our study, the 14-day 1N KOH and 4N KOH methods appear to be viable alternatives

to the NOAA protocol, both adapted originally from Foekema et al. (2013). Although it takes

longer for KOH to digest organic matter than wet peroxide oxidation methods, using 4N KOH

28

does not require the full 14 days to digest many types of organic matter. The KOH-based

methods are also less labor-intensive. Preliminary tests in our laboratory using plant-based

cellulosic material have demonstrated that 4N KOH is not caustic enough to break down this

material, but fish digestive tracts are digested with relative ease. The effectiveness of 1N/4N

KOH in our study is consistent with Dehaut et al. (2016), where a test to determine a benchmark

protocol for microplastic extraction and characterization of microplastics from tissues showed

that an adapted methodology based on Foekema et al. (2013) and Rochman et al. (2015) using

1N KOH incubated at 60 oC for 24 hours was most effective. Incubated 1N KOH appears to be

viable as the duration is much shorter; however, we caution against applying heat as 60 oC is

very close to the temperature we have shown to induce melting.

Plastic microbeads in this study (SB1, SB2, SB3, irregularly shaped microbeads, and fragments)

were isolated from store-bought personal care products that included polyethylene in the product

ingredient listing. The cleanser used to isolate SB1 also included ingredients consisting of

“microcrystalline wax” or “cera microcristallina”, or “synthetic wax”. These materials are also

used as thickeners, binders, and emulsifiers in cosmetics so it is not clear from ingredient listings

if these refer to beads or other forms. In this study, both SB1 and irregularly shaped microbeads

were obtained from the same product, with irregularly shaped microbeads being polyethylene

and SB1 likely being cera microcristallina. The FT-IR of both materials resemble polyethylene,

but with the SB1 spectra having differing crystallinity (Figure 2b; different intensities of the

paired bands at approximate wave numbers 720 and 1460). Microcrystalline waxes have a

melting point range of 62-102 oC (FAO 1995). This is consistent with the temperature range

within which SB1 melted during the evaluation of treatments; however, SB3 also melted and is

29

from a product with only polyethylene listed as the likely microbead composition. Varying

molecular weight ranges of polyethylene may be used to form beads, with lower molecular

weight polyethylene having a lower melting range. These materials may also be referred to as

“synthetic wax” or “polyethylene wax”. We subsequently conducted boiling tests of microbeads

from a wider range of personal care products, with 6 of 14 products containing beads that melted

at temperatures ranging from 70-98 oC. While it was generally the less numerous and colorful

spherical beads that melted, one product listed only polyethylene as the microbead ingredient. A

more in-depth analysis may be merited to better characterize the materials used as microbeads

and potential replacements. There is little information on the use of microcrystalline wax or

other waxes used as microbeads in personal care products, their stability in the environment and

when ingested in organisms, and whether they could have similar impacts to organisms as

polyethylene microbeads. Controlling temperature during sample processing will become even

more important should wax-based microbeads become alternatives to polyethylene and other

“plastic” microbeads and if tracking occurrence in the environment is warranted.

2.5 Conclusions

It is essential to test chemical digestion protocols for recovery of microplastics before proceeding

with sample processing and analysis. Given the results of our study, we caution against any

chemical digestion methods requiring heat or generating temperatures greater than 60 °C, or

specifying the addition of heat to similar or greater temperatures, in either the digestion or drying

stages. It is possible that other microplastics may exhibit fusion or melting at lower temperatures

than those used in this study. Based on varying recoveries depending on microplastic type and

method used, we recommend the use of KOH (1N or 4N) at room temperature, or incubated at

30

temperatures less than 60 °C for fish tissue digestions. Wet peroxide oxidation remains an

effective method for digesting samples with plant matter in marine and freshwater samples

(water, sediment), but temperatures must be controlled, eliminating spikes in temperature during

reactions, at or below 60 °C in order to minimize the loss of any constituent microplastics.

Assessments on the occurrence, types, sources and impacts of microplastics may be incomplete

if method-processing conditions selectively remove some types of materials.

31

Chapter 3

Microplastic Ingestion and Retention by Type in Three 3Species of Fish from the Great Lakes

3.1 Introduction

Microplastics are ubiquitous in marine and freshwater ecosystems. Different types of

microplastics may behave differently in the water column. Polymer composition makes some

types of microplastics, such as polyethylene microbeads, more buoyant and these may remain

floating within surface waters for longer periods of time relative to other types, such as polyester

fibres, which tend to sink (Ballent et al. 2016). Density is one of the major factors affecting

buoyancy. Polyethylene has a density ranging from 0.92-0.97 g per cm3 versus polyester, which

has a density of 1.2-2.3 g per cm3 (Hidalgo-Ruz et al. 2012). Differences in buoyancy related to

polymer type are likely why a higher concentration of fibrous particles are observed in

subsurface tows (Moore et al. 2001) and sediment samples (Corcoran et al. 2015; Ballent et al.

2016). As mentioned, there are a number of other factors that can affect buoyancy of

microplastics debris, including biofouling, entrapment in detritus and ingestion by aquatic

organisms if they are subsequently excreted in fecal pellets (Eriksen et al. 2014). In theory, the

position of different types of microplastics in the water column may determine which organisms

come in contact and subsequently ingest the microplastics.

Ingestion of microplastics has been observed to occur in hundreds of species of aquatic

organisms varying in size, position in the water column, feeding strategy and trophic level, from

zooplankton (Cole et al. 2013) to whales (Fossi et al. 2012; Fossi et al. 2014). Zooplankton from

32

the Northeast Atlantic coastal region exposed to polystyrene spheres (7.3-30.6 µm) showed the

capacity to ingest microplastics in 13 of 15 taxa tested, though some taxa showed varying

degrees of size-specific selectivity (Cole et al. 2013). Copepods and euphausids generated

currents using rapid movement of swimming legs or feeding appendages which allowed them to

draw in microplastics via filter-feeding; whereas doliolids siphoned microplastics into their body

cavity and entrapped them (Cole et al. 2013). Similarly, the dinoflagellates detected

microplastics using flagella and engulfed them (Cole et al. 2013), showing microplastic ingestion

despite having a different feeding strategy.

Ingestion has been observed in larger filter-feeding organisms as well. Bivalves and other filter-

feeding organisms living in shallower pools and intertidal zones have been observed to have

ingested and retained microplastics in the environment (Browne et al. 2008; Van Cauwenberghe

& Janssen 2014; Sussarellu et al. 2016). Commercially farmed mussels (Mytilus edulis) and

oysters (Crassostrea gigas) had ingested microplastics prior to harvesting, with average loads of

0.36 ± 0.7 and 0.47 ± 0.2 microplastics per gram of soft tissue respectively (Van Cauwenberghe

& Janssen 2014). Microplastic retention was observed in bivalves ranging from 12 hours

(Browne et al. 2008) to three days (Van Cauwenberghe & Janssen 2014). Polystyrene spheres

were observed in the gut cavity and digestive tubules of mussels (Browne et al. 2008), and the

stomach and intestine of oysters (Sussarellu et al. 2016). Some particles were also observed to

have translocated into the circulatory fluid of mussels (Browne et al. 2008). Oysters showed

some evidence of size-based selectivity for 6 µm microplastics (69 ± 6%) over 2 µm

microplastics (14 ± 2%) (Sussarellu et al. 2016). Very large filter-feeding organisms are assumed

to have ingested microplastics as well. Mediterranean Fin Whales (Balaenoptera physalus) and

33

Mediterranean Basking Sharks (Cetorhinus maximus) contained relatively high concentrations of

phthalates, a class of chemicals added to plastics, in their blubber and muscle tissue respectively

(Fossi et al. 2012; Fossi et al. 2014). In these studies, the detection of phthalates in tissues is

used as an indicator of microplastic ingestion (Fossi et al. 2012; Fossi et al. 2014).

Ingestion of microplastics by dozens of species of marine fish has been observed in the

environment at a number of locations (Boerger et al. 2010; Fossi et al. 2012; Foekema et al.

2013; Lusher et al. 2013; Fossi et al. 2014; Avio et al. 2015; Romeo et al. 2015). The ranges in

percentages of fish digestive tracts containing microplastics vary among studies. Only

approximately 3% of fish from the North Sea contained microplastics (Foekema et al. 2013).

Foekema et al. (2013) found textile fibres (~1 mm) in nearly every sample, though these were

excluded from analysis under the assumption that they may have been the result of airborne

contamination. Approximately 37% of fish were found to contain microplastics in the English

Channel (Lusher et al. 2013). In the Mediterranean Sea, 13-32% of fish contained microplastics,

with some species having much higher levels than others (Romeo et al. 2015). In the Adriatic

Sea, 19-67% of fish contained microplastics (Avio et al. 2015). In the NPCG, 35% of fish

collected had microplastics in their stomachs (Boerger et al. 2010). The average number of

microplastics per fish also varies. The overall average number of microplastics in the digestive

tracts of fish from the English Channel was 1.9 ± 0.1 pieces per fish; however, the range went

from 1-15 pieces per fish (Lusher et al. 2013). Fish from the NPCG had an average of 21

microplastics per stomach (Boerger et al. 2010). An average of 1.8 ± 0.7 microplastics was

found in the digestive tracts of European Pilchard (Sardina pilchardus) in the Adriatic Sea (Avio

34

et al. 2015). Clearly, a great deal of variation exists, and some of the variation may be attributed

to differences among species.

Microplastic ingestion has been observed in a range of trophic levels for fish species. Lower

trophic-level fish, including Golden Lanternfish (Myctophum aurolaternatum) and Bigfin

Lanternfish (Symbolophorus californiensis), were observed with microplastics in their digestive

tracts (Boerger et al. 2010). Large, pelagic predators including Swordfish (Xiphias gladius),

Bluefin Tuna (Thunnus thynnus) and Albacore (Thunnus alalunga) were found to have ingested

fragments in 13%, 32.% and 13% of individuals respectively (Romeo et al. 2015). Microplastics

also appear to be present in fish digestive tracts at all positions in the water column. Studies have

attempted to compare the abundance of microplastics in fish digestive tracts between pelagic and

demersal species (Foekema et al. 2013; Lusher et al. 2013; Avio et al. 2015). No significant

difference was found in the abundance of microplastics in fish digestive tracts of five pelagic and

five demersal species in the English Channel (Lusher et al. 2013). Both Blue Whiting

(Micromesistius poutassou) and Red Gurnard (Aspitrigla cuculus) were found to have

microplastics in the digestive tracts of over 50% of collected individuals, and are pelagic and

demersal species respectively (Lusher et al. 2013). Seven pelagic and demersal species were

collected from the North Sea and microplastics were detected in all but two species, with the

largest percentage found in Atlantic Cod (Gadus morhua), a demersal species (13%) (Foekema

et al. 2013). Of five species of fish collected in the Adriatic Sea, the pelagic species showed the

lowest percentage of fish containing plastics (19%) (Avio et al. 2015). Red Mullet (Mullus

barbatus) and Tub Gurnard (Chelidonichthys lucernus), two benthic species, showed the highest

percentage of fish containing microplastics (64% and 67%); however, the average number of

35

microplastics per fish showed the opposite relationship (Avio et al. 2015). The highest mean

number of microplastics per fish was found in European Pilchard (Sardina pilchardus), a pelagic

species (Avio et al. 2015). The significance of trophic level and position in the water column in

relation to abundance of microplastics in fish digestive tracts is unclear based on current

literature.

The types of microplastics found in fish digestive tracts vary among species. Epipelagic and

mesopelagic species of fish, generally occupying the upper levels of the water column, are more

likely to come in contact with more buoyant plastic debris. Less buoyant debris that sinks below

the surface of the water column, or sinks to combine with sediments may prove more

problematic for benthopelagic and demersal species of fish, generally occupying intermediate

and low depths respectively. One study determined that approximately 68% of the microplastics

in fish digestive tracts were fibres (Lusher et al. 2013), whereas others have found the

predominant microplastic type to be fragments (Boerger et al. 2010; Avio et al. 2015; Romeo et

al. 2015). It has been suggested that fibres are abundant in sediments, perhaps related to their

composition of low-density polymers relative to other microplastics (Thompson et al. 2004;

Ballent et al. 2016), and are abundant in wastewater effluent (Mason et al. 2016a; Mason et al.

2016b; Sutton et al. 2016). A higher abundance of fibres in fish may indicate a higher

contribution of wastewater effluent in a given region, or it may be related to the feeding

strategies of the fish sampled if they are demersal species frequently foraging in sediments

containing less buoyant microplastics. Alternatively, an abundance of fragments may indicate a

higher contribution of secondary microplastics in the environment. Fragments in fish digestive

tracts may also be related to feeding strategies as more buoyant fragments are often dominant in

36

surface waters that pelagic fish forage in (Moore et al. 2001; Eriksen et al. 2014; Yonkos et al.

2014; Mason et al. 2016a). The significance of feeding strategy and position in the water column

in relation to the proportions of different types of microplastics in fish digestive tracts is unclear.

Clearly, a great deal of variation exists from study to study around whether position in the water

column, trophic level and feeding strategies actually affect the amount of microplastics or types

of microplastics identified in digestive tracts of marine fish caught in the wild. A great deal of

variation also exists in the methods used to isolate microplastics (as mentioned in Chapter 2), and

the abundance of microplastics in the environment at these locations. Laboratory-based

exposures are conducted to reduce some of this variation. An early attempt at laboratory-based

microplastics exposure in fish subjected six species of larval teleost fish to polystyrene

microspheres (100-500 µm), and only two of six species (Spot (Leiostomus xanthurus) and

Striped Mullet (Mugil cephalus)) were observed with polystyrene microplastics in their gut

(Hoss & Settle 1990), both being demersal species. Of six species tested, five ingested plastics

after 48 hours of food deprivation, and some particles were retained in the gut; however, no

blockage or mortality was observed (Hoss & Settle 1990). The species observed to have ingested

plastics (Atlantic Menhaden (Brevoortia tyrannus), Pinfish (Lagodon rhomboides), Spot, Striped

Mullet and two species of Flounder (Paralichthys spp.)) were an array of pelagic, bentho-pelagic

and demersal species (Hoss & Settle 1990). On the other hand, demersal Common Goby

(Pomatoschistus microps) collected from the field were exposed to polyethylene microspheres

(420-500 µm) either alone or in combination with brine shrimp prey (Artemia franviscana

nauplii) and were found to ingest the polyethylene microspheres in all treatments (de Sa et al.

2014). Eurasian Perch (Perca fluviatilis) juveniles (10-day old larvae) exposed to polystyrene

37

particles (90 µm) at environmentally relevant concentrations for pelagic zones (10,000 particles

per m3 and 80,000 particles per m3) in combination with brine shrimp prey, displayed

significantly altered feeding behaviour (Lönnstedt & Eklöv 2016). Larvae from the high

microplastics treatment consumed an average of 7.2 ± 1.2 polystyrene particles over 24 hours

and their stomachs contained only microplastics and none of the food source provided

(Lönnstedt & Eklöv 2016). Fish fed on polystyrene microplastics, and not on a traditional food

source (Lönnstedt & Eklöv 2016). Again, variation exists from study to study; however, in every

study fish were observed consuming microplastics.

Several species of fish have been observed ingesting microplastics in both field and lab settings.

Relatively few studies have identified significant differences in microplastic ingestion between

species. One suggests there may be a difference between pelagic and demersal species (Avio et

al. 2015), while another study appears to contradict these results (Lusher et al. 2013).

Microplastics are abundant both in surface water and in sediment, particularly within close

proximity to urbanized and industrialized areas (Eriksen et al. 2013; Corcoran et al. 2015;

Ballent et al. 2016; Sutton et al. 2016). Various sources of microplastics typically contribute

different types of microplastics to surrounding aquatic environments (Castaneda et al. 2014;

Ballent et al. 2016). The variation in sizes of microplastics may have differences in ultimate fate

upon ingestion (Cole et al. 2013; Sussarellu et al. 2016). Since size-based selectivity has been

demonstrated (Cole et al. 2013; Sussarellu et al. 2016), it is possible that selectivity may exist for

microplastic type as well. It may be more meaningful to assess whether a feeding preference

exists using fish that are more representative of those present and feeding in the natural

environment to investigate differences in short-term retention among ingested microplastic types.

38

In this laboratory-based experiment, fish were exposed to a range of microplastic types from

various consumer and industrial products to determine whether microplastic type affects the

potential for retention of microplastics in the digestive tracts following short-term exposure. This

thesis considered three fish species that inhabit the Great Lakes and other lakes in this region,

which have been shown to have relatively high concentrations of microplastics in surface waters

and sediments (Eriksen et al. 2013; Corcoran et al. 2015; Ballent et al. 2016; Mason et al.

2016b). These species were selected as they use different positions in the water column and

feeding strategies. White Sucker (Catostomus commersonii) are a benthic species that

preferentially feed on detritus, aquatic insects, fish eggs, small crustaceans, molluscs and other

invertebrates (Holm & Mandrak 2009). White Sucker in Ontario were found to have

cladocerans, or water fleas, as 60-90% of their gut contents (Scott & Crossman 1973). Fathead

Minnow (Pimephales promelas) are a benthopelagic species that feed on algae, detritus, aquatic

insect larvae, small crustaceans and zooplankton (Scott & Crossman 1973; Laurich et al. 2003).

Fathead Minnow have been described as opportunistic filter feeding forage fish that sift through

mud and silt to find prey items (Laurich et al. 2003). Rainbow Trout (Oncorhynchus mykiss)

will feed anywhere from the lake bottom to the surface (Scott & Crossman 1973). Rainbow

Trout are opportunistic predators, as they will eat a number of things, including small fish,

crustaceans, snails, aquatic insects, insect larvae, fish eggs and leeches (Scott & Crossman 1973;

Holm & Mandrak 2009).

39

3.2 Materials and Methods

Microplastic Particles

Spherical polyethylene microbeads, irregular-shaped polyethylene microbeads and polyethylene

fragments were isolated from different rinse-off personal care products readily available in

Ontario, Canada, by rinsing the product through a 125 µm metal sieve using deionized water.

Shavings were mechanically generated from a polyethylene block using a pipe-threader in a drill

press. Foam spheres were taken from a sheet of polystyrene foam insulation board. Synthetic

fibres were cut from a nylon carpet sample. Twenty particles were used in individual replicates

for the feeding experiments. Microplastics within the 125-1000 µm range were selected for use

in the feeding experiments.

Food-Microplastic Preparation

For the Rainbow Trout feeding trial, 2 mm food pellets prepared by the toxicology group at the

Ontario Ministry of the Environment and Climate Change (MOECC) were placed in glass petri

dishes and soaked in deionized water for 24 hours. After 24 hours, microplastics were embedded

in the soaked pellets, with one microplastic particle per pellet. Pellets were placed in aluminum

trays and dried in a drying oven at approximately 55 °C to harden for 30 minutes. Once dried,

pellets were coated in oil from wild salmon oil capsules prior to feeding. For White Sucker and

Fathead Minnow treatments, microplastics for each replicate were placed in individual plastic

petri dishes. Approximately 15 mL of brine shrimp slurry, prepared by the toxicology group at

the MOECC, was added to each petri dish. Lids were placed on each petri dish and the dishes

were placed in a freezer until feeding, resulting in a microplastic-containing frozen disc. Six

40

treatments were assigned, one for each microplastic type. Fish were fed in this manner to allow

them to feed more naturally as opposed to force-feeding. A control treatment of clean,

microplastic-free food was included as the seventh treatment to assess levels of airborne

contamination due to experimental design processing.

Water Quality Protocol

City of Toronto municipal water, dechlorinated in the laboratory to make laboratory dilution

water, was used for the feeding experiments. This is the same water used for culturing of fish.

Water was aerated using glass filters at approximately 20 mL per minute per gram of fish. The

pH (7.8-8.0) and conductivity (320-338 µS/cm) of the laboratory dilution water was monitored

throughout feeding experiments.

Rainbow Trout 24-hour Exposure

Experimental conditions. Prior to acclimation, 10 L food-grade polyethylene bioassay pails were

filled with lab dilution water. Translucent lids covered the buckets to prevent any airborne

contamination of the buckets. Natural daylight visible through the laboratory windows was used.

Temperature in the laboratory was maintained at 15 ± 1 °C. Treatment pails were randomly

assigned within the laboratory to ensure that conditions did not vary among treatments. A total of

seven treatments were used, including the control treatment. Eight replicates were used for each

of the experimental treatments, and three replicates were used for the control treatment. Fewer

replicates were used for the control treatment as it was included to verify that microplastics of

41

the types and character that were added through the food were also not present as the result of

contamination through the feeding, incubation, and processing procedures.

Experimental procedure. One juvenile Rainbow Trout (9-26 g) was placed in each pail for an

acclimation period of 72 hours. Fish were not fed during acclimation. After 72 hours, 20 food

pellets, each containing one microplastic particle of a single type, were added to each pail. Fish

were allowed three hours to feed. After the feeding period, fish were removed and transferred to

a clean bucket containing fresh laboratory dilution water. Fish were allowed 21 hours to pass any

food and microplastics they had consumed (24 hours following initial exposure to microplastics).

After 24 hours, fish were euthanized. The contents of the initial acclimation/feeding pail and the

contents of the second pail were run through a 125 µm metal sieve separately. The sieved

contents were stored in 80% ethanol in plastic vials. The contents of the initial

acclimation/feeding pail were referred to as W1 and the contents of the second pail were referred

to as W2.

Dissection. Following euthanization, the total length of each fish was recorded in centimetres,

and the wet weight of the fish was recorded in grams. The digestive tract was extracted by

cutting along the ventral side of the fish from the gills to the anal pore. The digestive tract was

extracted from the top of the esophagus to the anal pore. The wet weight of the digestive tract

was recorded in grams. Any immediately visible microplastics or blockages were photographed

using an Apple iPhone 5S camera. Three fish from each treatment were selected at random for

examination using a microscope to determine whether any microplastics were visible through the

42

walls of the digestive tract. For the selected fish, the contents of the digestive tract were

massaged out from the esophagus downwards and into a plastic vial. Any feces excreted during

euthanization and dissection were collected and placed in the same plastic vial. The contents of

the digestive tract were stored in 80% ethanol. The emptied digestive tracts were placed in

plastic petri dishes and stored on ice until examination, when they were prepared for storage in

80% ethanol. For the remaining five fish, digestive tracts were placed directly into plastic vials

and stored in 80% ethanol.

Microscope examination. The three randomly selected digestive tracts from each treatment were

examined using a dissecting microscope (Leica S8 APO Stereozoom; Leica Microsystems

Canada, Inc., Richmond Hill, ON, Canada) at 10-80X magnification. Visible microplastics were

noted and photographed using the microscope.

Digestion. Each of the W1, W2 and digestive tract and digestive tract content samples were

rinsed through a 125 µm metal sieve using deionized water. Contents of the sieve were rinsed

into plastic vials. The room temperature 4N KOH method described in Chapter 2, originally

adapted from Foekema et al. (2014), was used to digest samples, with the volume of the 4N

KOH solution being approximately three times the volume of the sample in each vial. After 14

days, the contents of the vials were rinsed through a 125 µm metal sieve using deionized water,

captured on a 10 µm polycarbonate filter (Fisher Scientific, Ottawa, ON, Canada), and counted

under the dissecting microscope at 10-80X magnification.

43

White Sucker and Fathead Minnow 48-hour Exposures

Juvenile White Sucker (7-15 g) and adult Fathead Minnow (1-6 g) were used in 48-hour

exposures. White Sucker treatments were maintained at an ambient temperature of 15 ± 1 °C and

Fathead Minnow treatments were maintained at an ambient temperature of 23 ± 2 °C. White

Sucker and Fathead Minnow 48-hour exposures followed the same experimental design as the

Rainbow Trout 24-hour exposure, with only minor modifications. Instead of 10 L food-grade

polyethylene pails, 4 L glass jars were used. After acclimation for 72 hours, fish were fed by

adding the frozen brine shrimp slurry discs combined with 20 microplastics of a single type to

each jar. After three hours allotted for feeding, fish were transferred to a clean jar for 21 hours to

allow digestion and egestion of material. After 21 hours (24 hours post initial microplastic

exposure), fish were transferred a second time to a third clean jar. Fish were fed again, but with

15 mL of clean, microplastic-free brine shrimp slurry frozen discs. The experimental design was

amended to include this second feeding, which may aid in moving any microplastics remaining

after 24 hours through the digestive tract. This is more environmentally relevant. Twenty-four

hours after the second feeding, and 48 hours after the initial microplastic exposure, fish were

euthanized. The contents of this third jar were referred to as W3. Three randomly selected

replicates from each treatment were selected for microscope examination, and any microplastics

visible through the walls of the digestive tracts were noted and photographed under the

microscope. The digestive tracts of all fish were stored in 80% ethanol for digestion in 4N KOH

at a later date using the same extraction methods as described above.

44

Microplastic Counts and Data Analysis

The number of microplastics in each sample was counted using a dissecting scope. For the three

randomly selected replicates that underwent examination using a microscope prior to digestion

from each treatment, digestive tract and digestive tract contents samples were combined for

analysis (herein referred to as digestive tract samples). In each treatment, W1 was excluded from

statistical analysis, as it was not of interest in determining the fate of microplastics upon

ingestion. Microplastic counts from the W1 samples were only of interest in determining the

number of microplastics that were not ingested by the fish during the three-hour feeding period

to ensure no major loss of microplastics had occurred. The control treatment was also excluded

as no microplastics were found in the controls that matched the types added to each treatment.

Any microplastics that were observed in the control treatments were noted. Since the

microplastic type in each treatment was known, any microplastics in the control treatments could

be attributed to contamination, as they were visibly different in appearance and composition

from the microplastics used in the experimental treatments and only fibres were observed in

control treatments. Eight independent replicates were used in each feeding experiment. The mean

number of microplastics and standard deviation in each phase (W2, W3 and digestive tract) were

calculated across all eight replicates from each treatment. Any microplastics that were visibly

different from the microplastics put into the experimental treatments were attributed to airborne

or processing contamination, and were not included in the counts for each treatment. A G-test,

similar to a chi-square test, was used for each species to determine whether microplastic counts

differ among treatments, with treatments representing different microplastic types (Woolf 1957).

The G-Test is more accurate than a chi-square test when counts are small (Woolf 1957). No post-

hoc analysis was performed on the Rainbow Trout data as there was no appropriate post-hoc

statistical test for this study design. For the 24-hour Rainbow Trout exposure, the microplastic

45

counts from two phases (W2 and digestive tract) were compared using a G-test. For the White

Sucker and Fathead Minnow exposures, three phases (W2, W3 and digestive tract) were

compared using a G-test. The G-tests were performed in R using the “GTest” function to

determine whether there were common responses or whether there is evidence of statistically

significant differences (α= 0.05) among groups of microplastic types for each feeding

experiment.

3.3 Results

Control Treatments

Minimal contamination was observed in control treatments for all three species. No

contamination was observed in Rainbow Trout. Higher levels of contamination were observed in

White Sucker, ranging from zero to four particles observed in individual replicates. The highest

mean number of particles observed in the White Sucker control treatment was in W3 (3.0 ± 1.0

particles) and the lowest mean number of particles was in the digestive tracts (0.5 ± 0.5

particles). Contamination could be associated with a number of factors. For example, the

laboratory in which the experiment was conducted had a greater frequency of staff pass through

than for the Rainbow Trout experiment. The lids on the glass jars were not a tight fit, allowing

greater opportunity for airborne contamination. There were also differences in assistants helping

with processing on different days that could contribute different fiber material from clothing. The

Fathead Minnow experiment was run in the same setting at a later date, using the same glass jars.

Contamination was lower in the Fathead Minnow experiment, ranging from zero to one particle

in individual replicates. The mean number of contamination particles was consistent among

phases for the Fathead Minnow treatment (0.7 ± 0.6 particles). All particles observed as

46

contamination in control treatments were fibres, and were typically opaque and black in colour

and resembling clothing fibres or blue in colour resembling net fibres used to transport fish in

and out of the treatment pails and jars. The fibres observed were visibly different in composition

as the synthetic fibres used in the feeding experiments were brightly coloured, translucent and

wiry. Fibre contamination was visibly distinguished from the fibres used in experimental

treatments with ease.

Rainbow Trout 24-hour Exposure

The differences among number of microplastic treatments observed to be ingested and retained

after the initial feeding period among treatments were significant (G-Score= 64.4, 5 d.f., α= 0.05,

p <0.0001) (Figure 3). For half of the microplastics tested, including spherical microbeads,

fragments and polystyrene foam beads, a higher mean number of microplastics was found in the

digestive tract than in W2, indicating more microplastics were retained than excreted (Figure 4a,

4c, 4e). For irregularly shaped microbeads, shavings and synthetic fibres, the opposite was true

(Figure 4b, 4d, 4f). More microplastics were present in the W2 than in the digestive tract. The

highest mean number of microplastics retained in the digestive tract was for polystyrene foam

beads, followed by spherical microbeads (Figure 3). For polystyrene foam beads, five of 8 fish

ingested microplastics, and up to 14 polystyrene foam beads were retained in one fish (Figure 5).

Of all the polystyrene foam beads ingested (22 particles), only one single polystyrene foam bead

was excreted within 24 hours. In the fragments treatment, 7 of 8 fish ingested fragments. Three

of 7 fish retained all of the fragments they ingested over 24 hours, with up to 8 fragments being

retained by a single fish. At least some microplastics were observed in digestive tracts for every

microplastic type after 24 hours (Table 2). The highest number of ingested microplastic particles

47

was for synthetic fibres (82 particles), and 22% of the ingested synthetic fibres were retained (18

particles) (Table 2). While polystyrene foam beads had the second lowest number of ingested

microplastic particles (22 particles), it was also highest for number of particles retained in the

digestive tract (21 particles, 95%) (Table 2). Spherical microbeads and fragments also had

relatively high percentages of microplastics retained in Rainbow Trout digestive tracts (77% and

67% respectively) (Table 2). Overall, approximately 44% of ingested microplastic particles were

retained over 24 hours (Table 2).

Figure 3. A summary of the mean number of microplastics of each type observed in water 24 hours after microplastic exposure (W2) and in the digestive tracts (G+GC) of Rainbow Trout (n=8). Fish were exposed to spherical microbeads (SB), irregular microbeads (IB), fragments (F), shavings (S), polystyrene foam beads (PSF) and synthetic fibres (SF).

48

Figure 4. The mean number of microplastics of each type observed in water 24 hours after microplastic exposure (W2) and in the digestive tracts (G+GC) of Rainbow Trout (n=8). Fish were exposed to spherical microbeads (SB), irregular microbeads (IB), fragments (F), shavings (S), polystyrene foam beads (PSF) and synthetic fibres (SF).

49

Figure 5. Polystyrene foam beads accumulated near the top of the esophagus and stomach region of one Rainbow Trout.

Table 2. Sum of spherical microbeads (SB), irregularly shaped microbeads (IB), fragments (F), shavings, (S), polystyrene foam beads (PSF) and synthetic fibres (SF) across all replicates for Rainbow Trout (n=8). The number of microplastics excreted within 24 hours (W2) and the number of microplastics in the digestive tract (G+GC) are also summed. The percentage of microplastics ingested (% ingested) out of the total number of microplastics the fish were exposed to for each treatment (160 particles) is included, as well as the percentage of microplastics retained in fish digestive tracts (% retained) out of the total number of microplastics fish ingested for each treatment.

SB IB F S PSF SF Total

W2 6 26 7 17 1 64 121

G+GC 20 12 14 9 21 18 94

Total 26 38 21 26 22 82 215

% Ingested

16 24 13 16 14 51 22

% Retained

77 32 67 35 95 22 44

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White Sucker 48-hour Exposure

For White Sucker, irregularly shaped microbeads, polystyrene foam beads and synthetic fibres

had decreasing numbers of microplastics observed in samples with increasing time following

exposure (Figure 6b, 6e, 6f). In each treatment, the majority of microplastics were passed within

24 hours, and most of the remaining microplastics were passed by the 48-hour mark; however,

some still remained in the digestive tract after 48 hours. The opposite was true for spherical

microbeads, fragments and shavings, where a higher mean number of microplastics were

observed in digestive tracts than W3 (Figure 6a, 6c, 6d), suggesting more microplastics were

retained than excreted between 24 and 48 hours. Microplastics in the digestive tract after 48

hours indicates short-term retention, though the differences among microplastic types were not

significant (G-Score= 12.0, 10 d.f., α= 0.05, p >0.2) (Figure 7). Excluding polystyrene foam

beads, at least some microplastics were observed in digestive tracts for every microplastic type

after 48 hours. Across all treatments, the highest numbers of ingested microplastics were for

irregularly shaped microbeads (55 particles), with 16% of the ingested irregularly shaped

microbeads being retained in the digestive tract after 48 hours (9 particles) (Table 3). Overall,

approximately 17% of all ingested microplastic particles were retained over 48 hours (Table 3).

Some microplastics were observed in the control treatments. The highest level of contamination

was observed in the W3 phase, with a mean of 3.0 ± 1.0 particles. The lowest levels of

contamination were observed in the digestive tracts, with a mean of 0.5 ± 0.5 particles. These

few particles were visually identified as fibres, and likely originated from the nets used to

transfer fish into and out of the experimental jars.

51

Figure 6. The mean number of microplastics of each type observed in water 24 hours after microplastic exposure (W2), 24 hours after feeding with clean food and 48 hours after microplastic exposure (W3) and in the digestive tracts (G+GC) of White Sucker (n=8). Fish were exposed to spherical microbeads, irregular microbeads, fragments, shavings, polystyrene foam beads and synthetic fibres.

52

Figure 7. A summary of the mean number of microplastics of each type observed in water 24 hours after microplastic exposure (W2), 24 hours after feeding with clean food and 48 hours after microplastic exposure (W3) and in the digestive tracts (G+GC) of White Sucker (n=8). Fish were exposed to spherical microbeads (SB), irregular microbeads (IB), fragments (F), shavings (S), polystyrene foam beads (PSF) and synthetic fibres (SF).

53

Table 3. Sum of spherical microbeads (SB), irregularly shaped microbeads (IB), fragments (F), shavings, (S), polystyrene foam beads (PSF) and synthetic fibres (SF) across all replicates for White Sucker (n=8). The number of microplastics excreted within 24 hours (W2), the number of microplastics excreted within 24 hours of a subsequent feeding of microplastic-free food, and the number of microplastics in the digestive tract (G+GC) are also summed. The percentage of microplastics ingested (% ingested) out of the total number of microplastics the fish were exposed to for each treatment (160 particles) is included, as well as the percentage of microplastics retained in fish digestive tracts (% retained) out of the total number of microplastics fish ingested for each treatment.

SB IB F S PSF SF Total

W2 14 37 12 13 2 12 90

W3 1 9 0 2 1 5 18

G+GC 2 9 1 5 0 5 22

Total 17 55 13 20 3 22 130

% Ingested

11 34 8 13 2 14 14

% Retained

12 16 8 25 0 23 17

Fathead Minnow 48-hour Exposure

Fathead Minnow had fewer mean microplastics observed with increasing time following

exposure for every treatment (Figure 8). This indicated that Fathead Minnow were able to pass

the majority of microplastics within the first 24 hours post exposure, though some still remain in

the digestive tract (Figure 9a-f). Differences among groups were not significant (G-Score= 13.9,

10 d.f., α= 0.05, p >0.1). The microplastic types retained in digestive tracts most abundantly

were synthetic fibres, followed by shavings and irregularly shaped microbeads (Figure 9). No

spherical microbeads were retained in digestive tracts in any of the 8 individual fish. With the

exception of spherical microbeads, some microplastics were observed in digestive tracts for

54

every microplastic type after 48 hours. The most abundant ingested microplastic particle was

synthetic fibres (40 particles), and 18% of synthetic fibres were retained after 48 hours (7

particles) (Table 4). While only four polystyrene foam beads were ingested, 25% were retained

(1 particle) (Table 4). Approximately 20% of ingested fragments were retained, and 19% of

ingested shavings were retained (Table 4). Overall, approximately 14% of ingested microplastic

particles were retained over 48 hours (Table 4). Contamination was observed in control

treatments, with a consistent mean of 0.7 ± 0.6 particles found in each phase.

Figure 8. A summary of the mean number of microplastics of each type observed in water 24 hours after microplastic exposure (W2), 24 hours after feeding with clean food and 48 hours after microplastic exposure (W3) and in the digestive tracts (G+GC) of Fathead Minnow (n=8). Fish were exposed to spherical microbeads (SB), irregular microbeads (IB), fragments (F), shavings (S), polystyrene foam beads (PSF) and synthetic fibres (SF).

55

Figure 9. The mean number of microplastic of each type observed in water 24 hours after microplastic exposure (W2), 24 hours after feeding with clean food and 48 hours after microplastic exposure (W3) and in the digestive tracts (G+GC) of Fathead Minnow (n=8). Fish were exposed to spherical microbeads, irregular microbeads, fragments, shavings, polystyrene foam beads and synthetic fibres.

56

Table 4. Sum of spherical microbeads (SB), irregularly shaped microbeads (IB), fragments (F), shavings (S), polystyrene foam beads (PSF) and synthetic fibres (SF) across all replicates for Fathead Minnow (n=8). The number of microplastics excreted within 24 hours (W2), the number of microplastics excreted within 24 hours of a subsequent feeding of microplastic-free food, and the number of microplastics in the digestive tract (G+GC) are also summed. The percentage of microplastics ingested (% ingested) out of the total number of microplastics the fish were exposed to for each treatment (160 particles) is included, as well as the percentage of microplastics retained in fish digestive tracts (% retained) out of the total number of microplastics fish ingested for each treatment.

SB IB F S PSF SF Total

W2 14 15 6 19 2 27 83

W3 7 12 2 10 1 6 38

G+GC 0 3 2 7 1 7 20

Total 21 30 10 36 4 40 141

% Ingested

13 19 6 23 3 25 15

% Retained

0 10 20 19 25 18 14

3.4 Discussion

These short-term feeding tests demonstrate that retention time in the digestive tract differs

among microplastic types and across different types of fish. For Rainbow Trout, the differences

among microplastic types were significant (G-Score= 64.4, 5 d.f., α= 0.05, p <0.0001). It is

important to note that a great deal of variation existed among replicates for most treatments. For

example, the mean number of polystyrene foam beads retained after 24 hours in Rainbow Trout

was 1.9 ± 4.1 microplastic particles, and up to 14 of the initial 20 microplastic particles were

retained in individual fish In the case of polystyrene foam beads, it was observed that ingested

microplastics were typically concentrated near the top of the esophagus and stomach region, and

57

an abundance of food particles were observed in the mouth region. The ingested polystyrene

foam beads did not move through the digestive tract easily, and appeared to cause a blockage

during the 24-hour time period (Figure 10). The fish that consumed 14 polystyrene foam beads

did not excrete any polystyrene foam beads during the 24-hour period. A total of 22 polystyrene

foam beads were ingested across the entire treatment, and only one was excreted. Over 95% of

the ingested polystyrene foam beads were retained in the digestive tract after 24 hours. While

this is an atypical example and the mean number of polystyrene foam beads retained across the

entire treatment is lower, it suggests that there is potential for individual fish to ingest larger

amounts of microplastics and apparently experience a blockage preventing the passage of

microplastics and further ingestion of food. Based on these observations, the ingestion of

polystyrene foam beads may be problematic for juvenile Rainbow Trout. Further studies are

necessary to examine this phenomenon in greater detail. One study has examined differences in

retention of microbeads and fibres in the digestive tracts of Goldfish (Carassius auratus), and

determined that no significant difference exists (Grigorakis et al. 2017). The significant

difference among microplastic types in the digestive tracts of juvenile Rainbow Trout is a novel

finding.

58

Figure 10. Polystyrene foam beads accumulated at the top of the esophagus caused apparent blockage of the digestive tract in one Rainbow Trout.

Hoss & Settle (1990) tested six species of fish for ingested plastics and determined that five of

six species would ingest microplastics after being deprived of food for 48 hours; however, fish

were also observed ingesting brine shrimp following microplastic exposure and no blockages or

mortality was observed despite observing microplastics in the digestive tracts of four species.

Goldfish retained approximately 0-3 of 50 microplastics six days after initial exposure, and

authors determined the potential for retention in the digestive tracts of Goldfish was low

(Grigorakis et al. 2017). Microplastic counts in Goldfish are similar to the counts observed in

this feeding study for all species, though no blockages were observed for Goldfish (Grigorakis et

al. 2017). This may be related to body size of the Goldfish used (24.8-27.1 g), which is generally

larger than the largest species of fish in this study (Rainbow Trout, 15.3 ± 5.0 g), and the particle

size Goldfish were exposed to which was as small as 50 µm (Grigorakis et al. 2017). Because the

59

body sizes of the Goldfish were larger, and the exposed particles were smaller, the likelihood of

observing a blockage is reduced. Other studies have shown ingestion of polystyrene

microplastics in particular but have not noted obvious blockages (Browne et al. 2008; Cole et al.

2013; Van Cauwenberghe & Janssen 2014; Lönnstedt & Eklöv 2016; Sussarellu et al. 2016).

Also, the studies mentioned are for a much smaller size class of polystyrene microplastics, the

particles were not polystyrene foam, and most are not using fish as a study organism (Browne et

al. 2008; Cole et al. 2013; Van Cauwenberghe & Janssen 2014; Sussarellu et al. 2016). This

study notes a blockage in the digestive tract of one Rainbow Trout as a result of polystyrene

foam beads microplastics ingestion. It is possible that the blockage observed in Rainbow Trout is

related to particle size as well as type because polystyrene foam beads tended to be on the larger

end of the 125-1000 µm size range.

In many individual Rainbow Trout, microplastic particles were retained, but with no obvious

blockage at the top of the esophagus. For example, 6 of 8 fish ingested spherical microbeads, and

only one fish excreted more particles than were retained after 24 hours. Some spherical

microbeads were trapped in the mouth and gills, requiring forceps to extract them. The particles

did not appear to be obstructing the top of the esophagus in the same manner as polystyrene

foam beads. This is not the first observation of this kind. Microplastics were observed adhering

to external appendages (feeding appendages, swimming legs, antennae, furca and filamental

hairs between carapace), and coating the carapace of molting or deceased copepods (Cole et al.

2013). Approximately 77% of the ingested spherical microbeads were retained across the entire

treatment, suggesting that the potential for spherical microbead retention is also high for juvenile

Rainbow Trout (Table 2).

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Similar observations were made for fragments as approximately 67% of ingested fragments were

retained across all 8 individuals (Table 2). For the remaining microplastic types (irregularly

shaped microbeads, shavings and synthetic fibres), microplastic particles were excreted more

frequently than they were retained, but there were still individuals where the number of

microplastic particles retained was higher than the number of microplastic particles excreted.

Overall, some microplastics were retained in the digestive tract in juvenile Rainbow Trout, and

the type of microplastics ingested likely impacts potential for retention in the digestive tract after

24 hours. Goldfish took approximately 10 hours to evacuate 50% of the contents of the digestive

tract and 33.4 hours to evacuate 90% of digestive tract contents, including fibres (50-500 µm)

and microbeads (>63 µm) (Grigorakis et al. 2017). This is similar to Clearnose Skate (Raja

eglanteria), requiring 48 hours to completely clear digestive tracts (Stehlik et al. 2015), and less

than Yellow Perch (Perca flavescens) which required 19.8 hours to clear 50% of digestive tract

contents and 65 hours to clear 90% of digestive tract contents (Gingras & Boisclair 2000).

Preliminary studies to assess the amount of time needed to clear pellet food and brine shrimp

from juvenile Rainbow Trout and Fathead Minnow digestive tracts determined that 24 hours was

sufficient. It is reasonable to suggest that the majority of digestive tract contents would be passed

within the 24-hour time frame of this feeding study.

Differences among microplastic types were not significant for White Sucker (G-Score= 12.0, 10

d.f., α= 0.05, p >0.2), though half of the microplastic tested had a higher mean abundance in the

digestive tract after 48 hours than in the water. The microplastic types that tend to be retained

more in the digestive tract than excreted in water after 24-48 hours were not consistent for

Rainbow Trout and White Sucker. The only microplastic type that was consistent for both

61

species was spherical microbeads. For Fathead Minnow, the majority of microplastics were

excreted within the 48 hour period and differences among microplastic types were not significant

(G-Score= 13.9, 10 d.f., α= 0.05, p >0.1); however, at least some microplastics were retained in

the digestive tract for every microplastic type except spherical microbeads. Clearly, there is

potential for at least some microplastics to be retained in the digestive tract as approximately

17% and 14% of ingested microplastic particles were retained across all treatments for juvenile

White Sucker and adult Fathead Minnow respectively after a period of 48 hours and additional

feeding of clean, microplastic-free food. However, no microplastics were visually observed to be

obstructing the esophagus during dissection of these two species. The type of microplastics

retained most abundantly in the digestive tract differs among species. The differences among

species cannot be statistically compared for the differences among microplastic types retained, as

the study design was amended following completion of the Rainbow Trout feeding experiment to

include an additional feeding of clean, microplastic-free food. The additional 24 hours added to

the time period and additional food in the feeding experiments may have allowed more time and

physical stimulus for the fish to pass microplastics. Also, the body size of White Sucker (10.8 ±

2.7 g) and Fathead Minnow (3.3 ± 1.1 g) used in these feeding experiments were smaller than for

Rainbow Trout (15.3 ± 5.0 g). The size range of microplastics was consistent among species

(125-1000 µm) for these feeding experiments, though individual particle sizes varied within this

range. The size of some particle types may have been too large for some of the White Sucker and

Fathead Minnow to consume, though this is not likely as some ingestion was still observed. The

method of feeding differed among species as well. Rainbow Trout were fed with 2 mm food

pellets prepared with microplastics baked into the pellets, whereas White Sucker and Fathead

Minnow were fed with microplastics frozen into brine shrimp slurry discs. The frozen brine

shrimp slurry discs melted fairly quickly once dropped in the jars of water containing the fish.

62

The more buoyant microplastics likely separated from the brine shrimp slurry and floated toward

the water surface more quickly than the microplastics that were baked into food pellets, which

tended to hold together under water fairly well for the three hour feeding period. This may

explain why rainbow trout consumed a greater percentage of the overall microplastics they were

exposed to (22%, Table 2) relative to White sucker (14%, Table 3) and Fathead Minnow (15%,

Table 4).

It is possible that the second feeding assisted in moving microplastics that were retained in the

first 24 hours as the two species tested using the amended design did not show significant

differences among microplastic types following the second feeding; however, at least some

microplastics were retained in the digestive tract after 48 hours despite including this additional

feeding. A study on microplastic ingestion in copepods determined that microplastics ingested in

the absence of food were retained up to 7 days, whereas microplastics ingested in combination

with food were typically egested within hours in fecal pellets (Cole et al. 2013). Fish larvae that

had ingested plastics passed the plastics and then fed on brine shrimp (Hoss & Settle 1990). Most

microplastics were likely egested by fish as five of six species were observed ingesting

microplastics, but only two of six species were observed with microplastics retained in their

digestive tracts (Hoss & Settle 1990). Goldfish fed 50 microplastics, followed by feeding to

satiation on pellet food, were observed to retain up to three microplastics for a period of six days

(Grigorakis et al. 2017). This is consistent with the White Sucker and Fathead Minnow feeding

experiments as the number of microplastics retained after feeding on microplastic-free food is

low relative to the initial number of microplastics exposed, but in almost every treatment some

microplastics still remained in the digestive tract 24 hours after the second feeding.

63

Implications

If microplastic particles are retained in the digestive tracts of aquatic organisms, they may

accumulate over time. The potential build-up of microplastic particles may cause physical effects

to the organisms that consume them. Physical effects include changes in the overall size and

condition of the organisms that ingest them. There were significant differences between fish with

and without microplastics in their digestive tracts for weight, length and condition factor in fish

collected from the English Channel (Foekema et al. 2013). Six fish contained more than one

microplastic particle, and the maximum number per fish was four particles (Foekema et al.

2013). These are similar quantities to the fish on the lower end of the spectrum for microplastic

retention from this study. Growth of larval Eurasian Perch also appears to be significantly

impacted by polystyrene microplastic exposure as fish were significantly smaller than control

fish (9.2 ± 0.1 mm) two weeks after hatching (Lönnstedt & Eklöv 2016). The Rainbow Trout and

White Sucker used in this feeding study were at the juvenile life stage. Fish were shown to have

ingested polystyrene foam beads during a life stage that may have implications for growth.

Feeding on polystyrene may have reduced the amount of nutritious food the fish could take in,

resulting in fewer resources for growth. The size of fish may impact their susceptibility to

predation, which in turn will affect population sizes.

Physical effects have also been observed at a lower level of biological organization. Changes in

gene transcript expression in the digestive glands of oysters exposed to polystyrene may indicate

a response to glucocorticoid stimulus; fatty acid catabolic processes, respiratory burst and

cellular response to mechanical stimulus were enriched (Sussarellu et al. 2016), which are signs

indicative of stress response and reduced energy intake from food. In addition, the increased size

64

of hemocytes in exposed oysters indicates oxidative stress response (Sussarellu et al. 2016).

Also, a direct energy-budget model revealed that simulated differences existed in the final dry

flesh mass and oocyte production for exposed oysters, which were higher in the budget model

than observed measurements (Sussarellu et al. 2016). Energy appears to be shifted toward

structural growth and maintenance versus reproduction (Sussarellu et al. 2016). While this

finding differs from the finding of Lönnstedt & Eklöv (2016), both indicate that growth and size

are impacted in some manner.

Microplastic ingestion has been shown to alter the behaviour of aquatic species as well.

Exposure to microplastics can alter feeding behaviour. Oysters exposed to microplastics had

significantly higher algal consumption and absorption efficiency, perhaps increasing food intake

to compensate for increased intake of non-nutritious polystyrene (Sussarellu et al. 2016).

Similarly, the presence of algae was shown to increase microplastic uptake in filter-feeding

copepods (Cole et al. 2013). Larval Eurasian Perch feeding behaviour was also significantly

impacted by exposure to polystyrene as larvae from the high microplastics treatment (80,000

particles per m3) consumed an average of 7.2 ± 1.2 polystyrene particles over 24 hours and their

stomachs contained only plastics and none of the brine shrimp food source provided, which the

authors interpreted as indicating a preference for polystyrene over food (Lönnstedt & Eklöv

2016). While feeding behaviour was not specifically investigated in this study, the potential for

blockage as observed in Rainbow Trout with ingested polystyrene foam beads and the spherical

microbeads accumulating in the mouth and gills could alter or prohibit feeding behaviour.

65

The behavioural effects of ingesting microplastic particles may extend beyond feeding

behaviour. For example, Eurasian Perch larvae exposed to polystyrene particles displayed

significantly altered behaviour, including lower activity rates, lower total distance moved, and

greater amount of time spent immobile (Lönnstedt & Eklöv 2016). Increased immobility in fish

that have consumed microplastics may make them less viable competitors for potential prey

items, or for potential mates. Juvenile Eurasian Perch also displayed a significant reduction in

activity in response to conspecific olfactory chemical alarm cues following polystyrene

exposure, and significantly reduced survival rates of 2-week-old larvae exposed to a natural and

common predator, Northern Pike (Esox lucius) (Lönnstedt & Eklöv 2016). Reduced survival in

fish consuming microplastics may directly affect population sizes. If the species used in this

study are similarly affected, there could be ecological implications. Rainbow Trout are preyed

upon mainly by diving birds and mammals in the Great Lakes, but are a known prey item for

other trout, char and Coho Salmon (Oncorhynchus kisutch) in other communities (Scott &

Crossman 1973). White Sucker are an important food item for other predatory fish species,

including Northern Pike, Muskellunge (Esox masquinongy), basses (Centrarchidae spp.),

Walleye (Sander vitreus) and Burbot (Lota lota), while juvenile White Sucker are an important

food source for Brook Trout (Salvelinus fontinalis) and Atlantic Salmon (Salmo salar) (Scott &

Crossman 1973). Fathead Minnow are an important food source for Smallmouth Bass

(Micropterus dolomieu) and other game fish, and is of extreme importance for its role in

converting algae and other organic detritus to food sources for other fish (Scott & Crossman

1973).

66

The feeding experiments were conducted over a short time frame. The main objective of these

feeding experiments was to determine whether some microplastic types have more or less

potential for retention in the digestive tract because those that do have potential for retention in

the digestive tract should be of greater concern when considering possible physical harm and

other potential effects on the fish that have ingested them, including reproductive effects.

Japanese Medaka (Oryzias latipes) exposed to environmentally relevant concentrations of

contaminants sorbed to polyethylene, and uncontaminated polyethylene showed significant

down-regulation of genes linked to female fecundity (Rochman et al. 2014). Also, one male

exposed to plastic with environmental contaminants displayed characteristics possibly leading to

sex-reversal or intersex (Rochman et al. 2014). Though this study in particular was not extended

to examine the actual implications on reproductive output, other studies have. In female oysters,

oocyte abundance and diameter were significantly lower, as was larval yield arising from

exposed females (Sussarellu et al. 2016). These indicate severe negative effects on reproductive

health in exposed female oysters. In male oysters, sperm velocity was significantly lower, which

may reduce the ability of sperm to fertilize oocytes (Sussarellu et al. 2016). Moreover, larval

growth resulting from exposed female oysters was significantly slower (Sussarellu et al. 2016).

Also, exposure significantly reduced egg-hatching rates in Eurasian Perch over a three-week

period (Lönnstedt & Eklöv 2016). The organisms directly ingesting microplastic particles were

affected, as well as the offspring produced from affected females. This demonstrates the

potential for impacts of microplastic particle exposure beyond a single generation. If the ability

of females to produce viable eggs and offspring is reduced, and the ability of the offspring to

survive is lowered, the overall population levels will decline.

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3.5 Conclusions

Retention of some microplastic particles occurred in all three species tested. The difference

among microplastic types was significant only in Rainbow Trout. In Rainbow Trout and White

Sucker, some microplastic types were retained more abundantly in digestive tracts than they

were excreted over a 24-48 hour period, and the types of microplastics that were retained more

abundantly than excreted varied between species. Apparent blockage of the esophagus was

observed in a juvenile Rainbow Trout that had ingested several polystyrene foam beads. The

potential consequences of microplastic ingestion could be severe if the ingested microplastics

block the esophagus and prevent feeding and digestion. Further studies are needed to examine

this in greater detail. Ingestion of all other microplastic types occurred in each species, though

blockage was not observed. Spherical microbeads were retained often in both juvenile Rainbow

Trout and in juvenile White Sucker. The spherical microbeads were observed adhering to gills

and accumulating in the mouth of juvenile Rainbow Trout. While other ingested microplastic

particles were not observed to accumulate in the same manner, a portion of the ingested particles

still remained in the digestive tracts of all three species after 24-48 hours.

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Chapter 4

Conclusions and Future Directions 4

4.1 Conclusions

In this thesis I have determined that methods involving the use of heat or generating heat 60 °C

or higher may cause the disappearance of some spherical microbeads. The difference among

microplastic types was very significant for percent recovery. For digestion, 1N and 4N KOH at

room temperature do not alter the integrity of polymers beyond recognition. Either of these

chemical digestion methods should be used in the digestion of fish tissue to remove organic

matter from microplastics.

There were significant differences among microplastic type for retention of microplastics in

Rainbow Trout digestive tracts. Polystyrene foam beads appeared to cause blockage in the

digestive tract of Rainbow Trout. Spherical microbeads appeared to accumulate in the gills and

mouth of Rainbow Trout. The differences among microplastic types were not significant for

retention of microplastics in White Sucker and Fathead Minnow digestive tracts; however,

microplastics of nearly all types were retained to some degree in the digestive tracts of all three

species tested.

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4.2 Changes to Experimental Design

Some alterations could be made to the experimental design of the fish feeding experiments.

Uniform particle sizes would provide more clarity in determining whether differences observed

are due to differences among microplastic types rather than differences due to particle sizes.

Similarly, using a smaller range in size of fish among species and consistent experimental design

would allow for statistical comparison among species, rather than simply within each individual

species. Using larger sample sizes for fish would allow for more robust statistical comparisons.

4.3 Future Directions

The feeding experiments in this thesis lead to several possible future experiments. Polystyrene

foam beads and spherical microbeads appeared to cause novel observations of potentially

harmful physical effects (blockage). Polystyrene foam beads should be used for long-term

feeding experiments to determine whether blockage is a common phenomenon in Rainbow

Trout, and whether blockage leads to nutritional effects and mortality. Spherical microbeads

should be used for long-term effects on Rainbow Trout to determine whether adhesion and

entrapment in gills leads to potential respiratory effects. It is important to explore the potential

relationships between particle sizes and potential for harm to determine the size range that should

be considered most problematic, and whether this range differs among life stages within the

same species, or among species. All microplastic types should be observed in long-term feeding

experiments to determine whether retention persists beyond 24-48 hours, and whether

accumulation occurs. Repeated exposure to microplastic particles may cause both accumulation

and retention of microplastic particles, leading to possible physical, behavioural and

toxicological effects on the organisms that have consumed them.

70

Additional research steps in this field are to investigate the ecological implications of

microplastics given that some microplastics have the potential to be retained in the digestive

tracts of fish for at least 24 to 48 hours. Though this study provided the basis for future

ecological studies, large-scale studies must be done to fully understand the ecological

implications. Future studies should examine trophic transfer of microplastic particles through the

consumption of other organisms. The microplastic particles detected in field-collected fish

cannot be attributed to direct consumption with certainty. The microplastic particles could have

been ingested by a prey species, and subsequently the predator species caught and sampled. It

has been established that filter-feeding invertebrates, mussels and oysters consume microplastics,

as well as the minnow and juvenile fish species in this thesis. Trophic transfer could occur if any

of these species are consumed, and the predatory organisms also retain the microplastics

particles. Humans can ingest microplastics particles by ingesting contaminated bivalves.

Assuming an average concentration of 0.42 particles per gram of tissue, humans who consume

mussels and oysters frequently will consume up to 11,000 microplastics per year, and infrequent

mussel and oyster consumers will consume up to 1,800 microplastics per year from these

shellfish alone (Van Cauwenberghe & Janssen 2014). Developing an understanding of ecological

implications all the way up to the human level will generate a greater understanding of the

severity of microplastics contamination, and ideally generate broader interest from the general

public. Public interest is essential in motivating people to reduce microplastics input into aquatic

ecosystems because microplastics are an anthropogenic contaminant, perhaps by substituting

other biodegradable materials for plastics wherever possible. For example, fruit seeds are

commonly used in place of microbeads in exfoliants, and corn or cassava starch products can be

used in place of polystyrene foam packaging (Salgado et al. 2008). Biodegradable feedstock can

71

be used as an alternative to pellets, and are made of materials like wood, straw, maize, cassava or

algae (InnProBio 2016).

Because of these potential ecological implications of microplastics exposure and retention, it is

essential that the most problematic microplastic types be identified in order to target policy

development toward reducing the input of the most problematic microplastics into aquatic

ecosystems. Results of these studies show that polystyrene foam and spherical microbeads are of

particular concern in terms of retention in Rainbow Trout, though none of the microplastic types

were consistently egested among all species. Policy has been developed in Canada and the

United States to ban the use of microbeads in rinse-off personal care products (United States

Congress 2015; Canada 2016). This legislation does not address microplastic in personal care

products that are not rinse-off, in household or industrial cleaning supplies or any of the

secondary microplastic types present in the environment. There is a need for policy development

to regulate the input of microplastics into aquatic ecosystems besides microbeads, as this

category represents only a fraction of the microplastics observed.

72

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Appendices

A. 1. The mean total length ± the standard deviation (cm), wet weight ± the standard deviation (g) and wet weight of the digestive tract ± the standard deviation (g) for Rainbow Trout, White Sucker and Fathead Minnow used in the 24-48 hour feeding experiments. For each species, a total of 51 fish were used.

Total Length (cm) Wet Weight (g) Digestive Tract Wet Weight (g)

Rainbow Trout 11.9 ± 1.5 15.3 ± 5.0 6.5 ± 0.9

White Sucker 10.0 ± 1.1 10.8 ± 2.7 0.2 ± 0.1

Fathead Minnow 6.5 ± 0.8 3.3 ± 1.1 0.1 ± 0.1

A. 2. G-test results comparing differences among microplastic types for each species.

G-Score Degrees of Freedom

p-Value

Rainbow Trout 64.4 5 <0.0001

White Sucker 12.0 10 0.282

Fathead Minnow 13.9 10 0.180

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A. 3. The mean number of microplastic particles ± the standard deviation in each phase for Rainbow Trout (n=8). The initial water sample (W1) represents the number of microplastic particles out of the initial 20 particles that were not consumed, and W2 is the number of microplastic particles present in the water after 24 hours, presumably excreted. The numbers of microplastic particles in the digestive tracts are under G+GC. Treatments include control (C), spherical microbeads (SB), irregularly shaped microbeads (IB), fragments (F), shavings, (S), polystyrene foam beads (PSF) and synthetic fibres (SF).

W1 W2 G+GC

C 0 ± 0 0 ± 0 0 ± 0

SB 14.4 ± 3.6 0.8 ± 0.9 1.8 ± 2.2

IB 14.4 ± 3.7 3.3 ± 2.1 1.1 ± 0.9

F 14.3 ± 4.1 0.9 ± 1.1 1.23 ± 2.5

S 14.3 ± 5.0 2.1 ± 1.8 0.8 ± 1.6

PSF 17.8 ± 4.2 0.2 ± 0.4 1.9 ± 4.1

SF 10.3 ± 3.4 8.0 ± 3.5 1.6 ± 2.5

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A. 4. The mean number of microplastic particles ± the standard deviation in each phase for White Sucker (n=8). The initial water sample (W1) represents the number of microplastic particles out of the initial 20 particles that were not consumed, and W2 is the number of microplastic particles present in the water after 24 hours, presumably excreted. Fish were fed a second time with microplastic-free food, and allowed an additional 24 hours to pass any microplastic particles (W3). The numbers of microplastic particles in the digestive tracts are under G+GC. Treatments include control (C), spherical microbeads (SB), irregularly shaped microbeads (IB), fragments (F), shavings, (S), polystyrene foam beads (PSF) and synthetic fibres (SF).

W1 W2 W3 G+GC

C 2.0 ± 1.0 2.0 ± 0 3.0 ± 1.00 0.50 ± 0.5

SB 17.3 ± 2.1 1.8 ± 2.1 0.1 ± 0.4 0.2 ± 0.6

IB 13.6 ± 1.2 4.6 ± 2.9 1.1 ± 1.5 0.8 ± 0.9

F 18.0 ± 2.5 1.5 ± 2.0 0 ± 0 9.1 e-2 ± 0.3

S 17.4 ± 0.9 1.63± 1.2 0.3± 0.5 0.5 ± 0.9

PSF 19.0 ± 1.3 0.3 ± 0.5 0.1 ± 0.4 0 ± 0

SF 16.3 ± 1.0 1.5 ± 1.1 0.6 ± 0.5 0.5 ± 0.8

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A. 5. The mean number of microplastic particles ± the standard deviation in each phase for Fathead Minnow (n=8). The initial water sample (W1) represents the number of microplastic particles out of the initial 20 particles that were not consumed, and W2 is the number of microplastic particles present in the water after 24 hours, presumably excreted. Fish were fed a second time with microplastic-free food, and allowed an additional 24 hours to pass any microplastic particles (W3). The numbers of microplastic particles in the digestive tracts are under G+GC. Treatments include control (C), spherical microbeads (SB), irregularly shaped microbeads (IB), fragments (F), shavings, (S), polystyrene foam beads (PSF) and synthetic fibres (SF).

W1 W2 W3 G+GC

C 0.7 ± 0.6 0.7 ± 0.6 0.7 ± 0.6 0.7 ± 0.6

SB 16.5 ± 2.5 1.8 ± 2.4 0.9 ± 1.4 0 ± 0

IB 14.9 ± 2.8 1.9 ± 0.8 1.5 ± 1.2 0.3 ± 0.5

F 15.6 ± 6.4 0.8 ± 1.0 0.3 ± 0.5 0.2 ± 0.6

S 15.5 ± 3.4 2.4 ± 1.3 1.3 ± 1.2 0.6 ± 0.9

PSF 19.0 ± 0.6 0.3 ± 0.5 0.2 ± 0.4 9.1 e-2 ± 0.3

SF 12.4 ± 3.4 3.4 ± 1.8 0.8 ± 0.7 0.7 ± 0.8

85

A. 6. Examples of microplastic types used in the assessment of chemical digestion methods in Chapter 2.

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A. 7. An experimental fibre from a carpet sample and used in the assessment of chemical digestion methods (Chapter 2) and the 24-48 hour feeding experiments (Chapter 3), next to a visibly different fibre identified as a contamination fibre in the 24-48 hour feeding experiments. Synthetic fibres used in this thesis appear brightly coloured and wiry, whereas contamination fibres are typically black in colour and resemble clothing fibres.