MSc Chemistry

Analytical Sciences Track

Literature Thesis

The fate of engineered nanomaterials (ENMs) forming eco-

corona in aquatic systems

Characterization and impacts

by

Dimitra Ntagiou

12983586

July 2021

12 ECTS

April 21-July 14

Daily Supervisor: Examiners:

Dr. Sebastian Kuehr Dr. Alina Astefanei

Dr. Saer Samanipour

Second supervisor:

Dr. Anastasia Georgantzopoulou

University of Amsterdam/Norwegian Institute for Water Research

1

Contents

Abstract ................................................................................................................................................. 2

1.Introduction ....................................................................................................................................... 3

2. Eco-corona formation: environmental conditions, NOM and ENM properties ................................ 6

2.1 Environmental properties and media composition .................................................................... 6

2.2 ENM properties (surface functionalization and size) .................................................................. 7

2.3 Composition and size of biomolecules ........................................................................................ 8

3. Characterization and quantification of eco-corona .......................................................................... 9

3.1 Principal methods of bio-nano size and morphology characterization ...................................... 9

3.1 Principal methods of protein/NOM identification and quantification ..................................... 11

4.Exposure conditions that influence eco-corona formation and the effects in aquatic organisms .. 14

4.1 Summary of bio-nano interface ................................................................................................ 14

4.2 NOM concentration and ENM surface coatings ....................................................................... 17

4.3 ENM ageing ............................................................................................................................... 17

4.4 Conditioning medium and incubation time .............................................................................. 18

5. Conclusions and perspectives ......................................................................................................... 20

List of abbreviations ............................................................................................................................ 22

Acknowledgments ............................................................................................................................... 23

References ........................................................................................................................................... 24

2

Abstract

Engineered nanomaterials (ENMs) are inorganic or organic materials, sized 1 to 100 nm in at least

one dimension, with large surface area and various unique physical and chemical properties,

including high surface area, high chemical reactivity, and electric conductivity. Due to these unique

properties they have found applications in several fields such as, medicine, electronics, textile,

cosmetics and personal care products, and food processing. Once released into the aquatic

environment via e.g. industrial wastewater or agricultural leachates, ENMs can interact with

several surrounding ecological macromolecules leading to the adsorption of such molecules onto

their surface and the formation of eco-corona or environmental corona. These macromolecules

can be natural organic matter (NOM), organism’s and cell’s secreted proteins and metabolites, or

extracellular polymeric substances (EPS). In turn, the eco-corona confers another identity to ENMs,

while the eco-nano interface can alter the ENM physico-chemical characteristics, bioreactivity,

bioavailability, and thus toxicity to aquatic organisms. The understanding of the eco-corona

formation in the environment and its relevance in the ecotoxicological studies is still challenging

and emerging, and limited research has been performed on the ENM biotransformation and the

analytical methodologies for quantification. Therefore, in this review the various conditions (pH,

ionic strength, concentration of eco-molecules) according to protocols and reported studies, as

well as the ENM properties (size, coatings, hydrophobicity) that facilitate the eco-corona formation

are discussed. Next, the analytical techniques that have been used to characterize and quantify the

biomolecules bound on the ENM surface in the environment are described. In addition, the fate,

and transformations of ENMs and biomolecules after their interaction (bio-nano interface) are

summarized. Furthermore, the role of eco-corona on ENM interaction with cells and organisms and

its link to bioavailability, uptake and potential ecotoxicological effects in combination with

reported exposure conditions (NOM/biomolecule concentration, ENM coatings and ageing,

conditioning, and incubation time) are highlighted. Finally, the gaps, the challenges and the needs

for future research are discussed both from an analytical and ecotoxicological perspective.

3

1.Introduction

Nanotechnology is the most evolving and promising field of the 21st century, counting many

applications and new materials in medicine, cosmetics, textiles, energy, electronics, and the

environment. The engineered nanomaterials (ENM)s are materials with at least one dimension in

the size of 1 to 100 nm designed to have advanced properties compared to their bulk material and

can be used efficiently in various applications of our daily life, due to their small size and large

surface area [1]–[3]. Due to these special properties the total nanomaterial-containing product

number increased from 54 in 2005 to 1814 in 2014 as reported by the Nanotechnology Consumer

Product Inventory (CPI) [4]. In 2021, the number of 5110 nanomaterials is reported by The

Nanodatabase, an online inventory of commercial products [5].

The types of ENMs developed, according to the literature, are inorganic or metal-based (Ag, Au, Zn,

Ni, Fe, Cu, TiO2, Fe3O4, SiO2, CeO2, MnO2 and Al2O3), quantum dots and silica, and organic such as

polymer-based, carbon-based materials (nanotubes and fullerene) [6], with the most used ENMs

these of TiO2, Au and Ag ENMs [4]. Their advanced properties include among others their electric

conductivity, optic and magnetic properties but also their chemical reactivity and could thus be

found in wide range of usage [7], [8]. Especially the properties to catalyze reactions, to break down

organic compounds or to bind/adsorb contaminants or heavy metals to their surface make these

materials useful [9], [10]. For instance, carbon- based materials, such as modified graphene oxide

and carbon nanotubes (CNTs) are reported to be used for remediation of aqueous organic

contaminants and metals, by adsorption on the ENM layer or their pore structure [11]. Another

example of adsorption is Fe ENMs which can remove heavy metals or chlorinated compounds from

aqueous environments [12], [13]. Inorganic nanomaterials, such as Ag ENM own antibacterial and

antiviral properties through the release of Ag+ ions and thus applied as water disinfectants as well

as, in many personal care and consumer products, textiles, and food processing [14], [15]. Also,

Ti/TiO2 ENMs act as photocatalysts by absorbing light, oxidizing toxic organic pollutants into

harmless species and turning them into H2O and CO2 [16]. During this process reactive oxygen

species (ROS) can be generated, such as hydrogen peroxide (H2O2), singlet oxygen (O2), or hydroxyl

radicals (OH∙), which can lead to the induction of oxidative stress in living organisms [17].

Due to their unique properties and enhanced reactivity, ENMs can interact with different

constituents present in the biological and natural environment. When ENMs get in contact with the

biological environment and biofluids, several biomolecules are adsorbed onto their surface [18].

This formation is called ‘’biocorona’’, or ‘’protein corona’’ when the biomolecules correspond to

proteins. Studies have observed that this formation provide the ENM with a new identity [19]. The

protein corona can play a critical role on bio-nano interactions, thus, the bioavailability, immune

response and clearance, and potential toxicity [20], [21]. The bio-nano research has shown that

two protein layers can be formed on ENM surface depending on their affinity, the high affinity

layer (hard corona, inner layer of hard-bound proteins) and the low affinity layer (soft corona,

outer layer of weakly bound proteins) [22]. Through the years plenty of studies have focused on

the dynamic interactions and behavior of hard and soft protein corona and its importance in

applications of nanomedicine and nano-safety highlighting the complexity of this formation on

ENM transformation, fate and effects [23], [24].

4

In the natural environment, entry of ENMs is possible via direct release from manufacturing

processes waste or indirect release from product disposal [25]. One part of the released ENMs

ends up in the aquatic environment as they either reach it directly via wastewater or by leachates

from agricultural fields that have been fertilized with ENM containing sewage sludge [26]. The

aquatic environments contain ecological macromolecules, the natural organic matter (NOM) such

as dissolved organic matter (DOM), plant and animal decomposed material derived from soil, as

well as surfactants coming from wastewater. NOM in the aquatic environment is an heterogeneous

mixture and mainly composed of humic substances (humic acid (HA) and fulvic acid (FA)),

extracellular polymeric substances (EPS organic biopolymer aggregates rich in proteins and

polysaccharides) which compose the DOM content, but also large biomacromolecules, smallest

living organisms such as prokaryotes, viruses and cell debris (refractory organic cell walls) [27]. In

fresh and coastal waters, humic substances (compounds from plant decomposition) derive from

soil (pedogenic) and represent the largest fraction of NOM content [28]–[30]. The aquatic HA

composition differs from FA since it is a high molecular weight (MW) substance and a negative

charged molecule due to its carboxylic and phenolic function groups, while FA is a lower MW

molecule with higher functionality.

Released ENMs can interact with NOM by π-π bonding mechanism on their surface [31]. For the

first time, such kind of ENM-biomacromolecule formation was called ‘’eco-corona’’ [32]. It is also

known as macromolecular corona or ecological corona [32], [33]. The formation of an eco-corona,

similar to protein-corona may cause alterations in the physicochemical properties of both: ENMs

and biomolecules bound to them - and thus their fate, bioavailability and potential ecotoxicological

effects [34]–[36]. According to a recent study it was observed that toxic effects are mediated

though the eco-corona formation [37]. The fate of ENMs when NOM and other colloids are

adsorbed onto their surface in aquatic environments is depicted in the Figure 1. The eco-nano

interaction can stabilize or destabilize the ENMs resulting in aggregation and sedimentation in

aquatic systems, or in other case surrounding biomolecules can bind to the existing adsorbed ones

upon the ENM surface.

Figure 1. Interaction of ENMs with NOM and other colloids, behavior, and fate in aquatic environments

(from:[38])

5

Even though adequate research is conducted on the entry path, fate, and the effects of ENMs in the

aquatic ecosystem (e.g. [38]–[40]), the ENM property to bind/ adsorb organic compounds and

proteins (eco-corona) and the protein-corona- ENM bio-transformation are only taken into

account to a limited extent.

Also, only few studies focus on or include the characterization of eco-corona and the induced

alterations on the ENM fate and ecotoxicological effects. The same applies to (standardized)

approaches to artificially induce eco-corona formation in order to use these transformed ENMs in

laboratory studies combined with extensive characterization [41]–[43].

This literature review aims to elaborate on the ENM-biomolecule interaction, its role on the

interaction with cells and organisms, the factors that govern such interactions and provide a

comprehensive overview of the analytical techniques that can be used to study the process of

eco-corona formation. Thus, in this review, the following aspects will be covered and discussed:

- the environmental conditions (e.g. pH, presence of proteins, DOM, and NOM) and ENM

properties that influence the eco-corona formation, as well as the existing protocols for artificial

induction of corona formation under experimental /lab conditions

- the analytical methods for the characterization and quantification of the biomolecules forming

the eco-corona that have been successfully used in (ecotoxicological) studies

- the biomolecule-ENM alteration as well as how this may impact the environmental fate and

(eco-) toxicological effects of ENM in aquatic systems

- the relevance of the corona characterization for (eco-) toxicological studies as well as

recommendations for future research and implementation of eco-corona induction and

characterization in regulatory risk assessment and associated test protocols regarding the B

(bioconcentration) and T (toxicity) criterion [44].

6

2. Eco-corona formation: environmental conditions, NOM and ENM properties Recently studies taking into account the chemical and physical transformation of ENM in the

environment, for example by using environmentally relevant exposure media (sewage plant

effluents or sewage sludge) or artificially transformed ENMs [45]–[49] have shown that such

transformations can result in altered toxicological effects highlighting the importance of

incorporation of ENM transformation in effect studies [50]. This is also in line with the call for the

adaptation and development of standard test for regulatory risk assessment [51], [52]. The

Organization for Economic Co-operation and Development (OECD) and the International

Organization for Standardization (ISO) have established the required test guidelines (TG) related to

sample preparation and dosimetry for ecotoxicological testing. However, the presence of DOM has

not been taken into account, avoiding the complexity of the solution, even if natural waters

contain always organic matter [53]. Also, the European Chemical Agency (ECHA) has reported the

need to include the biotransformation into PBT (persistent, bioaccumulative, toxic) and risk

assessment of chemicals into the aquatic environments [54].

In order to understand the primary process of eco-corona formation and the acquired

characteristics of ENMs that govern their fate and effects in the aquatic ecosystems, it is crucial to

consider the most important parameters controlling the eco-nano behavior, bioavailability and

toxicity. For this reason, laboratory settings attempt to reproduce the process under a more

realistic environment and exposure conditions. The main parameters controlling the bioavailability

and toxicity are environmental parameters (the exposure media composition, pH, ionic strength,

and incubation time) ENM properties (size, charge, ageing and hydrophobicity) and DOM

concentration and composition.

2.1 Environmental properties and media composition

Natural organic matter, pH, salinity, and ionic strength of the aquatic media determine the surface

charge that can influence the eco-corona formation, the aggregation rate, the hydrodynamic size

and dispersion of ENMs. According to OECD TG 318, the NOM added into the TiO2 and Ag ENMs

dispersion acts as a stabilizer, at high pH values [55]. The reason is that at pH values lower than the

point of zero charge (pH<pHpzc=6), the TiO2 ENMs have a positive surface charge. Thus, by the

addition of NOM like Suwannee River humic acid (SRHA), which bear a negative charge in water,

the ENM surface gets neutralized leading to a decreased electrostatic repulsion between the ENMs

and an increase in the aggregation rate. By the excessive addition of HA, the surface charge is

reversed making the TiO2 ENMs more stable [56], [57]. A similar behavior was observed when ZnO

ENMs (pHpzc=9.3) and FeO (pHpzc=7.8) were tested [58], [59]. Therefore, higher pH provides ENM

dispersion stability in presence of NOM and controlled aggregation behavior through electrostatic

forces.

Salinity is also another parameter which influences the potential eco-corona formation

interactions, especially when the media is seawater. Regarding the salts, Ca(NO3)2 salt with the

multivalent cations of Ca2+ has been selected as major electrolyte, at low ionic strength (IS=1.7

7

mM), representing the natural water better than other ions [55]. The presence of CaCl2 salt at high

concentrations (0.08 M) promotes aggregation of CeO2 ENMs because of the bridging effect

between HA-coated ENMs, in contrast to KCl which provides stability at the same concentration

[60]. Additionally, TiO2 ENMs showed higher adsorption rate for EPS at pH value higher than 6 and

at NaCl concentration lower than 1 mM [61]. In contrast, at higher ionic strengths (50 and 100 mM

NaCl) the various types of HA were adsorbed tightly on TiO2 ENMs due to the ‘’salting out’’ effect

or NOM shape changes from linear to spherical [62]. This result can be attributed to other factors

such as ENM size or heterogeneity [63]. An important study of Fernando et al examined the effect

of the combination of tightly and loosely-bound EPS and ionic strength on the agglomeration rate

of Ag ENMs [64]. The results indicated that the EPS corona reduced the agglomeration of ENMs in

the presence of NaNO3 and at low concentrations of Ca(NO3)2 equal to 0.05-1.0 mM because of the

increased steric repulsions. These results indicate that at high pH values and low ionic strength the

repulsion forces are increased between ENMs and the interactions between biomolecules and

ENMs are enhanced. However, there are studies that support the opposite, for example, the

increased salinity (100 mM) by the addition of NaCl and CaCl2 facilitated the eco-corona formation

between protein bovine serum albumin (BSA) and polystyrene (PS)-COOH ENM promoting the

heteroaggregation, compared to amidine polystyrene (NH2-PS) ENM where the aggregation was

depressed [65].

2.2 ENM properties (surface functionalization and size)

The surface functionalization of ENMs influences their charge and hydrophobicity, thus the way

that the macromolecules will bind on the ENM surface with electrostatic or steric interactions. For

example, Z- potential results showed that the presence of 2 mg/L Suwannee River Humic Acid

(SRHA) or alginate led to a decrease in the surface charge of positively charged (NH2-PS) ENMs (200

nm) from +50 mV to -39 and -47 mV, respectively. On the contrary, the charge of the negatively

charged carboxyl-PS ENMs did not change under the same conditions [43]. Also, another study

showed that negatively charged PS ENMs (PS-COOH) bound with EPS decreased their aggregation

in the high salt content water (high ionic strength) by promoting steric repulsion between colloids

[66]. A study of Wu et al. showed that the negatively charged NOM and EPS did not affect the

behavior of the negatively charged PS-COOH ENMs, and provided them stability, compared to the

positively charged NH2-PS ENMs, as their surface is firstly neutralized and then covered by the

adsorbed biomolecules [67].

The adsorption of biomolecules on the surface of ENMs also depends on the ENM size. For

instance, comparing 10, 60 and 100 nm Au ENM, in the presence of SRHA the molecules adsorbed

only on the surface of 10 nm spherical Au ENMs decreasing the ability to cause algal cell’s

membrane damage compared to the bigger-sized Au ENMs [68]. The reason is that the small-sized

ENMs have larger specific area and the active sites for binding with SRHA, decreasing their adverse

effects. Consistent to that, the HA and FA adsorption on the 20 nm SiO2 ENMs surface was stronger

compared to 100 and 500 nm [69]. The explanation of the higher affinity was the high reactivity

and the higher specific surface area of 20 nm which was 578 m2/g compared to 100 and 500 nm

which was 244 and 135 m2/g, respectively. Another study also supported that the eco-corona

formation is ENM size dependent, when 200 nm and 500/1000 nm PS ENM was tested with

positively charged lysozyme in seawater [70]. The results showed that the eco-corona was formed

8

only on 200 nm PS ENMs through colloidal steric hindrance and electrostatic repulsions.

Hence, the surface functionalization and the size are significant factors that determine the

potential of biomolecule adsorption on ENM surface and the eco-corona formation since smaller

and positively charged ENMs enhance the adsorption rate.

2.3 Composition and size of biomolecules

The NOM in the aquatic environment mainly consists of carbon, oxygen, nitrogen, hydrogen, and

sulfur. Utilizing Fourier-transform ion cyclotron resonance mass spectrometry (FT-ICR-MS), the

NOM corona formed on the surface of Ag ENMs by two different high molecular weight NOMs

(Yukon river(YRNOM) and Milwaukee River (MRNOM)) in 0.1 mM NaHCO3 showed that the bond

was rich in N and S-containing compounds, which was different from the initial NOM composition.

Also, NOMs of high molecular weight, high saturation and high number of oxygenated groups

preferentially form the corona [71]. The interactions of NOM with Au depend on the origin of the

NOM, the aromaticity and molecular weight [72]. In a similar way, EPS adsorption is associated

with their molecular weight and aromaticity. High molecular weight EPS (>100 kDa) exhibit

stronger adsorption on TiO2 ENMs surface than the low molecular weight EPS (<1 kDa) due to their

higher aromatic composition [73]. Similar to NOMs, EPS also behave differently comparing the two

types of algae-derived EPS (cell surface and soluble EPS) on anatase and rutile TiO2. Specifically, cell

surface bound EPS (loose binding polymers, microbial sheath) adsorbed stronger on anatase TiO2

than the soluble EPS (soluble macromolecules, colloids) from Chlorella pyrenoidosa, causing higher

toxicity due to heteroagglomeration (loose binding formation) with the algae. The soluble EPS

promoted homoaggregation (strong binding formation) of NPs due to their higher hydrophobic

interactions (higher adsorption because of higher EPS polysaccharide content), leading to

sedimentation [74]. Alginate, which is a linear biopolymer EPS can enhance the stability of CuO and

Ag ENMs, suspensions indicating electrostatic stabilization with the contribution of steric repulsion

[75], [76]. EPS contains proteins and carbohydrates as well, which determine their hydrophobicity

and the interactions with the ENMs. For instance, when the secreted EPS have low protein content

they are more hydrophobic (contain more carbohydrates) and a higher EPS concentration (100

ppb) is needed for the ENM-EPS aggregate formation compared to protein-rich EPS which need

only 10 ppb to form the eco-corona [77]. Therefore, the composition and the origin of the various

biomolecules present in aquatic ecosystems should be further specified in eco-corona formation

with ENMs.

3. Characterization and quantification of eco-corona

The analytical methods needed to understand how the eco-corona is formed and how it affects the

behavior of ENMs can be classified into two categories:

i. Methods that characterize the ENM size and morphology prior to and after the corona formation

ii. Methods that quantify and identify the adsorbed proteins or NOM molecules on the

nanomaterial surface

3.1 Principal methods of bio-nano size and morphology characterization

A recent comprehensive review reported the main methods for the characterization of DOM

adsorbed onto nanomaterial surface [78]. Here, relevant characterization methods, besides ENM

size and surface charge changes, are introduced so as to provide the main processes and aspects of

eco-corona formation such as adsorption rate, composition of biomolecules, and affinity degree. In

Table 1 the analytical techniques reported in literature and the respective properties can be

measured in order to characterize the eco-corona formation are mentioned.

Starting with the adsorption rate, a very simple technique to measure the ENM ability to NOM or

protein adsorption is UV-visible spectrophotometry, by calculating the signal of the dissociative

amounts of aromatic or aliphatic molecules unbound in freshwater and seawater medium [62],

[79]. A more sensitive and selective method compared to UV-Vis spectroscopy is the fluorescence

spectroscopy. Besides monitoring the fate of fluorescently labelled ENMs in aquatic organisms, and

the interaction with NOM and EPS, [80], [81] the changes of DOM fluorophores can also be

demonstrated upon adsorption. For instance, changes in fluorescence index (FI) revealed that the

terrestrial derived decomposed NOM preferentially binds to carbon nanotubes (CNT) and only high

molecular weight fractions of NOM were adsorbed due to hydrophilicity of low molecular weight

NOM [82]. In order to cover a large range of DOM fluorophores, three dimensional-excitation-

emission matrix (3D-EEM) is used, however the several fluorophore peaks may overlap each other

and particle coupling parallel factor analysis (PARAFAC) can be combined to avoid these limitations

[83]. In addition, the composition and concentration of the NOM medium that affects the

properties of ENMs, as well as the adsorption kinetics and thermodynamics can be evaluated by

measurement of the total or dissolved organic carbon (TOC/DOC) [84]. TOC determination is based

on the catalytic oxidation combustion technique at high temperature to convert the organic carbon

into carbon dioxide (CO2) which is measured. For instance, the maximum adsorption capacity,

according to Langmuir, of surface water NOM on the surface of raw carbon nanotube (CNT),

heated CNT and granulated active carbon was 22.0, 26.1 and 14.7 mg NOM/g CNT, respectively

[85]. In addition, thermal gravimetric analysis (TGA) is a useful tool to determine the amount of

organic matter adsorbed on ENMs. Based on the TGA curve, it has been shown that the amount of

Aldrich and Leonardite HA on Fe3O4 was 15.7 and 18.4,% respectively [86] while, in another study

the amount was 50% [87].

10

Table 1. Analytical methods used for characterization of ENM properties during the eco-corona formation and determination of (bio)molecules adsorbed onto ENM surface

Technique Measured properties

ENM size, concentration, and charge

Dynamic light scattering (DLS) Hydrodynamic diameter

Transmission electron microscopy (TEM) Diameter, morphology, shape, composition

Scanning electron microscopy (SEM)

SEM-EDX

Diameter, morphology, shape

Elemental composition of NOM bound

Nanoparticle tracking analysis (NTA) Diameter

Flow field-flow fractionation (FlFFF)

Hydrodynamic diameter, size distribution, hard/soft

corona

FlFFF-multi-angle light scattering (MALS) Size distribution

High-performance size exclusion chromatography (HPSEC) Size polydispersity, hydrodynamic diameter

Electrophoretic mobility analysis Z-potential

Inductively coupled plasma-optical emission spectrometry

(ICP-OES)

Total metal concentration

Inductively coupled plasma-mass spectrometry (ICP-MS) Total metal concentration

Single particle ICP-MS (spICP-MS) Metallic particle size and concentration

Pyrolysis-gas chromatography-mass spectrometry (py-GC-MS) Organic ENM identification and concentration,

composition

Thermal desorption GC-MS (TDS-GC-MS) Organic ENM identification and concentration,

composition

Raman spectroscopy ENM concentration, composition

Photothermal absorption correlation spectroscopy (PhACS) Light-absorbing ENM composition

Fluorescence correlation spectroscopy (FCS) Fluorescently labelled-ENM composition

Confocal microscopy Fluorescently labelled-ENM uptake visualization

Adsorption rate

UV-Visible spectrophotometry Aggregation/disaggregation, adsorption rate

Fluorescence spectroscopy

Three-dimensional-excitation-emission matrix (EEM)

Adsorption rate, NOM size

Thermal gravimetric analysis (TGA) Organic matter mass absorbed

Affinity

Isothermal titration calorimetry (ITC) Affinity determination

Surface plasmon resonance (SPR) Affinity determination, protein concentration

Biolayer interferometry (BLI) Affinity determination, DOM concentration

Quartz Crystal Microbalance with Dissipation (QCM-D) Affinity determination

Affinity capillary electrophoresis (ACE) Size, protein affinity determination (Au, Ag)

Morphology

X-ray photoelectron spectroscopy (XPS) Size, morphology, chemical bonds, changes in

crystallization

Atomic force microscopy (AFM) Morphology, NOM sorption location, structure

Circular dichroism (CD) Protein secondary structure changes

NOM composition

Fourier transform infrared spectroscopy (FTIR) NOM composition

Fourier transform-ion cyclotron resonance-mass spectrometry NOM composition

Nuclear magnetic resonance (NMR) NOM composition

Total/ Dissolved organic carbon (TOC/DOC) analysis NOM concentration, carbon concentration

11

Three methods that quantify the affinity of the biomolecule interaction with the nanomaterials is

isothermal titration calorimetry (ITC), surface plasmon resonance (SPR), biolayer interferometry

(BLI) and quartz crystal microbalance with dissipation (QCM-D) [19], [88]. SPR measures the

frequency and amplitude shifts of protein or NOM binding on the surface of immobilized silver or

gold-coated NPs. In the same way, in BLI, the ENM is immobilized onto a biosensor tip and the

binding rate is monitored for the interpretation of the binding affinity and concentration of DOM

on ENM surface. For example, in a study using BLI, it was observed that as the MW of the HA

fractions increased, the adsorption rate of HA onto the graphene oxide (GO) surface increased as

well as, the HA had greater affinity than BSA on GO surface [89]. The reverse is happening in QCM-

D, where the NOM or EPS solution is immobilized for assessing the affinity with Ag ENMs, where it

was revealed that it was dependent on the composition of DOM bound on silica and the coatings

of Ag ENMs (citrate, polyvinylpyrrolidone (PVP)) [90].

Furthermore, nuclear magnetic resonance (NMR) spectroscopy and X-ray crystallography can

depict the 3-D structure of the proteins or the organic matter associated with ENMs, while circular

dichroism (CD) can reveal the changes in the secondary structure of protein molecules upon

binding [91][92]. Regarding the types of NMR, 13C-NMR is used to depict the changes in aromatic

and aliphatic fractions of DOM [93], 1H-NMR can show the NOM content depending on their

derivation [94] and liquid state 1H-NMR can reveal displacement changes between NOM and EPS

with ENM coatings (PVP and citrate) [90]. However, any 2D-NMR was applied to date in order to

provide comprehensive information on the adsorption and bio-nano interaction. Also, Fourier

transform infrared spectroscopy (FTIR) is a tool for the demonstration of changes like intensity or

shift of DOM spectra peaks upon interaction with ENMs. An example of such changes is this of

multi-walled carbon nanotubes (MWCNTs) showing a decrease in peak intensity of 1628 cm-1 and

an absence of other peaks (1212 and 1720 cm-1) indicating the corresponding functional groups of

C=C, COOH and OH to get involved in the adsorption of NOM on MWCNs [31]. Generally, the

interpretation of IR results is challenging due to overlapping of bands so it can be used as

supplementary.

An important spectrometric method, the ultra-high resolution-Fourier transform-ion cyclotron

resonance-mass spectrometry (FT-ICR-MS) is a quite accurate tool (error <0.5 ppm) for the

identification of NOM molecular composition, including the functional groups of CHO,CHOS, CHON

and CHONS, on the surface of metal ENMs [95], [96]. However, it provides mostly qualitative

information and other kind of ENMs such as organic ENMs have not been tested by this technique.

3.1 Principal methods of protein/NOM identification and quantification

Two important steps are needed for the quantification of biomolecules; the preconcentration and

the separation of ENMs from the biomolecules bound. The preconcentration, as reported, can be

carried out using techniques such as ultrafiltration, cloud point extraction (CPE), and continuous

flow centrifugation (CFC). The separation can be achieved using methods such as membrane

filtration, size-exclusion chromatography (SEC) and flow field flow fractionation (FlFFF) [97].

Centrifugation only isolates all the proteins bound on the surface without any distinction of the

bio-nano properties of corona [98] although, there is study which reported that the centrifugation

remove efficiently the hard corona proteins from ENMs [99]. In Table 2 are shown some examples

of techniques for the isolation/ separation and the determination of biomolecules, forming eco-

12

corona, as well as the number of identified proteins.

The three methods combined for the determination of biomolecules concentration on the

nanomaterial surface that were adopted by some studies is centrifugation-sodium dodecyl sulfate

polyacrylamide gel electrophoresis (SDS-PAGE)-mass spectrometry [42], [66], [100]. Monopoli et al.

have described a protocol for protein isolation and quantification after the incubation process,

used by many experiments and is well-established [101]. According to this, after centrifugation and

washing steps with phosphate buffered saline (PBS) for the isolation of proteins, the proteins are

recovered by the addition of SDS surfactant buffer (disruptor agent) and separated by 1D SDS-

PAGE analysis. A similar and advanced approach has been described by Docter et al. including

short-term centrifugation, with the addition of sucrose cushion, to prevent additional binding of

secreted proteins with the ENMs due to their contact [102]. The disadvantages of the PAGE

analysis are the important loss of low abundance or low-intense proteins, while they elute in gel

buffer. Instead of SDS-PAGE, an LC-MS method can be implemented, identifying the corona

proteins after trypsin digestion, without providing information on the binding affinity [103]. The

Minimum Information about Nanomaterial Biocorona Experiments (MINBE) has established the

necessary guidelines for bio-corona characterization experimental setups, utilizing LC-MS for

obtaining reliable and highly reproducible results, although it is not yet applied in eco-corona

determination [104]. Specifically, these guidelines are referred to the experimental design for the

bio-nano interaction determination, regarding the exposure media (organisms, processing, and

storage), the reaction conditions (sample conditions and concentrations), the ENM

characterization, the separation techniques and processing (isolation and protein digestion) as well

as, the biomolecule identification and quantification.

Instead of LC-MS, capillary electrophoresis coupled with electrospray ionization mass spectrometry

(CESI-MS) can also determine the concentration of proteins bound on ENM surface [105].

However, CE has not yet been used for the determination of proteins in aquatic environments

[106]. The last step of quantification can be replaced by protein concentration assays such as

Bradford [107] or bicinchoninic acid (BCA) assay [108].

Table 2. Flow of methods for the determination of eco-corona biomolecules used in ecotoxicological studies

ENM Organism Separation method Quantification

method

Number of

biomolecules

Reference

Citrate-Au D. magna Disk centrifugation,

SDS-PAGE

LC-MS 175 proteins [109]

(graphene oxide

nanosheets)

GONS

Danio rerio

zebrafish

Centrifugation, SDS-

PAGE

GC-MS,

UV-Vis, Bradford

Protein

concentration 11.3

mg/g

DNA (8.3 mg/kg)

[110]

(polystyrene)

NH2-PS

D. magna Centrifugation, SDS-

PAGE

LC-MS/MS 281 proteins [42]

NH2-PS D. magna Centrifugation, SDS-

PAGE

BCA assay, MS 14 proteins [111]

COOH-PS D. magna Centrifugation, SDS-

PAGE

BCA assay, MS 14 proteins [111]

13

Regarding the identification and quantification of eco-corona formation, a recent review has

reported the main challenges of the centrifugation step such as the bound and unbound fraction

collection (losses due to ENM precipitation, soft and hard corona bound proteins), the small size of

ENMs (need for ultracentrifugation) as well as, the challenges of SDS-PAGE and the number of

proteins bound on ENM surface when EPS were adsorbed onto organic ENMs in the aquatic

environment [112].

4.Exposure conditions that influence eco-corona formation and the effects in

aquatic organisms

4.1 Summary of bio-nano interface

The eco-corona formation, as it was described in section 2, relies on the high surface free energy

and the special characteristics of ENMs (size, shape, charge, surface coatings, rigidity, and

roughness) and the properties of the adsorbed biomolecules (charge, hydrophobicity, molecular

weight, shape). Therefore, all these characteristics determine the eco-nano interactions, the fate,

and the toxicity of ENMs. The corona generally, at the interface, is formed firstly from the most

abundant molecules and then, they are gradually replaced by the highest affinity environmental

molecules. This is explained by the Vroman effect, where the loosely bound biomolecules are

replaced by the tightly bound biomolecules over time [113]. For the environmental proteins, even

if the protein corona is mostly studied in biofluids, it is governed by the kinetics and energetics of

the interaction, and so, the main principles can also be applied to environmental ecosystems [114].

In Figure 2.a, the potential types of eco-nano interaction (left) that correspond to hydrophobic

(due to hydrophobic amino acids), protein-protein, electrostatic (charged amino acids), cation

bridging (salt solution), ligand exchange, hydrogen bonding (ENM surface functionalization) and

chelation (metal ENM) are presented. The displacement by NOM/DOM or other biomolecules (due

to affinity power) and the protein misfolding are some of the possible alterations during eco-

corona formation. The inner cycle (Figure 2.b) (right) depicts the conditions (media, ENM surface

functionalization, biomolecule composition, size or PTMs) that influence the protein-corona in

biofluids, which are also connected with the environmental outer cycle regarding ENMs and

biomolecules, and transformations during eco-corona.

Figure 2. Main interactions between ENM and proteins during eco-corona formation (a) and the conditions influencing the corona formation (b) (from: [115]).

Firstly, the interactions between environmental proteins and nanomaterials can give a rise to

transformations such as misfolding of the protein leading to fibrillation [115], chelation if the ENM

is made of metal, or binding with other molecules while they are already adsorbed on the ENM

surface (NOM, other proteins) [98]. In turn, these transformations may mediate agglomeration,

catalyze the dissolution of ENM, and cause sulfidation or precipitation/sedimentation [34].

Additionally, the unique composition of proteins may have an impact on corona formation and

15

cellular uptake, such as the post-translational modifications (PTMs), in the case of deglycosylation

[116]. PTMs are the modifications of protein structure such as phosphorylation, glycosylation and

lipidation which provide a unique functionality to proteins. It is demonstrated that some proteins

upon interaction may increase the stability of ENMs while others increase their agglomeration

[117]. Moreover, the metal ENMs can react with proteins via photo-oxidation [118] or metal

binding [119], while proteins binding can lead to dissolution and reprecipitation of ENM metal

oxides and inorganic ENMs [120]. Misfolding of proteins results in alteration of their functional

activity and cellular recognition. A study showed that the bovine serum albumin protein (BSA)

interaction with various coated PS ENMs, caused unfolding in different levels. Especially at the

higher unfolding level it was observed that the protein was not recognized by the cell receptor,

lessening the cell uptake [121]. Regarding the protein-NOM competition on the surface of ENMs, it

was observed that when BSA and NOM were present in the solution in mixture, a monolayer was

formed mainly composed by NOM. However when the TiO2 ENMs were exposed to pure BSA, a

multilayer was formed and the protein covered all the free sites of ENM surface [122].

EPS and NOMs in the aquatic environment, interact with ENMs, competing for more hydrophobic

surfaces [66]. The NOM/EPS corona results in a layer of tightly bound molecules on the surface of

ENMs or to aggregates, increasing the ENM hydrodynamic size. This interaction is the primary

interface that determines ENM characteristics such as reactivity (free ions on the surface),

adsorption capacity, charge, and size as well as, the acquired behavior (aggregation, dissolution

and sulfidation). The aggregation and the dissolution depend on the ENM charge and surface

coatings and the conditions of the media as it is described previously. The concentration of NOM

plays a crucial role to sulfidation, since in a study of Liu et al the rate of sulfidation increased by the

increasing NOM presence [123] while in another study the rate increased by the decrease of NOM

amount into the solution [124].

To examine the eco-nano interactions, toxicity tests have been performed in order to

demonstrate the toxicological effects of ENM exposure like genotoxicity, immunotoxicity,

oxidative stress, behavioral alterations as well as lethal effects using different lifeforms ranging

from microbes to algae, invertebrates to fish and in vitro models [38], [45], [125], [126].

Up to date, sixteen standard toxicity tests guidelines for water OECD have been developed as OECD

TG (TG 202, 211, 241, 243), four utilizing invertebrates, nine for fish (TG 203, 236, 210, 212, 215,

229, 230, 234, 240), two for amphibian (TG 231, 241) and one for algae (TG 201) [53]. However, the

ageing of the ENMs prior to corona formation and exposure has not been considered by OECD, as a

factor that should be rationalized in the ecotoxicity testing. Another variable that also influence the

eco-corona formation, and subsequently the toxicological effects, is the exposure incubation and

conditioning time. Some of the previously mentioned toxicological studies have shown that the

eco-corona formation mitigates the adverse effects of ENMs to the aquatic organisms by

promoting disaggregation or homoagglomeration. In this section the main exposure conditions

that affect the eco-corona formation, and potentially, the adverse effects on different organisms

are shown in Table 3.

Table 3. Conditions that influence the eco-corona formation and the effects of ENMs on various aquatic organisms. no: adverse effect; yes: effect because of the applied condition; -: not applied;

Type ENM Size (nm) Biomolecule/NOM Concentration of

biomolecules

ENM incubation

time (hours) Higher pH

value Higher DOC

content Higher Ionic

strength

Surface

functionalization Ageing Species Effects Ref.

Inorganic CeO2 5 Alginate/Chitosan 10 and 100 μg/L 48 - - - - - D. magna Reduced oxidative stress,

reduced CAT activity with

chitosan

[127]

Ce02 14 NOM 0-10 mg C/L 48 yes - no - - P. subcapitata Decreased growth inhibition [128]

ZnO 5 Secreted proteins 0.055 mg/L 0, 12, 24

- - - - - D. magna Higher Toxicity

(EC50 0.193 mg/L) [129]

TiO2 12 EPS 50 mg C/L 24 - - - anatase - Chlorella sp Higher toxicity

(IC50 24.8 mg/L with Cell EPS,

19.9 mg/L without Cell EPS)

[74]

TiO2 5-10 HA 0-1500 mg HA/L 72 - no - anatase - Chlorella sp. Decreased growth inhibition

(IC50 from 4.9 mg/L to 18 mg/L) [130]

TiO2 6 7 NOM 0-4 mg C/L 24 - yes - - - D. magna Decreased acute toxicity

96h-EC50 [131]

Ag/Ag2S 18/44 NOM 4.6 mL/L 0-6 months - - - PVP yes D. magna High survival rates [132]

Ag 6.24 HA 10,20 mg/L 72 yes yes no - - M. aeruginosa Decrease of Ag+ release and

photosynthetic activity [133]

Ag 23.5,22.1 PLFA 5,30 mg/L 0-72 - - - Citrate, BPEI - D. magna 70% lower toxicity [134]

Au 10 SRHA 10 mg/L 24 - - - - - P. subcapitata No membrane damage [68]

Au 25 Secreted proteins TOC:2.1 mg/L 24 - - - Citrate/BSA - D. magna Detoxification due to

disaggregation [109]

Organic PS 220 EPS - 12, 24, 48 - - - COOH yes Chlorella sp. Decrease of ROS, SOD and CAT

activity, and

membrane permeability

[37]

PS 89 Secreted proteins 0-400 μg/mL 1, 3, 6 - yes - NH2 - D. magna Increased toxicity

(0.0258 mg/mL) [111]

PS 200 SRHA

Alginate 2 mg/L - - - - amidine - D. magna

T. platyurus

B. calyciflorus

Increased toxicity

(EC50 150 mg/L,

280 mg/L,160 mg/L)

[43]

PS 60 EPS 10 mg/L 3 - - - COOH - P. tricornutum Reduction of ROS, no toxicity [66]

GONS - Secreted proteins 11.3 mg/g 72 - - - - - D. rerio Membrane damage, toxicity [110]

PS: polystyrene; GONS: graphene oxide nanosheets EPS: extracellular polymeric substances, NOM: natural organic matter, HA: humic acid; EPS: extracellular polymeric substances PLFA: Pony Lake fulvic acid; SRHA: Suwannee River humic acid; TOC: total organic carbon; PVP: polyvinylpyrrolidone; ROS:reactive oxygen species SOD: superoxide dismutase enzyme CAT: chloramphenicol acetyltransferase activity; EC50: median effective concentration; IC50: half maximal inhibitory concentration

4.2 NOM concentration and ENM surface coatings

The OECD has established the minimum NOM concentration on ENM dispersion stability as 10

mg/L to be the most representative of natural aquatic environment. According to Fadare et al the

concentration of HA equal to 10 mg/L is considered sufficient for the crustacean Daphnia magna to

be protected from ENM toxicity, therefore the presence of NOMs should be taken into account

before the conduction of exposure studies [135]. For instance, HA concentrations from 0 to 1500

mg/L used in the algae Chlorella sp. growth test (96h) using different concentrations of anatase

TiO2 ENMs. The results showed that the maximum HA concentration on ENM surface was 50 mg/g

and only at low HA concentrations (<5 mg/L) the algal growth was promoted. At this concentration

of dissolved HA or saturation surface bound HA, the half maximum inhibitory concentration (IC50)

increased from 4.9 to 18 mg/L or 48 mg/L, respectively, indicating toxicity mitigation and a

decrease in ROS production [130]. However, in another study testing using the algae Microcystis

aeruginosa it was observed that low concentrations of HA (<5 mg/L) had little impact on the Ag

ENMs toxicity, while at higher concentrations (10 or 20 mg/L) the algal cells presented higher

photosynthetic activity compared to that of lower concentration [133]. This indicates that the

ENMs should be surface saturated with HA content in order to lessen the toxicity adverse effects.

In a study using D. magna as a model organism, the influence of seven different NOMs (HA,

Suwannee River NOM, Suwannee River HA, Suwannee River FA, seaweed extract, Leonardite, and

Pahokee peat) and their impact on the TiO2 ENMs acute toxicity were examined. As it is already

mentioned, the aromaticity and the composition of NOMs affects the eco-corona formation. In this

case, the concentration of NOM was independent of the NOM type, but their presence decreased

the ENM toxicity (decrease in the 96 h-EC50 values), and it was apparent that the aromaticity and

the hydrophobicity played a crucial role [131]. In addition, applying a model to find the ErC20, the

predicted 48h-ErC20 for the alga Pseudokirchneriella subcapitata was a factor of 1.08-2.57 lower

compared to the experimental due to the derived NOM varied composition [128]. Using Pony Lake

FA (PLFA) the toxicity of Ag ENMs to D. magna decreased to around 70% [134]. Surface plasmon

resonance (SPR) measured the Ag ENMs incubated in the presence and absence of PLFA. At 6 min,

the spherical citrate-Ag ENMs and branched polyethylenimine (BPEI)-Ag ENM in solution showed

two peaks for each coated NP in absence of NOM, while over 3 days of incubation only a slight

decrease in peak intensity of citrate-Ag ENM was observed indicating stability in the presence of

PLFA. In contrast, BPEI-Ag ENM showed peak broadening indicating aggregation in PLFA presence.

CeO2 ENMs on the other hand, coated with alginate displayed increased agglomeration and

sedimentation rates causing ROS release and hyperactive activity to D. magna, in contrast to

chitosan coated CeO2 ENMs which show stability and mediated effects [127].

4.3 ENM ageing

Relatively recent research results have pointed out the need for adaptations in TGs, including the

ageing parameter, especially in long-term exposure, larger than already tested by OECD [136]. The

research investigated the exposure of D. magna to aged and pristine uncoated/coated TiO2 and Ag

ENMs, in the same medium (Class V water, which is a European river water containing NOM [137]).

18

The ageing was achieved by incubating ENM at 1000 mg/L in the Class V water for 6 months stored

in the dark to avoid any other undesired alterations due to the influence of light. Afterwards, by

using TEM imaging, the pristine and aged TiO2 ENMs were observed to be aggregated while

exposed to the medium, while the aged Ag ENMs underwent heteroagglomeration with NOM

molecules. Also, all the daphnid generations after the exposure of first generation to aged TiO2

ENMs showed higher than 70% survival rate, that means it did not affect their fitness and the toxic

effects were reduced. A previous study has shown that the effects of a 6-month aged Ag ENM

dispersion on D. magna, were reduced compared to those of the pristine ENM, in Class V water

[132]. It was observed that in all the media tested (HH combo (organism cultivation media), Class V

water (river water contain NOM)) the hydrodynamic size of aged Ag ENMs was increased and the

survival rates in HH combo media were higher in comparison to the pristine where the daphnids

did not survive after 24 h of exposure. The decreased toxicity of aged coated ENMs is attributed to

the fact that the daphnids ingested aggregated aged ENMs with NOMs, compared to the ingested

pristine ENMs which released Ag+ due to the still active surface sites. It is worth to be mentioned

that this study examined four generations, so when the pristine NPs induced morphological

changes which were epigenetic, so these remained to the next two generations. In consistence

with the above study, L. Natarajan et al examined 12, 24, 48 h time-aged PS ENMs towards marine

microalgae Chlorella sp. indicating enhanced eco-corona formation and reduced adverse effects

[37]. In particular, the size of all the aged COOH-PS ENMs increased significantly after the bio-nano

interaction compared to the pristine one. Also, the 24 h and 48 h aged coated PS NPs showed

increased cell viability while the pristine ENMs appeared to induce increased reactive oxygen

species (ROS) levels and enzyme chloramphenicol acetyltransferase (CAT) activity and decreased

cell viability. Increased levels of ROS can induce membrane damage and lead to cytotoxicity. In this

case, the lessened toxicity of aged PS ENMs is attributed to the fact that the reactive sites of ENM

surface decreased by increased EPS adsorption on their surface. Furthermore, several studies have

observed altered effects and bioavailability of ENMs that underwent much shorter transformed

processes e.g. [47], [48], [138].

Hence, ageing of ENMs is observed to mitigate the toxicity to aquatic organisms, because the

adsorption of biomolecules on their surface leads to the decrease of their active surface sites,

resulting in lower reactivity.

4.4 Conditioning medium and incubation time

Organisms such as D. magna feed by filtering the water for organic matter. However, they might

also release biomolecules to their surrounding environment. This process of biomolecule addition

in the media tested is called ‘’conditioning’’. These secreted proteins can be enzymes, antibodies or

kairomones, changing the concentrations of biomolecules into the environment and increasing the

content of NOM within the water body. The released proteins, depending on their type and origin,

could have high affinity to the surface of ENMs and potentially increase their trend to agglomerate.

By this they may become more available for benthic organisms, that e.g. graze the ground for

feeding like some amphipods or by being more available due to the increased size and size-

depending filtration efficiency of filter- feeding organisms like daphnids or bivalves [139], [140]. For

daphnids it has been described, that due to the morphology of the organs that are used to create

the water flow and sieve out particles from the water, they are only able to ingest particles with

sizes above 200 nm [141]. By this, the increased availability then is linked with an increased toxicity

19

potential [142], [143].

To determine the value of the bioconcentration factor (BCF) of substances in fishes, the exposure

medium should contain less than 5 mg/L organic matter and 2 mg/L total organic carbon, according

to the TG 305 [144]. Although, this concentration is not representative of real conditions and it is

not clear that this TG refers to ENMs. Proteins are also not considered. Mattsson et al tested how

the conditioning of media with secreted proteins mitigate the effects to D. magna comparing

citrate-Au ENMs and BSA-coated Au ENMs. After 24 h of conditioning time, the number of proteins

found on the surface of Au ENM was 175 compared to the protein number present to D. magna in

tap water (without previous conditioning) which was 90. Only 8 proteins were found to be in

common. The protein coated ENMs contributed to ENM detoxification due to their disaggregation

(decrease of aggregate size). Also it was commented that the BSA coating did not affect the

stability of ENMs in the different media as citrate coating did [109]. Nasser et al attributed the

decreased survival of D. magna to the increased protein layer on the NH2- and COOH-PS ENMs

surface when the neonates exposed to a 6-hour conditioning media (higher concentration of

secreted proteins) and 12-hour incubation time. The results showed that the coated NH2-PS were

more toxic compared to the COOH-PS ENMs since more positively charged nanoparticles interact

with the negatively charged membranes of cells and the formation of larger aggregates lead to

higher polydispersity index (PDI) and higher uptake. Using BCA assay, it was observed that in a 6-

hour incubation time around 100 mg/mL of secreted proteins (in total 400 mg/mL) were identified

as hard corona on the amino-functionalized PS ENMs, sized of 50-80 kDa (Type VI secretion system,

stress response protein and QseC sensor protein) [111].

A study using zebrafish embryos observed that exposure to graphene oxide nanosheets (GONS)

incubated in biological secretions for 72 h resulting in graphene oxide biological secreted (GOBS)

with a concentration of protein adsorbed equal to 11.3 mg/g. GOBS eco-corona led to morphology

transformation (thickness and distributed lateral lengths) as well as, higher toxicity than GONS

(death, membrane damage, malformation) due to the great number of functional groups adhered

to the surface of embryo’s cell [110]. Briffa et al. examined how the HH combo media conditioning

and ENM incubation time of 0, 12 and 24 hours influence the eco-corona formation and ENM

stability during tests with D. magna juveniles. They observed that the longer the time of protein

interaction with the coated ZnO ENMs is, the higher the toxicity of ENMs becomes due to changes

of size (aggregation). Opposite results were shown for CeO2 ENMs toxicity, due to no change in size

after incubation (no ENM ingestion from the solution led to no toxicity) [129].

20

5. Conclusions and perspectives

This review has focused on the eco-nano interaction of ENMs with biomolecules (including humic

and fulvic acid, proteins, EPS, lipids) and the conditions that govern such interactions, the

methodologies for characterization and quantification of the formed eco-corona as well as, its role

on ecotoxicological effects. It was observed that the environmental conditions (pH, ionic strength,

salinity) play an important role on the type of eco-nano interaction and the ENM fate, since they

determine the ENM charge and behavior in aquatic media. High pH values and low ionic strength

seem to provide stability to dispersions with NOM through electrostatic repulsions, while some

studies have shown that high ionic strength provides higher interaction with biomolecules and

aggregation. In addition, the ENM properties such as surface coatings and size demonstrated the

same results. Positively charged coated PS ENMs showed increased toxicity compared to the

negatively charged ones and the small ENM size enhanced the eco-nano interaction. Also, it was

observed that usually the high molecular weight EPS and NOM bind more easily than the low

molecular weight and the origin of NOM determines their composition and behavior.

The analytical techniques used for the characterization of eco-corona formation can provide

precise information about the assembly size, the ENM fate in organisms, the NOM composition,

the concentration, the DOM adsorption rate, the affinity potential, the ENM charge and the

assembly morphology. Furthermore, the ENM and DOM identification and quantification has been

achieved by separation methods coupled with mass spectrometry for more accurate results.

Although, these techniques cannot be applied separately and a combination of them is required.

Finally, the eco-nano interface in the aquatic environment was summarized and the parameters of

ENM ageing, NOM concentration, conditioning and incubation time were incorporated to the

ecotoxicological studies and effects. It was observed that bare or pristine ENMs induce high toxicity

to organisms. However, when biomolecules are adsorbed on the ENM surface, their toxicity may be

lessened because of their biotransformation and resulting, decreased reactivity and increased

trend to form aggregates. However, there are also studies that the eco-corona induced higher

toxicity, as reported for graphene oxide causing membrane damage to zebrafish [110]. Therefore,

many conditions may influence the potential of mediated or higher toxicity, and this is challenging

to comprehend.

In order to understand thoroughly the role of eco-corona on toxicity tests, more future studies

should be performed representing more environmentally realistic conditions for the aquatic

environment. Also, future protocols should include the NOM presence. To improve the

understanding of the eco-nano interaction and its impacts on the organisms, future research

should focus on aspects including:

- the environmental conditions: The salinity of water is still a challenging aspect since different

results related to aggregation and dissolution rate of ENMs were observed. Therefore, various salt

concentrations and electrolytes as well as seasonal water variability need to be studied further.

- the ENM and biomolecule properties: The impact of several coatings, ageing, shape is in part

considered. Therefore, there is a need to apply more hydrophobic coatings, various surface

functional groups, ageing by different conditions (light, time) and various ENM crystallinities.

Additionally, the impacts of ENM surface after the isolation of NOM was not determined at all by

the current analytical methods used. The ENM size can also determine the location and the

21

amount of NOM adsorbed on the surface, so further studies should be performed on size

variability for both ENM and biomolecules and its impacts on the organisms.

As it was shown the derivation of biomolecules plays a crucial role on their composition, thus, on

their behavior and eco-corona formation. Studies that implement NOM are referred or isolate

only humic and fulvic acids without considering the proteins or saccharides contained. Therefore,

the studies using analytical methods should focus on the determination of NOM and EPS derived

from different sources as well as the quantification of all the biomolecules contained in natural

water media (seawater, river water).

- SDS PAGE: The current method of separation is not so effective when it comes to the loss of

proteins or when the number of proteins onto the ENM surface is large. Therefore, more studies

are needed to compared result so as to enhance the procedure incorporating LC-MS and CESI-MS

which may provide more accurate results with high resolution.

- Environmental proteins: Plenty of studies have focused on the protein corona in biofluids, but

studies of proteins in the environment is limited. There is a need for more studies, utilizing media

with secreted proteins which are more realistic environments since the eco-corona is formed

dynamically around the surface and conditioning/incubation time influences in a high degree the

concentration of proteins adsorbed. Also, more studies should focus on the composition of these

proteins and their types.

- Hard and soft corona: To date, it is not clear how the molecules arrange onto the ENM surface,

how they assign the chains and what other molecules have higher affinity and replace the NOM or

EPS. Into the environment there are more than one type of biomolecules, thus a mixture of

different types should be examined and their affinity should be characterized further.

22

List of abbreviations 3D-EEM Three dimensional-excitation-emission matrix ACE Affinity capillary electrophoresis AFS Atomic force microscopy BCA Bicinchoninic acid BCF Bioconcentration factor BLI Biolayer interferometry BPEI Branched polyethylenimine BSA Protein bovine serum albumin CAT Chloramphenicol acetyltransferase activity CD Circular dichroism CESI-MS Capillary electrophoresis-electrospray ionization mass

spectrometry CNT Carbon nanotube DLS Dynamic light scattering DNA Deoxyribonucleic acid DOC Dissolved organic carbon DOM Dissolved organic carbon EC50 Median effective concentration ECHA European Chemical Agency ENM Engineered nanomaterials EPS Extracellular polymeric substances ErC50 Median effective concentration in reduction of growth rate FA Fulvic acid FCS Fluorescence correlation spectroscopy FFFF Flow field-flow fractionation FFFF-MALS Flow field-flow fractionation-multi-angle light scattering FI Fluorescence index FT-ICR-MS Fourier transform-ion cyclotron resonance-mass spectrometry FTIR Fourier transform infrared spectroscopy GOBS Graphene oxide biological secreted GONS Graphene oxide nanosheets HA Humic acid HPSEC High-performance size exclusion chromatography HR-LC-MS High resolution-liquid chromatography-mass spectrometry IC50 Half maximal inhibitory concentration ICP-MS Inductively coupled plasma-mass spectrometry ICP-OES Inductively coupled plasma-optical emission spectrometry IS Ionic strength ISO International Organization for Standardization ITC Isothermal titration calorimetry LC-MS Liquid chromatography-mas spectrometry MINBE Minimum Information about Nanomaterial Biocorona

Experiments MRNOM Milwaukee River natural organic matter MW Molecular weight NMR Nuclear magnetic resonance NOM Natural organic matter NTA Nanoparticle tracking analysis OECD Organization for Economic Co-operation and Development PARAFAC Parallel factor analysis PBS phosphate buffered saline PhACS Photothermal absorption correlation spectroscopy PLFA Pony lake fulvic acid POPs Persistent organic pollutants

23

PS Polystyrene PTMs Post-translational modifications PVP Polyvinylpyrrolidone Py-GC-MS Pyrolysis-gas chromatography-mass spectrometry QCM-D Quartz Crystal Microbalance with Dissipation ROS Reactive oxygen species SDS-PAGE Sodium dodecyl sulfate polyacrylamide gel electrophoresis SEM Scanning electron microscopy SEM-EDX Scanning electron microscopy-Energy Dispersive Xray SOD Superoxide dismutase enzyme Sp-ICP-MS Single-particle-Inductively coupled plasma-mass spectrometry SPR Surface plasmon resonance SRHA Suwannee River humic acid TDS-GC-MS Thermal desorption- gas chromatography-mass spectrometry TEM Transmission electron microscopy TG Test guideline TGA Thermal gravimetric analysis TOC Total organic carbon UV-Vis spectroscopy Ultraviolet-Visible spectroscopy XPS X-ray photoelectron spectroscopy YRNOM Yukon river natural organic matter

Acknowledgments The current review was supported by the research grant ENTRANS (Investigating the

ENvironmental impacts of TRANSformed engineered nanomaterials released from wastewater

treatment plants) awarded by the Research Council of Norway (Grant Agreement number

302378/F20)

References

[1] E. Casals, S. Vázquez-Campos, N. G. Bastús, and V. Puntes, “Distribution and potential

toxicity of engineered inorganic nanoparticles and carbon nanostructures in biological systems,” TrAC - Trends Anal. Chem., vol. 27, no. 8, pp. 672–683, Sep. 2008.

[2] E. Navarro et al., “Environmental behavior and ecotoxicity of engineered nanoparticles to algae, plants, and fungi,” Ecotoxicology, vol. 17, pp. 372–386, 2008.

[3] EC, “Scientific Basis for the Definition of the Term ‘ Nanomaterial ,’” Sci. Comm. Emerg. New. Identified Heal. Risks, no. July, pp. 1–43, 2010.

[4] M. E. Vance, T. Kuiken, E. P. Vejerano, S. P. McGinnis, M. F. Hochella, and D. R. Hull, “Nanotechnology in the real world: Redeveloping the nanomaterial consumer products inventory,” Beilstein J. Nanotechnol., vol. 6, no. 1, pp. 1769–1780, 2015.

[5] The Nanodatabase, “The Nanodatabase,” 2021. [Online]. Available: https://nanodb.dk/en/. [6] J. R. Baalousha, M., Lead, “Natural colloids and manufactured nanoparticles in aquatic and

terrestrial systems,” Wilderer,P. (Ed.). Treatise Water Sci. New York, Elsevier, pp. 89–129, 2011.

[7] E. J. Petersen et al., “Identification and avoidance of potential artifacts and misinterpretations in nanomaterial ecotoxicity measurements,” Environmental Science and Technology, vol. 48, no. 8. American Chemical Society, pp. 4226–4246, Apr-2014.

[8] F. A. Cupaioli, F. A. Zucca, D. Boraschi, and L. Zecca, “Engineered nanoparticles. How brain friendly is this new guest?,” Progress in Neurobiology, vol. 119–120. Elsevier Ltd, pp. 20–38, Aug-2014.

[9] Y. C. Sharma, V. Srivastava, V. K. Singh, S. N. Kaul, and C. H. Weng, “Nano-adsorbents for the removal of metallic pollutants from water and wastewater,” Environmental Technology, vol. 30, no. 6. Taylor and Francis Ltd., pp. 583–609, 2009.

[10] F. D. Guerra, M. F. Attia, D. C. Whitehead, and F. Alexis, “Nanotechnology for Environmental Remediation: Materials and Applications,” Molecules, vol. 23, no. 7, p. 1760, 2018.

[11] H. Saroyan, D. Ntagiou, V. Samanidou, and E. Deliyanni, “Modified graphene oxide as manganese oxide support for bisphenol A degradation,” Chemosphere, vol. 225, pp. 524–534, Jun. 2019.

[12] S. E. Ebrahim, A. H. Sulaymon, and H. Saad Alhares, “Competitive removal of Cu2+, Cd2+, Zn2+, and Ni2+ ions onto iron oxide nanoparticles from wastewater,” Desalin. Water Treat., vol. 57, no. 44, pp. 20915–20929, Sep. 2016.

[13] M. Guo, X. Weng, T. Wang, and Z. Chen, “Biosynthesized iron-based nanoparticles used as a heterogeneous catalyst for the removal of 2,4-dichlorophenol,” Sep. Purif. Technol., vol. 175, pp. 222–228, Mar. 2017.

[14] A. Gupta and S. Silver, “Silver as a biocide: Will resistance become a problem?,” Nat Biotechnol., vol. 16, no. 10, p. 888, 1998.

[15] M. Bosetti, A. Mass, E. Tobin, and M. Cannas, “Silver coated materials for external fixation devices: in vitro biocompatibility and genotoxicity,” Biomaterials, vol. 23, no. 3, pp. 887–892, 2002.

[16] S.-Y. Kim, T.-H. Lim, T.-S. Chang, and C.-H. Shin, “Photocatalysis of methylene blue on titanium dioxide nanoparticles synthesized by modified sol-hydrothermal process of TiCl4,” Catal. Letters, vol. 117, pp. 112–118, 2007.

[17] A. Kumar, A. K. Pandey, S. S. Singh, R. Shanker, and A. Dhawan, “Engineered ZnO and TiO2 nanoparticles induce oxidative stress and DNA damage leading to reduced viability of Escherichia coli,” Free Radic. Biol. Med., vol. 51, no. 10, pp. 1872–1881, Nov. 2011.

[18] D. Walczyk, F. B. Bombelli, M. P. Monopoli, I. Lynch, and K. A. Dawson, “What the Cell ‘Sees’ in Bionanoscience,” J. Am. Chem. Soc., vol. 132, no. 16, pp. 5761–5768, 2010.

[19] I. Lynch and K. A. Dawson, “Protein-nanoparticle interactions,” Nano Today, vol. 3, no. 1–2. Elsevier, pp. 40–47, 01-Feb-2008.

25

[20] C. C. Fleischer and C. K. Payne, “Nanoparticle−Cell Interactions: Molecular Structure of the Protein Corona and Cellular Outcomes,” Acc. Chem. Res., vol. 47, no. 8, pp. 2651–2659, 2014.

[21] P. Chun Ke, S. Lin, W. J. Parak, T. P. Davis, and F. Caruso, “A Decade of the Protein Corona,” ACS Nano, vol. 11, no. 12, pp. 11773–11776, 2017.

[22] R. García-Álvarez and M. Vallet-Regí, “Hard and Soft Protein Corona of Nanomaterials: Analysis and Relevance,” Nanomaterials, vol. 11, no. 4, p. 888, 2021.

[23] X. Cui, S. Xu, X. Wang, and C. Chen, “The nano-bio interaction and biomedical applications of carbon nanomaterials,” Carbon, vol. 138. Elsevier Ltd, pp. 436–450, 01-Nov-2018.

[24] Y. Wang, R. Cai, and C. Chen, “The Nano–Bio Interactions of Nanomedicines: Understanding the Biochemical Driving Forces and Redox Reactions,” Acc. Chem. Res., vol. 52, no. 6, pp. 1507–1518, Jun. 2019.

[25] F. Gottschalk and B. Nowack, “The release of engineered nanomaterials to the environment,” J. Environ. Monit., vol. 13, no. 5, pp. 1145–1155, 2011.

[26] M. Bundschuh et al., “Nanoparticles in the environment: where do we come from, where do we go to?,” Environ. Sci. Eur., vol. 30, no. 1, p. 6, Dec. 2018.

[27] P. G. Coble, J. Lead, A. Baker, D. M. Reynolds, and R. G. M. Spencer, Aquatic organic matter fluorescence. Cambridge University Press, 2014.

[28] J. Zumstein and J. Buffle, “Circulation of pedogenic and aquagenic organic matter in an eutrophic lake,” Water Res., vol. 23, no. 2, pp. 229–239, Feb. 1989.

[29] C. É. Line, G. Guen, L. Guo, D. Wang, N. Tanaka, and C.-C. Hung, “Chemical characteristics and origin of dissolved organic matter in the Yukon River,” Biogeochemistry, vol. 77, pp. 139–155, 2006.

[30] P. A. Lozovik, A. K. Morozov, M. B. Zobkov, T. A. Dukhovicheva, and L. A. Osipova, “Allochthonous and autochthonous organic matter in surface waters in Karelia,” Water Resour., vol. 34, no. 2, pp. 204–216, 2007.

[31] W. L. Sun, J. Xia, S. Li, and F. Sun, “Effect of natural organic matter (NOM) on Cu(II) adsorption by multi-walled carbon nanotubes: Relationship with NOM properties,” Chem. Eng. J., vol. 200–202, pp. 627–636, Aug. 2012.

[32] I. Lynch, K. A. Dawson, J. R. Lead, and E. Valsami-Jones, “Macromolecular Coronas and Their Importance in Nanotoxicology and Nanoecotoxicology,” in Frontiers of Nanoscience, vol. 7, Elsevier Ltd, 2014, pp. 127–156.

[33] G. Pulido-Reyes, F. Leganes, F. Fernández-Piñas, and R. Rosal, “Bio-nano interface and environment: A critical review,” Env. Toxicol Chem, vol. 36, no. 12, pp. 3181–3193, 2017.

[34] G. V Lowry, K. B. Gregory, S. C. Apte, and J. R. Lead, “Transformations of Nanomaterials in the Environment,” Environ. Sci. Technol., vol. 46, no. 13, pp. 6893–6899, 2012.

[35] S. Wagner, A. Gondikas, E. Neubauer, T. Hofmann, and F. Von Der Kammer, “Spot the Difference: Engineered and Natural Nanoparticles in the Environment-Release, Behavior, and Fate,” Angew. Chemie Int. Ed., vol. 53, no. 46, pp. 12398–12419, 2014.

[36] S. Liu et al., “The effects and the potential mechanism of environmental transformation of metal nanoparticles on their toxicity in organisms,” Environ. Sci. Nano, vol. 5, p. 2482, 2018.

[37] L. Natarajan et al., “Eco-corona formation lessens the toxic effects of polystyrene nanoplastics towards marine microalgae Chlorella sp.,” Environ. Res., vol. 188, p. 109842, Sep. 2020.

[38] H. T. Ratte, “Bioaccumulation and toxicity of silver compounds: A review,” Environ. Toxicol. Chem., vol. 18, no. 1, pp. 89–108, Jan. 1999.

[39] S. Kuehr, J. Klehm, C. Stehr, M. Menzel, and C. Schlechtriem, “Unravelling the uptake pathway and accumulation of silver from manufactured silver nanoparticles in the freshwater amphipod Hyalella azteca using correlative microscopy,” NanoImpact, vol. 19, p. 100239, Jul. 2020.

[40] S. Kuehr, V. Kosfeld, and C. Schlechtriem, “Bioaccumulation assessment of nanomaterials using freshwater invertebrate species,” Environ. Sci. Eur., vol. 33, no. 1, pp. 1–36, Dec. 2021.

26

[41] O. O. Fadare et al., “Humic acid alleviates the toxicity of polystyrene nanoplastic particles to Daphnia magna ,” Environ. Sci. Nano, vol. 6, p. 1466, 2019.

[42] O. O. Fadare, B. Wan, K. Liu, Y. Yang, L. Zhao, and L.-H. Guo, “Eco-Corona vs Protein Corona: Effects of Humic Substances on Corona Formation and Nanoplastic Particle Toxicity in Daphnia magna,” Environ. Sci. Technol., vol. 54, no. 13, pp. 8001–8009, Jul. 2020.

[43] J. Saavedra, S. Stoll, and V. I. Slaveykova, “Influence of nanoplastic surface charge on eco-corona formation, aggregation and toxicity to freshwater zooplankton,” Environ. Pollut., vol. 252, pp. 715–722, Sep. 2019.

[44] European Parliament Council, “Regulation (EC) No 1907/2006 of the European Parliament and of the Council of 18 December 2006 concerning the Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH), establishing a European Chemicals Agency, amending Directive 1999/4,” Off. J Eu, 2006.

[45] V. Galhano et al., “Impact of wastewater-borne nanoparticles of silver and titanium dioxide on the swimming behaviour and biochemical markers of Daphnia magna: An integrated approach,” Aquat. Toxicol., vol. 220, p. 105404, 2020.

[46] R. Zeumer et al., “Chronic effects of wastewater-borne silver and titanium dioxide nanoparticles on the rainbow trout (Oncorhynchus mykiss),” Sci. Total Environ., vol. 723, 2020.

[47] S. Kühr, S. Schneider, B. Meisterjahn, K. Schlich, K. Hund-Rinke, and C. Schlechtriem, “Silver nanoparticles in sewage treatment plant effluents: chronic effects and accumulation of silver in the freshwater amphipod Hyalella azteca,” Environ. Sci. Eur., vol. 30, no. 1, Feb. 2018.

[48] H. C. Poynton, C. Chen, S. L. Alexander, K. M. Major, B. J. Blalock, and J. M. Unrine, “Enhanced toxicity of environmentally transformed ZnO nanoparticles relative to Zn ions in the epibenthic amphipod Hyalella azteca,” Environ. Sci. Nano, vol. 6, pp. 325–340, 2019.

[49] B. Steinhoff et al., “Investigation of the Fate of Silver and Titanium Dioxide Nanoparticles in Model Wastewater Effluents via Selected Area Electron Diffraction,” Environ. Sci. Technol., vol. 54, no. 14, pp. 8681–8689, 2020.

[50] A. Georgantzopoulou et al., “Ecotoxicological Effects of Transformed Silver and Titanium Dioxide Nanoparticles in the Effluent from a Lab-Scale Wastewater Treatment System,” Environ. Sci. Technol., vol. 52, no. 16, pp. 9431–9441, 2018.

[51] S. M. Hankin et al., “FINAL Specific Advice on Fulfilling Information Requirements for Nanomaterials under REACH (RIP-oN 2) - Final Project Report,” 2011.

[52] P. Christian, F. Von der Kammer, M. Baalousha, and T. Hofmann, “Nanoparticles: structure, properties, preparation and behaviour in environmental media,” Ecotoxicology, vol. 17, no. 5, pp. 326–343, 2008.

[53] OECD, “GUIDANCE DOCUMENT ON AQUATIC AND SEDIMENT TOXICOLOGICAL TESTING OF NANOMATERIALS (No. 317),” Ser. Test. Assess., no. 317, 2020.

[54] ECHA, Chapter R.11: PBT/vPvB assessment, no. June. 2017. [55] Test No. 318: Dispersion Stability of Nanomaterials in Simulated Environmental Media.

OECD, 2017. [56] M. Zhu, H. Wang, A. A. Keller, T. Wang, and F. Li, “The effect of humic acid on the

aggregation of titanium dioxide nanoparticles under different pH and ionic strengths,” Sci. Total Environ., vol. 487, no. 1, pp. 375–380, Jul. 2014.

[57] F. Loosli, P. Le Coustumer, and S. Stoll, “TiO2 nanoparticles aggregation and disaggregation in presence of alginate and Suwannee River humic acids. pH and concentration effects on nanoparticle stability,” Water Res., vol. 47, no. 16, pp. 6052–6063, Oct. 2013.

[58] F. Mohd Omar, H. Abdul Aziz, and S. Stoll, “Aggregation and disaggregation of ZnO nanoparticles: influence of pH and adsorption of Suwannee River humic acid.,” Sci. Total Environ., vol. 468–469, pp. 195–201, Jan. 2014.

[59] M. Baalousha, “Aggregation and disaggregation of iron oxide nanoparticles: Influence of particle concentration, pH and natural organic matter,” Sci. Total Environ., vol. 407, no. 6,

27

pp. 2093–2101, Mar. 2009. [60] K. Li and Y. Chen, “Effect of natural organic matter on the aggregation kinetics of CeO2

nanoparticles in KCl and CaCl2 solutions: Measurements and modeling,” J. Hazard. Mater., vol. 209–210, pp. 264–270, 2012.

[61] D. Lin, S. D. Story, S. L. Walker, Q. Huang, W. Liang, and P. Cai, “Role of pH and ionic strength in the aggregation of TiO2 nanoparticles in the presence of extracellular polymeric substances from Bacillus subtilis,” Environ. Pollut., vol. 228, pp. 35–42, 2017.

[62] M. Erhayem and M. Sohn, “Stability studies for titanium dioxide nanoparticles upon adsorption of Suwannee River humic and fulvic acids and natural organic matter,” Sci. Total Environ., vol. 468–469, pp. 249–257, Jan. 2014.

[63] A. Praetorius et al., “Behavior and bio-Interactions of anthropogenic particles in marine environment for a more realistic ecological risk assessment,” This J. is Open Access Artic., vol. 7, pp. 351–367, 2020.

[64] I. Fernando, D. Lu, and Y. Zhou, “Interactive influence of extracellular polymeric substances (EPS) and electrolytes on the colloidal stability of silver nanoparticles,” Environ. Sci. Nano, vol. 7, no. 1, pp. 186–197, 2020.

[65] X. Li, E. He, K. Jiang, W. J. G. M. Peijnenburg, and H. Qiu, “The crucial role of a protein corona in determining the aggregation kinetics and colloidal stability of polystyrene nanoplastics,” Water Res., vol. 190, p. 116742, Feb. 2021.

[66] G. Grassi et al., “Interplay between extracellular polymeric substances (EPS) from a marine diatom and model nanoplastic through eco-corona formation,” Sci. Total Environ., vol. 725, p. 138457, Jul. 2020.

[67] J. Wu, R. Jiang, W. Lin, and G. Ouyang, “Effect of salinity and humic acid on the aggregation and toxicity of polystyrene nanoplastics with different functional groups and charges,” Environ. Pollut., vol. 245, pp. 836–843, Feb. 2019.

[68] F. Abdolahpur Monikh et al., “Do the joint effects of size, shape and ecocorona influence the attachment and physical eco(cyto)toxicity of nanoparticles to algae?,” Nanotoxicology, vol. 14, no. 3, pp. 310–325, 2020.

[69] L. Liang, L. Luo, and S. Zhang, “Adsorption and desorption of humic and fulvic acids on SiO2 particles at nano-and micro-scales,” Physicochem. Eng. Asp., vol. 384, pp. 126–130, 2011.

[70] Z. Dong, Y. Hou, W. Han, M. Liu, J. Wang, and Y. Qiu, “Protein corona-mediated transport of nanoplastics in seawater-saturated porous media,” Water Res., vol. 182, p. 115978, Sep. 2020.

[71] M. Baalousha, K. Afshinnia, and L. Guo, “Natural organic matter composition determines the molecular nature of silver nanomaterial-NOM corona,” Environ. Sci. Nano, vol. 5, no. 4, pp. 868–881, 2018.

[72] S. M. Louie, E. R. Spielman-Sun, M. J. Small, R. D. Tilton, and G. V Lowry, “Correlation of the Physicochemical Properties of Natural Organic Matter Samples from Different Sources to Their Effects on Gold Nanoparticle Aggregation in Monovalent Electrolyte,” Environ. Sci. Technol, vol. 49, p. 32, 2015.

[73] H. Xu, F. Li, M. Kong, X. Lv, H. Du, and H. Jiang, “Adsorption of cyanobacterial extracellular polymeric substance on colloidal particle: Influence of molecular weight,” Sci. Total Environ., vol. 715, 2020.

[74] X. Gao, K. Zhou, L. Zhang, K. Yang, and D. Lin, “Distinct effects of soluble and bound exopolymeric substances on algal bioaccumulation and toxicity of anatase and rutile TiO2 nanoparticles,” Environ. Sci. Nano, vol. 5, no. 3, pp. 720–729, 2018.

[75] A.-K. Ostermeyer, C. Kostigen, L. Semprini, and T. Radniecki, “Influence of Bovine Serum Albumin and Alginate on Silver Nanoparticle Dissolution and Toxicity to Nitrosomonas europaea,” Environ. Sci. Technol., vol. 47, no. 24, pp. 14403–14410, 2013.

[76] L. Miao et al., “Enhanced stability and dissolution of CuO nanoparticles by extracellular polymeric substances in aqueous environment,” J. Nanoparticle Res., vol. 17, no. 10, p. 404, 2015.

28

[77] C.-S. M. Anaya, J. M. Zhang, S. Spurgin, and J. Chuang, “Effects of Engineered Nanoparticles on the Assembly of Exopolymeric Substances from Phytoplankton,” PLoS One, vol. 6, no. 7, p. 21865, 2011.

[78] Q. V. Ly et al., “Characterization of dissolved organic matter for understanding the adsorption on nanomaterials in aquatic environment: A review,” Chemosphere, vol. 269. Elsevier Ltd, p. 128690, 01-Apr-2021.

[79] K. Rezwan, L. P. Meier, M. Rezwan, J. Vörös, M. Textor, and L. J. Gauckler, “Bovine Serum Albumin Adsorption onto Colloidal Al2O3 Particles: A New Model Based on Zeta Potential and UV-Vis Measurements,” Langmuir, vol. 20, no. 23, pp. 10055–10061, 2004.

[80] T. Maqbool, Q. V. Ly, M. B. Asif, H. Y. Ng, and Z. Zhang, “Fate and role of fluorescence moieties in extracellular polymeric substances during biological wastewater treatment: A review,” Sci. Total Environ., vol. 718, p. 137291, May 2020.

[81] J. Wu, R. Jiang, Q. Liu, and G. Ouyang, “Impact of different modes of adsorption of natural organic matter on the environmental fate of nanoplastics,” Chemosphere, vol. 263, p. 127967, Jan. 2021.

[82] Y. Shimizu, M. Ateia, and C. Yoshimura, “Natural organic matter undergoes different molecular sieving by adsorption on activated carbon and carbon nanotubes,” Chemosphere, vol. 203, pp. 345–352, Jul. 2018.

[83] D. D. Phong and J. Hur, “Using Two-Dimensional Correlation Size Exclusion Chromatography (2D-CoSEC) and EEM-PARAFAC to Explore the Heterogeneous Adsorption Behavior of Humic Substances on Nanoparticles with Respect to Molecular Sizes,” Environ. Sci. Technol., vol. 52, no. 2, pp. 427–435, Jan. 2018.

[84] K. M. James L. Weishaar, George R. Aiken, Brian A. Bergamaschi, Miranda S. Fram, Roger Fujii, “Evaluation of Specific Ultraviolet Absorbance as an Indicator of the Chemical Composition and Reactivity of Dissolved Organic Carbon,” Environ. Sci. Technol., vol. 37, no. 20, pp. 4702–4708, 2003.

[85] C. Lu and F. Su, “Adsorption of natural organic matter by carbon nanotubes,” Sep. Purif. Technol., vol. 58, no. 1, pp. 113–121, 2007.

[86] L. Carlos et al., “The effect of humic acid binding to magnetite nanoparticles on the photogeneration of reactive oxygen species,” Sep. Purif. Technol., vol. 91, pp. 23–29, 2011.

[87] X. Zhang, P. Zhang, Z. Wu, L. Zhang, G. Zeng, and C. Zhou, “Adsorption of methylene blue onto humic acid-coated Fe3O4 nanoparticles,” Physicochem. Eng. Asp., vol. 435, pp. 85–90, 2013.

[88] T. Cedervall et al., “Understanding the nanoparticle-protein corona using methods to quantify exchange rates and affinities of proteins for nanoparticles,” Proc. Natl. Acad. Sci., vol. 104, no. 7, pp. 2050–2055, 2007.

[89] Q. Zhou, S. Ouyang, Z. Ao, J. Sun, G. Liu, and X. Hu, “Integrating Biolayer Interferometry, Atomic Force Microscopy, and Density Functional Theory Calculation Studies on the Affinity between Humic Acid Fractions and Graphene Oxide,” Environ. Sci. Technol., vol. 53, no. 7, pp. 3773–3781, 2019.

[90] B. L. T. Lau, W. C. Hockaday, K. Ikuma, O. Furman, and A. W. Decho, “A preliminary assessment of the interactions between the capping agents of silver nanoparticles and environmental organics,” Colloids Surfaces A Physicochem. Eng. Asp., vol. 435, pp. 22–27, Oct. 2013.

[91] F. Caruso, T. Hyeon, V. Rotello, C. D. Walkey, and W. C. W. Chan, “Understanding and controlling the interaction of nanomaterials with proteins in a physiological environmentw,” Chem. Soc. Rev, vol. 5, pp. 2780–2799, 2011.

[92] Y. Yan et al., “Differential Roles of the Protein Corona in the Cellular Uptake of Nanoporous Polymer Particles by Monocyte and Macrophage Cell Lines,” ACS Nano, vol. 7, no. 12, pp. 10960–10670, 2013.

[93] D. Zhang, B. Pan, R. L. Cook, and B. Xing, “Multi-walled carbon nanotube dispersion by the adsorbed humic acids with different chemical structures,” Environ. Pollut., vol. 196, pp. 292–

29

299, Jan. 2015. [94] R. Zhang, H. Zhang, C. Tu, and Y. Luo, “The limited facilitating effect of dissolved organic

matter extracted from organic wastes on the transport of titanium dioxide nanoparticles in acidic saturated porous media,” Chemosphere, vol. 237, p. 124529, Dec. 2019.

[95] J. D’andrilli, T. Dittmar, B. P. Koch, J. M. Purcell, A. G. Marshall, and W. T. Cooper, “Comprehensive characterization of marine dissolved organic matter by Fourier transform ion cyclotron resonance mass spectrometry with electrospray and atmospheric pressure photoionization,” Rapid Commun. Mass Spectrom., vol. 24, no. 5, pp. 643–650, 2010.

[96] X. Gao, A. Middepogu, R. Deng, J. Liu, Z. Hao, and D. Lin, “Adsorption of extracellular polymeric substances from two microbes by TiO2 nanoparticles,” Sci. Total Environ., vol. 694, p. 133778, Dec. 2019.

[97] H. Cai, E. G. Xu, F. Du, R. Li, J. Liu, and H. Shi, “Analysis of environmental nanoplastics: Progress and challenges,” Chem. Eng. J., vol. 410, p. 128208, Apr. 2021.

[98] J. Ashby, S. Pan, and W. Zhong, “Size and Surface Functionalization of Iron Oxide Nanoparticles Influence the Composition and Dynamic Nature of Their Protein Corona,” ACS Appl. Mater. Interfaces, vol. 6, no. 17, pp. 15412–15419, 2014.

[99] M. Lundqvist, J. Stigler, G. Elia, I. Lynch, T. Cedervall, and K. A. Dawson, “Nanoparticle size and surface properties determine the protein corona with possible implications for biological impacts,” Proc. Natl. Acad. Sci., vol. 105, no. 38, pp. 14265–14270, 2008.

[100] L. Canesi et al., “Interactions of cationic polystyrene nanoparticles with marine bivalve hemocytes in a physiological environment: Role of soluble hemolymph proteins,” Environ. Res., vol. 150, pp. 73–81, Oct. 2016.

[101] M. P. Monopoli, A. S. Pitek, I. Lynch, and K. A. Dawson, “Nanomaterial Interfaces in Biology: Methods and Protocols,” vol. 1025, pp. 137–155, 2013.

[102] D. Docter et al., “Quantitative profiling of the protein coronas that form around nanoparticles,” Nat. Protoc., vol. 9, no. 9, pp. 2030–2044, 2014.

[103] Z. Hu, H. Zhang, Y. Zhang, R. Wu, and H. Zou, “Nanoparticle size matters in the formation of plasma protein coronas on Fe3O4 nanoparticles,” Colloids Surfaces B Biointerfaces, vol. 121, pp. 354–361, Sep. 2014.

[104] A. J. Chetwynd, K. E. Wheeler, and I. Lynch, “Best practice in reporting corona studies: Minimum information about Nanomaterial Biocorona Experiments (MINBE),” Nano Today, vol. 28, p. 100758, Oct. 2019.

[105] K. Faserl et al., “Corona Isolation Method Matters: Capillary Electrophoresis Mass Spectrometry Based Comparison of Protein Corona Compositions Following On-Particle versus In-Solution or In-Gel Digestion,” Nanomaterials, vol. 9, p. 898, 2019.

[106] A. J. Chetwynd et al., “Capillary electrophoresis mass spectrometry approaches for characterization of the protein and metabolite corona acquired by nanomaterials,” J. Vis. Exp., vol. 2020, no. 164, 2020.

[107] N. J. Kruger, “The Bradford Method For Protein Quantitation BT - The Protein Protocols Handbook,” J. M. Walker, Ed. Totowa, NJ: Humana Press, 2009, pp. 17–24.

[108] J. M. Walker, “The Bicinchoninic Acid (BCA) Assay for Protein Quantitation BT - The Protein Protocols Handbook,” J. M. Walker, Ed. Totowa, NJ: Humana Press, 2009, pp. 11–15.

[109] K. Mattsson et al., “Disaggregation of gold nanoparticles by Daphnia magna,” Nanotoxicology, vol. 12, no. 8, pp. 885–900, 2018.

[110] L. Mu, Y. Gao, and X. Hu, “Characterization of Biological Secretions Binding to Graphene Oxide in Water and the Specific Toxicological Mechanisms,” Environ. Sci. Technol, vol. 50, p. 55, 2016.

[111] F. Nasser and I. Lynch, “Secreted protein eco-corona mediates uptake and impacts of polystyrene nanoparticles on Daphnia magna,” J. Proteomics, vol. 137, pp. 45–51, Mar. 2016.

[112] M. Junaid and J. Wang, “Interaction of nanoplastics with extracellular polymeric substances (EPS) in the aquatic environment: A special reference to eco-corona formation and

30

associated impacts,” Water Research, vol. 201. Elsevier Ltd, p. 117319, 01-Aug-2021. [113] L. Vroman, A. L. Adams, G. C. Fischer, and P. C. Munoz, “Interaction of high molecular

weight kininogen, factor XII, and fibrinogen in plasma at interfaces.,” Blood, vol. 55, no. 1, pp. 156–159, Jan. 1980.

[114] K. E. Wheeler et al., “Environmental dimensions of the protein corona,” Nat. Nanotechnol., vol. 16, pp. 617–629, 2021.

[115] J. Li, I. Pylypchuk, D. P. Johansson, V. G. Kessler, G. A. Seisenbaeva, and M. Langton, “Self-assembly of plant protein fibrils interacting with superparamagnetic iron oxide nanoparticles,” Sci. Rep., vol. 9, p. 8939, 2019.

[116] A. Ghazaryan, K. Landfester, and V. Mailänder, “Protein deglycosylation can drastically affect the cellular uptake †,” Nanoscale, vol. 11, no. 22, pp. 10727–10737, 2019.

[117] J. S. Gebauer et al., “Impact of the Nanoparticle−Protein Corona on Colloidal Stability and Protein Structure,” Langmuir, vol. 28, no. 25, pp. 9673–9679, 2012.

[118] D. T. Jayaram, S. Runa, M. L. Kemp, and C. K. Payne, “Nanoparticle-induced oxidation of corona proteins initiates an oxidative stress response in cells †,” Nanoscale, vol. 9, no. 22, pp. 7595–7601, 2017.

[119] A. J. Martinolich, G. Park, M. Y. Nakamoto, R. E. Gate, and K. E. Wheeler, “Structural and Functional Effects of Cu Metalloprotein-Driven Silver Nanoparticle Dissolution,” Environ. Sci. Technol., vol. 46, no. 11, pp. 6355–6362, 2012.

[120] C. Xie et al., “Bacillus subtilis causes dissolution of ceria nanoparticles at the nano-bio interface †,” Environ. Sci. Nano, vol. 6, p. 216, 2019.

[121] A. Albanese et al., “Secreted Biomolecules Alter the Biological Identity and Cellular Interactions of Nanoparticles,” ACS Nano, vol. 8, no. 6, pp. 5515–26, 2014.

[122] S. Shakiba, A. Hakimian, L. R. Barco, and S. M. Louie, “Dynamic Intermolecular Interactions Control Adsorption from Mixtures of Natural Organic Matter and Protein onto Titanium Dioxide Nanoparticles,” Environ. Sci. Technol, vol. 52, p. 14, 2018.

[123] J. Liu, K. G. Pennell, and R. H. Hurt, “Kinetics and Mechanisms of Nanosilver Oxysulfidation,” Environ. Sci. Technol, vol. 45, pp. 7345–7353, 2011.

[124] Y. Zhang, J. Xia, Y. Liu, L. Qiang, and L. Zhu, “Impacts of Morphology, Natural Organic Matter, Cations, and Ionic Strength on Sulfidation of Silver Nanowires,” Environ. Sci. Technol., vol. 50, no. 24, pp. 13283–13290, 2016.

[125] T. Gomes, J. P. Pinheiro, I. Cancio, C. G. Pereira, C. Cardoso, and M. J. Bebianno, “Effects of copper nanoparticles exposure in the mussel Mytilus galloprovincialis,” Environ. Sci. Technol., vol. 45, no. 21, pp. 9356–9362, 2011.

[126] T. L. Rocha et al., “Immunocytotoxicity, cytogenotoxicity and genotoxicity of cadmium-based quantum dots in the marine mussel Mytilus galloprovincialis,” Mar. Environ. Res., vol. 101, pp. 29–37, 2014.

[127] S. Villa et al., “Natural molecule coatings modify the fate of cerium dioxide nanoparticles in water and their ecotoxicity to Daphnia magna,” Environ. Pollut., vol. 257, p. 113597, 2020.

[128] K. Van Hoecke, K. A. C. De Schamphelaere, P. Van Der Meeren, G. Smagghe, and C. R. Janssen, “Aggregation and ecotoxicity of CeO2 nanoparticles in synthetic and natural waters with variable pH, organic matter concentration and ionic strength,” Environ. Pollut., vol. 159, no. 4, pp. 970–976, 2011.

[129] S. M. Briffa, F. Nasser, E. Valsami-Jones, and I. Lynch, “Uptake and impacts of polyvinylpyrrolidone (PVP) capped metal oxide nanoparticles on Daphnia magna: role of core composition and acquired corona †,” Environ. Sci. Nano, vol. 5, p. 1745, 2018.

[130] D. Lin, J. Ji, Z. Long, K. Yang, and F. Wu, “The influence of dissolved and surface-bound humic acid on the toxicity of TiO2 nanoparticles to Chlorella sp.,” Water Res., vol. 46, no. 14, pp. 4477–4487, 2012.

[131] F. Seitz et al., “Quantity and quality of natural organic matter influence the ecotoxicity of titanium dioxide nanoparticles,” Nanotoxicology, vol. 10, no. 10, pp. 1415–1421, 2016.

[132] L.-J. A. Ellis et al., “Multigenerational Exposures of Daphnia Magna to Pristine and Aged

31

Silver Nanoparticles: Epigenetic Changes and Phenotypical Ageing Related Effects,” Small, vol. 16, no. 21, 2020.

[133] T. Huang, M. Sui, X. Yan, X. Zhang, and Z. Yuan, “Anti-algae efficacy of silver nanoparticles to Microcystis aeruginosa: Influence of NOM, divalent cations, and pH,” Colloids Surfaces A Physicochem. Eng. Asp., vol. 509, pp. 492–503, 2016.

[134] Y. Jung, G. Metreveli, C.-B. Park, S. Baik, and G. E. Schaumann, “Implications of Pony Lake Fulvic Acid for the Aggregation and Dissolution of Oppositely Charged Surface-Coated Silver Nanoparticles and Their Ecotoxicological Effects on Daphnia magna,” Environ. Sci. Technol., vol. 52, no. 2, pp. 436–445, Jan. 2018.

[135] O. O. Fadare, B. Wan, L.-H. Guo, Y. Xin, W. Qin, and Y. Yang, “Humic acid alleviates the toxicity of polystyrene nanoplastic particles to Daphnia magna,” Environ. Sci. Nano, vol. 6, no. 5, pp. 1466–1477, 2019.

[136] L.-J. A. Ellis, E. Valsami-Jones, and I. Lynch, “Exposure medium and particle ageing moderate the toxicological effects of nanomaterials to Daphnia magna over multiple generations: a case for standard test review? †,” Environ. Sci. Nano, vol. 7, p. 1136, 2020.

[137] J. Hammes, J. A. Gallego-Urrea, and M. Hassellöv, “Geographically distributed classification of surface water chemical parameters influencing fate and behavior of nanoparticles and colloid facilitated contaminant transport,” Water Res., vol. 47, no. 14, pp. 5350–5361, Sep. 2013.

[138] R. Zeumer, L. Hermsen, R. Kaegi, S. Kühr, B. Knopf, and C. Schlechtriem, “Bioavailability of silver from wastewater and planktonic food borne silver nanoparticles in the rainbow trout Oncorhynchus mykiss,” Sci. Total Environ., vol. 706, p. 135695, Nov. 2020.

[139] M. S. Othman and D. Pascoe, “Growth, development and reproduction of Hyalella azteca (Saussure, 1858) in laboratory culture,” Crustaceana, vol. 74, no. 2, pp. 171–181, Feb. 2001.

[140] C. A. García-Negrete et al., “Behaviour of Au-citrate nanoparticles in seawater and accumulation in bivalves at environmentally relevant concentrations,” Environ. Pollut., vol. 174, pp. 134–141, Mar. 2013.

[141] W. Geller and H. Müller, “The filtration apparatus of Cladocera: filter mesh-sizes and their implications on food selectivity,” Oecologia, vol. 49, no. 3, pp. 316–321, 1981.

[142] L.-J. A. Ellis and I. Lynch, “Mechanistic insights into toxicity pathways induced by nanomaterials in Daphnia magna from analysis of the composition of the acquired protein corona,” Environ. Sci. Nano, vol. 7, no. 11, p. 3343, 2020.

[143] Y. Hayashi et al., “Species Differences Take Shape at Nanoparticles: Protein Corona Made of the Native Repertoire Assists Cellular Interaction,” Environ. Sci. Technol., vol. 47, no. 24, pp. 14367–14375, Dec. 2013.

[144] OECD, Test No. 305: Bioaccumulation in Fish: Aqueous and Dietary Exposure. 2012.