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STOTEN-24347; No of Pages 14
Science of the Total Environment xxx (2017) xxx–xxx
Contents lists available at ScienceDirect
Science of the Total Environment
journal homepage: www.elsevier.com/locate/scitotenv
Review
Silver engineered nanoparticles in freshwater systems – Likely fate and
behaviour through natural attenuation processes
David Shevlin ⁎, Niall O'Brien, Enda Cummins
School of Biosystems and Food Engineering, University College Dublin, Belfield, Dublin 4, Ireland
H I G H L I G H T S
G R A P H I C A L
A B S T R A C T
• Natural attenuation processes influence
fate and behaviour in aquatic systems.
• Colloid interactions likely to dominate
and influence hetero-aggregation potential.
• Significant
challenges
detecting
nanosilver/ENMs following surface
modifications
• Complexity of freshwater aquatic systems varies fate and behaviour of
nanosilver.
• Environmental
persistence
of
nanosilver likely, due to elevated use
and release.
a r t i c l e
i n f o
Article history:
Received 4 July 2017
Received in revised form 12 October 2017
Accepted 13 October 2017
Available online xxxx
Editor: D. Barcelo
Keywords:
Nanosilver
Toxicity
Aquatic
Health
Environmental
a b s t r a c t
Growth in the nanotechnology sector is likely introducing unnatural formations of materials on the nanoscale
(10−9 m) to the environment. Disposal and degradation of products incorporating engineered nanomaterials
(ENMs) are likely being released into natural aquatic systems un-intentionally primarily via waste water effluents. The fate and behaviour of metallic based nanoparticles (NPs) such as silver (Ag) in aquatic waters is complex with high levels of variability and uncertainty. In-situ physical, biological and chemical (natural
attenuation) processes are likely to influence ENM fate and behaviour in freshwater systems. Surfaced functionalized particles may inhibit or limit environmental transformations which influence particle aggregation, mobility, dissolution and eco-toxic potential. This paper focuses on ENM characteristics and the influence of physical,
chemical and biological processes occurring in aquatic systems that are likely to impact metallic ENMs fate. A
focus on silver NPs (while for comparison, reporting about other metallic ENMs as appropriate) released to
aquatic systems is discussed relating to their likely fate and behaviour in this dynamic and complex environment.
This paper further highlights the need for specific risk assessment approaches for metallic ENMs and puts this
into context with regard to informing environmental policy and potential NP influence on environmental/
human health.
© 2017 Elsevier B.V. All rights reserved.
⁎ Corresponding author.
E-mail addresses: david.shevlin@ucdconnect.ie (D. Shevlin), niall.obrien@ucd.ie (N. O'Brien), enda.cummins@ucd.ie (E. Cummins).
https://doi.org/10.1016/j.scitotenv.2017.10.123
0048-9697/© 2017 Elsevier B.V. All rights reserved.
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
2
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
Contents
1.
2.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Nanomaterial characterization . . . . . . . . . . . . . . . . . . . . . . . . . . .
2.1.
Nanoparticle fate, behaviour linking to risk assessment . . . . . . . . . . . . .
2.2.
Particle surface stabilization coatings . . . . . . . . . . . . . . . . . . . . .
2.3.
Environmental risk assessment of ENMs . . . . . . . . . . . . . . . . . . .
3.
Engineered nanoparticle release to the environment . . . . . . . . . . . . . . . . .
3.1.
ENMs in the aquatic environment . . . . . . . . . . . . . . . . . . . . . .
4.
Fate and behaviour in aquatic systems. . . . . . . . . . . . . . . . . . . . . . . .
5.
Physical processes likely to influence fate and behaviour of metallic ENMs . . . . . . .
5.1.
Dispersion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.2.
Agglomeration and aggregation . . . . . . . . . . . . . . . . . . . . . . .
5.3.
Hetero-aggregation . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.4.
Adsorption. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
5.5.
Partitioning of nanomaterials . . . . . . . . . . . . . . . . . . . . . . . .
5.6.
Sedimentation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.
Chemical and biological processes likely to influence fate and behaviour of metallic ENMs
6.1.
Photochemical . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.2.
Redox reactions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.3.
Dissolution. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
6.4.
Precipitation and speciation . . . . . . . . . . . . . . . . . . . . . . . . .
6.5.
Bio-transformations. . . . . . . . . . . . . . . . . . . . . . . . . . . . .
7.
Environmental concerns of ENMs . . . . . . . . . . . . . . . . . . . . . . . . . .
8.
Nanometrology . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
9.
Modelling nanomaterial environmental fate in aquatic systems . . . . . . . . . . . .
10.
Human health concerns . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
11.
Quantification and risk assessment modelling. . . . . . . . . . . . . . . . . . . .
12.
Policy developments and legislation . . . . . . . . . . . . . . . . . . . . . . . .
13.
Knowledge and data gaps . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
14.
Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
References. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
1. Introduction
The natural environment is persistently being exposed to contaminants originating from anthropogenic sources. The presence of heavy
metal residues originating from these sources could present ecological
health concerns for the environment. Biologically non-essential elemental metals such as silver (Ag) could present toxicity concerns due
to unnatural levels accumulating in the environment (O'Brien and
Cummins, 2008). Furthermore, metallic oxides such as Ag are being utilized in the emerging field of Nanotechnology. Growth in the nanotechnology sector is likely introducing unnatural formations of materials of
nanoscale (10−9 m) to the environment. Nanotechnology involves the
synthesis and manipulation of materials on the nano-scale for a diverse
range of nano-enabled products and processes (Liu et al., 2014; Brar
et al., 2010). Advancements in particle characterization techniques led
industry to identify commercially beneficial properties of materials at
the nano-scale (Buzea et al., 2007; Petersen et al., 2014;
Kunhikrishnan et al., 2015). Materials at the nanoscale can possess enhanced properties, with superior physicochemical properties, in contrast to bulk material forms. These enhanced properties are
attributable to their small size (surface area and size distribution).
Chemical composition (purity, crystallinity, electronic properties, etc.),
surface structure (surface reactivity, surface groups, inorganic or organic coatings, etc.), solubility, shape, and aggregation ability which all contribute to the activity of NMs (Nel, 2007). Acknowledging that these
materials possess unique properties, concerns have been growing as
to the acute and chronic health effects that increasing levels of metallic
engineered nanomaterials (ENMs) such as Ag released to the environment may pose in the immediate too long term. Evidence is scarce as
to how these unnatural materials with highly reactive surface properties will react in the natural environment and the likely risks they may
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pose. Therefore, research focus has somewhat shifted to assess the potential risks of ENMs (Meesters et al., 2014; Quik et al., 2014).
The objective of this review paper was to identify and review the current available literature relating to the fate and behaviour of metallic
ENM's with a focus on Ag, while for comparison, reporting data and discussion about other metallic ENMs as appropriate. Silver ENMs are
highlighted because of their bactericidal properties and use in disposable
consumer products and likely release to the environment during use and
disposal. Engineered nAg entering wastewater networks could release
AgNPs in wastewater effluents discharges to freshwater environments
and could pose a potential risk to environmental/human health. The review will focus on the environmental processes that nAg (as a representative metallic NP) may experience upon entry to primarily freshwater
aquatic environments in an effort to assess ENM transport, fate and
behaviour in complex and variable natural environments.
2. Nanomaterial characterization
Nanomaterials are generally engineered in three different dimensional forms: nano-films and coatings (nano 1D), nano-tubes, nanofibres and nano-wires (nano 2D), and nanoparticles (nano 3D) (Wang
et al., 2015c). Nanoparticles (NPs) and materials are categorized by
the European Commission (2015) as:
A natural, incidental or manufactured material containing particles, in
an unbound state or as an aggregate or as an agglomerate and where,
for 50% or more of the particles in the number size distribution, one or
more external dimensions is in the size range 1 nm–100 nm.
The distinctive physical and chemical properties of ENMs may pose
toxicity concerns for the environment and human health due to their
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
enhanced physiochemical properties (Nel, 2007). Furthermore, surface
reactivity increases due to the large surface area to volume ratio, thus
allowing greater numbers of surface atoms to freely interact with
other atoms in close proximity. These surface related properties enhance the commercially beneficial chemical (catalysts), biological
(Ag), magnetic (medical), electronic (quantum dots) and optical abilities (TiO2) of ENMs (Liu et al., 2014).
ENMs are incorporated in a diverse range of consumable goods such
as textiles, household cleaning products, personal care products and
medical applications (Brar et al., 2010; Buzea et al., 2007). Production
of ENMs can be derived in two different ways: one approach is mechanical reduction of bulk materials through milling, repeated quenching,
laser abolition or photolithography - processes known as top-down
(Kunhikrishnan et al., 2015; Mafune et al., 2000). Secondly, a process referred to as bottom-up with the manipulation of molecular components
into complex clusters resulting in the formation of desired materials
(Kunhikrishnan et al., 2015).
Nanoparticles are incorporated as surface coatings or as an embedded structural component in products (N1600 consumer products containing NMs) (PEN, 2016). The growth of the nanotechnology sector
and increasing use of metallic-based ENMs may lead to significant levels
of metallic based NP contamination in all environmental compartments
(Barton et al., 2015a). Concerns about the volumes (500 tons/year of Ag
NMs) (Gottschalk et al., 2009; Sun et al., 2014) being produced globally
and potentially entering the various environmental compartments over
their life cycle have hastened the need to evaluate the associated environmental risk (Brunetti et al., 2015). Silver (14.5% of available products) and titanium (37% of available products) ENMs represent the
majority of nano-functionalized products commercially available. Furthermore, many products may contain a number of different nanocomposite combinations in varying concentrations (Vance et al., 2015). The
fate and behaviour of metallic AgNPs and other ENMs entering into
aquatic systems could influence their potential toxicity to biota and
the ecosystem as a whole (Rocha et al., 2015).
likely interactions of coated particles with humic substances in freshwater environments and their entrapment in formation of heteroaggregates (Wang et al., 2015a). While toxicity of stabilizing coatings
such as PVP have been shown to be toxic to marine microalgae under
experimental conditions (Schiavo et al., 2017). The omni presence of
cysteine in freshwater systems is assumed to suppress the toxicity effects of PVP coatings (Navarro et al., 2015).
Electrostatic stability is achieved through the application of negatively charged Cit coatings, while PVP coatings present steric stabilization (Topuz et al., 2014). Particle surface charge is determined by
measuring the zeta potential using a zeta analyser which indicates the
magnitude of electrostatic charge (Montano et al., 2014; Rocha et al.,
2015). Zeta potential is an effective tool for measuring coagulation potential as it measures the level of repulsion/attraction between colloids
which influences formation of aggregates.
2.3. Environmental risk assessment of ENMs
Metallic based NPs may pose a health concern due to possible toxicity to aquatic biota and a direct route for entry into the food chain
(Buzea et al., 2007). Therefore it would seem prudent to access the
risks they may pose to the environment and human health. While
AgNPs can be found naturally in the environment (Wagner et al.,
2014), the increasing use and eventual disposal of engineered nAg will
elevate concentrations beyond natural levels. Release rates, potential
routes/pathways through the environment, physiochemical form,
transformation potential, fate and behaviour upon entry to aquatic systems remain uncertain. These are vital factors in determining the predicted environmental concentration (PEC) and predicted no effect
concentration (PNEC). These PEC and PNEC are essential for reliable environmental risk assessment modelling (Praetorius et al., 2012). Current
methods for chemical risk assessment are calculated using a risk quotient (RQ) using the PEC of a substance and the PNEC of that same chemical substance using the following formula:
2.1. Nanoparticle fate, behaviour linking to risk assessment
RQ ¼
Key reviews conducted by Hotze et al. (2010), Schaumann et al.
(2015), Wang et al. (2015a) and Vale et al. (2016) have focused on
metal and metal oxides in aquatic environments focusing on factors
influencing NP fate and behaviour. Other reviews have focused on
ENMs in saltwater environments (Minetto et al., 2016). These reviews
highlight the many factors that influence the behaviour of ENMs in natural waters such as interactions with organic matter (OM) and the
destabilizing influence of increasing electrolyte concentrations. It is postulated that metallic ENMs could undergo physical and chemical processes in natural waters similar to other contaminants: dispersion,
agglomeration/aggregation, oxidation, sulfidation, dissolution and sedimentation (Kunhikrishnan et al., 2015; Nowack et al., 2012; Louie et al.,
2014).
2.2. Particle surface stabilization coatings
Stabilization of ENMs during production and storage is achieved
through application of surface coatings to prevent aggregation (homoaggregation) (Baker et al., 2014). Surface coatings such as polyvinylpyrrolidone (PVP) (López-Serrano et al., 2014), citrate (Cit) (Navarro et al.,
2014) and polyethylene glycol (PEG) (Hartmann et al., 2014) can also
have a functional purpose by enhancing the performance of particles
by altering the reactivity (charge repulsion) or attachment of surface
molecules (functionalization) (Montano et al., 2014). These coatings referred to above, are frequently applied to AgNPs to enhance particle stability but invariably also may inhibit/limit their agglomeration/
aggregation potential in natural waters, increasing particle mobility
and dispersal potential (Badawy et al., 2010). Toxicity potential of stabilization coatings may be somewhat unclear and mostly mitigated by the
3
PEC
PNEC
ð1Þ
In calculating a risk level using the RQ, values N1 indicate a risk and
cause for concern while those below and closer to 0 indicating no risk
concern (Arvidsson et al., 2011). Gottschalk et al. (2010) used the RQ
formula in their study of nAg and TiO2, they observed values N1 in treated wastewater but values were much lower for other environmental
compartments (soil, water, and air). These values were modelled on release rates in the year 2008 so one could expect increased concentrations if evaluated again.
3. Engineered nanoparticle release to the environment
Point and non-point sources of ENM exposure to various environmental compartments are mainly derived from landfill leachate (degradation of nano-incorporated materials), waste water treatment plants
(WWTPs), agriculture (bio-solids in land application) and industry
(smoke, cooling systems) (Nowack et al., 2012; Barton et al., 2015a,
2015b; Dale et al., 2015a, 2015b; Gottschalk and Nowack, 2011).
ENMs are utilized in environmental remediation and water treatment
processes for their absorption capacity and high affinity for contaminants which could also contribute to direct environmental release
(Adeleye et al., 2013; Yang et al., 2013; Gehrke et al., 2015). Removal
of ENPs during water treatment could be hindered by the operational
efficiency of conventional waste water treatment processes (chemical,
physical, and biological). Treatments may be ineffective at removing/recovering or capturing ENMs due to NP intrinsic properties (sizes, shape,
core composition, surface properties, and concentration) potentially
resulting in direct release in effluent discharges (Liu et al., 2014).
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
4
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
3.1. ENMs in the aquatic environment
The PEC/PNEC risk quotient used to quantify contaminant exposure
is difficult to calculate for ENMs as reputable data on the environmental
aquatic loading are not adequately evaluated and predictions are mainly
based on probability distributions (Gottschalk et al., 2013; O'Brien and
Cummins, 2011). In the absence of actual, measured aquatic ENM concentrations and fate data, it is essential to develop effective fate models
to determine ENM transport, partitioning, degradation and transformation (Bour et al., 2015). In these environmental fate models, particle
characteristics such as particle charge; size; morphology; coating and
chemical composition will be of critical importance (OECD, 2014).
The European Chemicals Agency (ECHA) provides guidance on Registration, Evaluation, Authorisation and Restriction of Chemicals
(REACH) regulation guidelines for a chemical substance entering a
water body. The concentration in the water can be calculated (Eq. (2))
using the different fate processes that are likely to impact on that substance under environmentally relevant conditions (Quik et al., 2015).
Rate constants must be measured or estimated to calculate a steadystate concentration (Cw) of a substance dissolved in water (Quik et al.,
2011)
Cw ¼
E
Kadv þ Kvol þ Kdeg þ Ksed V
ð2Þ
where:
Cw = concentration in water, E = balance of an emission (kg s−1),
Kadv = advection, Kvol = volatilization, Kdeg = degradation/transformation, Ksed = deposition to sediments, V = volume of water body m3.
This approach may be somewhat limited in the assessment of metallic ENMs as they may not undergo the same fate processes due to their
complex surface properties and nanospecific behaviour (Quik et al.,
2015). Most notably, data on measured removal rate constants for metallic ENMs are scarce and while some parameters may not be relevant
such as “volatilization”, in the case of this equation, they can simply be
omitted (Quik et al., 2011).
4. Fate and behaviour in aquatic systems
Natural formation of AgNPs can occur in the aquatic environment
through the precipitation of ionic Ag species. Although interactions
with NOM can make it difficult to distinguish them from natural or anthropogenic origins due to surface modifications (Li et al., 2016) Transformation or surface modifications of AgNPs could occur within natural
waters through in-situ natural attenuation processes. Natural attenuation is defined by the Environmental Protection Agency (EPA) as:
“A variety of physical, chemical, or biological processes that, under
favourable conditions, act without human intervention to reduce the
mass, toxicity, mobility, volume, or concentration of contaminants in
soil or groundwater. These in-situ processes include biodegradation;
dispersion; dilution; sorption; volatilization; radioactive decay; and
chemical or biological stabilization, transformation, or destruction of
contaminants”.
[(Newell et al., 2002).]
Transformational processes are dependent on material surface, environmental characteristics, presence of natural substances (bio and geocolloids), environmental conditions (ionic strength, pH, and temperature), and seasonal variations (Wang et al., 2015a, 2015b; Dale et al.,
2015a, 2015b; Louie et al., 2014). In oxygen rich waters AgNPs are likely
to be oxidized releasing particle surface ions (Ag+) (Garner and Keller,
2014). Where particles form aggregates, gravitational settling should
occur when aggregates reach critical mass and migrate through the
water column eventually settling in the sediments. Sulphur (S) in the
sediments can bind with depositing AgNPs. Complexing and
transforming Ag with S should form Ag2S under anoxic conditions rendering particles chemically stable and unreactive (Dwivedi et al., 2015).
Although, setting particles may be bioavailable to sediment dwelling organisms and a possible entry route to the aquatic trophic system (Ellis
et al., 2016). Also, bio uptake, ingestion, and eventual excretion by organisms may modify particle surface properties through biological processes by the removal of stabilizing organic or polymeric coatings
(Schaumann et al., 2015)
Particle physiochemical characteristics and water chemistry (e.g.
pH, ionic strength, dissolved OM content) have been identified as determining factors in ENM mobility, fate, and behaviour in receiving waters
(Boxall et al., 2007; Park et al., 2014). The following section considers
the physical (Table 1.), chemical and biological (Table 2.) processes
that are likely to impact on nAg and other metallic ENMs in natural
waters.
5. Physical processes likely to influence fate and behaviour of metallic ENMs
5.1. Dispersion
Particle dispersion in aqueous media is influenced by kinetic processes through Brownian motion, while particle characteristics (surface
properties, density and mass) and physiochemical conditions of receiving media (pH, ionic strength, and OM) also contribute to dispersion efficiency (Table 1) (Römer et al., 2011). Surface stabilized metallic ENMs
could persistence and be subjected to wide spread dispersal in natural
waters (Yang et al., 2014). However, widespread dispersal should dilute
ENM concentrations and result in a lower toxicity potential within
aquatic systems (Hartmann et al., 2014; Li et al., 2011). Electrolyte concentration in the media can influence dispersion potential as the presence of mono and di-valent ions can significantly impact on
dispersability (Chinnapongse et al., 2011).
5.2. Agglomeration and aggregation
Agglomeration is a mass-conserving process whereby particles of
similar properties interact to form larger clusters held together by attractive van der Waals forces (Hartmann et al., 2014). In colloid science,
the processes of agglomeration and aggregation are sometimes used interchangeably to mean the same thing. Aggregation is essentially a nonreversible process resulting in increased particle clusters as particles adhere to one another by strong chemical (covalent or ionic bonds) or
electrostatic interactions (Hartmann et al., 2014; Garner and Keller,
2014). In natural waters, homo-aggregation is unlikely to control particle – particle collisions as natural colloids are typically present in concentrations several orders of magnitude higher than NPs (Sun et al.,
2014). Within complex media such as river/stream systems, AgNPs
are likely to interact with a variety of bio and geo-colloids of varying
concentrations which could substantially modify particle surface
chemistry.
Collision frequencies between particles leading to aggregate formation are influenced by Brownian motion (perikinetic aggregation),
fluid motion (orthokinetic aggregation) and differential settling
(Hartmann et al., 2014). Successful collisions result in the formation of
aggregates with increasing size and mass (Petosa et al., 2010; Wagner
et al., 2014). Derjagiun Landau Verwey Overbeek (DLVO) theory describes the interaction energies experienced by attraction forces when
spherical particles come into close contact to form aggregates (Salieri
et al., 2015; Praetorius et al., 2012). The point zero charge (PZC) represents the point of instability where a particle and solution are in a state
of equilibrium, repulsive forces are suppressed (Dastafkan et al., 2015;
Garner and Keller, 2014). The neutralisation of repulsive charges allows
particles to aggregate through dominant van der Waals interactions
(Dastafkan et al., 2015; Garner and Keller, 2014).
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
5
Table 1
Physical processes - dispersion, agglomeration, aggregation, heteroaggregation, adsorption and sedimentation studies for selected ENMs.
Particle
Average
particle size
(nm)
Dispersion
AgNP-Cit ~20
Agglomeration
AgNP-Cit ~25
AgNP-PVP
TiO2
AgNP-Cit ~61
AgNP-PVP ~42
TiO2
~317
AgNP-Cit ~60
Aggregation
TiO2
~21
TiO2
400
AgNP
40
AgNP-PVP ~100
Media
Main findings
Natural freshwater (PW), synthetic sea
aquatic (SW) & simulated estuarine
water
Citrate stabilized AgNPs nominally 20 nm in diameter partially agglomerated in the three Chinnapongse
PW samples tested, with varying fractions of AgNPs remaining in solution at equilibrium. et al. (2011)
AgNPs were least stable in SW, losing electrostatic colloidal stabilization rapidly due to
charge screening by greater NaCl concentrations and the presence of divalent cations.
Natural water
Agglomeration rate of Ag NP-Cit and TiO2 NP dependant on Ca2+ conc. AgNP-PVP
agglomeration in the presence of Mg2+ ions.
Topuz et al.
(2014)
Natural water
Ag NP PVP stable after 1 week (in all fractions). Ag Np-Cit and TiO2 positively correlated
with Ca2+ conc. Biopolymers provided stabilization in natural waters.
Topuz et al.
(2015)
Ultrapure water
Presence of alginate or humic acid differentially affects the kinetic of the agglomeration
process
António et al.
(2015)
River water
Aggregation of TiO2 is controlled by presence and concentration of DOM. Increasing
surface-adsorbed DOM leading to more compact structures.
Kaolin promotes NP aggregation via hetero-aggregation HA promotes aggregation of
TiO2 but decreases Ag aggregation due to the presence of a surface coating
Aggregation PVP-Ag NPs enhanced by addition of cysteine at high IS. Aggregation
inhibited by the presence of HA reducing deposition rate.
Chekli et al.
(2015)
Wang et al.
(2015b)
Yang et al.
(2014)
Millipore water
Fresh & saline waters
Reference
Hetero-aggregation
AgNP-PVP 90.5
SiO2 – Ag 124
CeO2
175
C60
Bare Ag
~60
Natural water
Hetero-aggregation rates ranged between 0.007 and 0.6 L mg−1 day−1. Highest observed Quik et al.
in seawater.
(2014)
Milli-Q
No significant differences in the stability of bare silver nanoparticle and clays at pH 7
when compared to the single particle systems of clay or silver at the same pH. High
electrolyte concentrations were needed to overcome the energy barrier to form
aggregates. Under neutral pH and moderate to elevated electrolyte concentrations
approaching the measured CCC values that binary systems of montmorillonite/illite and
silver nanoparticles can be treated as single component clay systems.
Adsorption
AgNP-Cit ~13
Distilled water
Higher ionic strength destabilized AgNPs while increased NOM concentration hindered
Bae et al.
soil adsorption. AgNPs in 50 mM NaNO3 continued to aggregate, however in the presence (2013)
of HA (50 mg/L), the size of AgNPs did not change and remained stable. The results
showed that both aggregation and soil adsorption increased with increasing ionic
strength. The presence of NOM had an opposite effect: both aggregation and soil
adsorption decreased in the presence of humic acid.
Natural water
The sedimentation rates ranged from 0.0048 m d−1 for PVP-Ag to 0.12 m d−1 for C60.
The apparent non-settling fractions (given as C15/Co ∗ 100%) after 15 d varied from
0.01% to 92% for the metal based ENMs. Sedimentation rates ranged from 0.0001 m d−1
for SiO2-Ag to 0.14 m d−1 for C60.
Ag and CeO2 NPs appear to bind less strongly to sediment constituents in comparison to
their ionic counterparts, rendering the nanoparticles more mobile.
Sedimentation
AgNP-PVP 90.5
SiO2 – Ag 124
CeO2
175
C60
AgNP
~10
CeO2
~6
Milli-Q water
Aggregation rates are influenced by the particle collision frequency, a high attachment efficiency to suspended solids is likely to result
in reduced transport potential and subsequently reduced bioavailability (ECHA, 2014). Particle surfaces emit an electrical double
layer (EDL) which maybe either negatively or positively charged repelling similar charged particles. This repulsion can be suppressed by
altering the ionic strength (IS) (Wagner et al., 2014; Petosa et al.,
2010). An increase in IS causes the EDL to become suppressed
resulting in a greater attractive surface potential. Increasing the electrolyte concentration influences aggregation kinetics, with the critical coagulation concentration (CCC) a measure of the minimum
electrolyte concentration required for suppression of the electrostatic energy barrier to allow aggregation to proceed (Dwivedi et al.,
2015). These closer interactions between particle surfaces allow
attracting bonds such as van der Waals interactions and magnetic
bonds to dominate (Wagner et al., 2014).
Liu et al.
(2015)
Quik et al.
(2014)
Van Koetsem
et al. (2015)
Non-DVLO processes (hydro-phobicity, steric repulsion and polymer bridging) also impact on the collision and attachment efficiency
preventing aggregation and subsequent deposition (Petosa et al.,
2010). Colloidal interactions of ENMs with natural colloids
(suspended solids (SS) and particulate matter (PM)) in the water
column could lead to larger aggregates of increased mass which
could migrate to the sediments in low velocity water (Huynh and
Chen, 2012). However, aggregates of larger fractal dimension or
chains will undergo limited settling due to their reduced relative
density (Chekli et al., 2015).
Chen and Zhang (2012) observed increased aggregation rates for
nano Ag PVP capped in CaCl2 and NaCl solution in the presence of
humic acid. The CCC was 40 times higher in NaCl solution compared
to CaCl2. Particle surface charges were eliminated under high electrolyte
concentrations allowing the particles to aggregate readily. They suggest
that nano Ag could aggregate more readily in seawater than lake water
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
6
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
Table 2
Chemical & biological processes – photochemical, biological transformation and dissolution studies of nAg.
Particle
Average
particle size
(nm)
Photochemical
Ag-PVP
~27
Media
Main findings
Reference
River, lake & spring
water
Yin et al.
(2015)
Ag-PVP
~25
Ultrapure water
Sunlight irradiation accelerated the photo-transformation of AgNPs in environmental waters. Larger
particles were more susceptible than smaller particles to photo-transformation. DOM inhibited both
aggregation and photo-transformation.
Photo oxidation resulted in the degradation of PVP coating enhancing Ag+ release.
Biological
n/a
n/a
AgCl
Reduction of silver Ag+ by Geobacter and Shewanella spp. To form nanosilver particles.
Distilled water &
ethanol solutions
PEO in the brush shells is bioavailable to bacteria through brush degradation.
Artificial natural
media
Relatively small fractions of Ag nanoparticles dissolved. The dissolved silver was stable from day 5 to
day 19.
Artificial seawater
& freshwater.
Freshwater and
seawater
Reconstituted
freshwater
The dissolution of silver was slow. Both Ag-cit and Ag-PVP dissolved progressively throughout a 30 day Thio et al.
period in natural seawater. Less than 1% of either coated Ag dissolved in freshwater after 30 days.
(2012)
Presence of Cl− enhances Ag NP dissolution.
PEO-brushed ~30
NPs
Dissolution studies
10–100
Ag-PVP
Ag-Cit
Ag-gelatin
Ag-chitosan
Ag-Cit
~40
Ag-PVP
Ag-Cit
20,50 & 80
Ag-PEG
6,9,13 & 70
AgNO3
n/a
Degradation of capping agents cause AgNPs to become unstable and release Ag+. Similar toxicity for 50
nm and 80 nm suspended in 0.003 medium and for 20 nm suspended in 0.015. Primary particle size of
50 and 80 nm in 0.015 medium played a major role in toxic responses to D. magna. AgNPs suspended in
0.015 were less stable than particles in 0.003.
Ultrapure deionized Extent of dissolution depends significantly on the particle size and pH of the aqueous system. A decrease
water
of pH from 7 to 3 increases the rate of dissolution.
Sterile distilled
Synthesis of AgNPs by plant systems. Ag+ ions reduced in plant extracts. Particles of nAg formed in the
water
size range of 2–5 nm
due to the high presence of divalent ions in seawater and the bridging
effects of HA.
Yu et al.
(2014)
Law et al.
(2008)
Kirschling
et al. (2011)
Odzak et al.
(2014)
Harmon et al.
(2014)
Peretyazhko
et al. (2014)
Jha et al.
(2009)
‘realistic’ predicted ENM concentrations (i.e. ppb) due to limitations of
measurement techniques and the fluctuations in natural colloids over
spatial and temporal conditions.
5.3. Hetero-aggregation
5.4. Adsorption
Hetero-aggregation is likely to be more relevant than aggregation
due to increased levels of naturally occurring geo and bio-colloids,
such as viruses, bacteria, proteins, DNA, spores, algae, protozoa and
other microorganisms, are expected to dominate (Wang et al., 2015a,
2015b; Keller and Auset, 2007).
Wang et al. (2015a, 2015b) studied the hetero aggregation efficiency
rates of engineered TiO2 and Ag NPs to kaolin clays under varying conditions of pH, IS and HA to assess the potential of kaolin clays as a coagulant for the removal of metallic ENMs. Kaolin clay was shown to
promote hetero-aggregation of TiO2 and Ag NPs to the kaolin clays by
destabilizing the NP surface charge. However, the presence of HA was
also shown to reduce hetero-aggregation, enhancing mobility and dispersal of the particles which would negatively impact sedimentation
rates.
Praetorius et al. (2012) modelled nano-TiO2 and suspended particulate matter (SPM) hetero-aggregation in the Rhine River. The model
highlighted the importance of the characterization of the SPM and the
affinity of the ENM and SPM to form hetero-aggregates. They found
that 0.36 kg day− 1 which represented 92% of the input flow of
0.39 kg day−1 of the first box of the model. This indicated that particles
are likely to remain in suspension and be transported in the water
column.
Quik et al. (2014) experimentally determined heteroaggregation
rate constants (K) and sedimentation rates for PVP and silicon dioxide
coated Ag, cerium dioxide (CeO2) and fullerene (C60) ENMs in natural
waters. A range of water samples were investigated representing
water bodies with varying salinity, acidity and OM content. They
noted that direct measurements of aggregation and sedimentation are
problematic or impossible to determine in actual natural systems at
Adsorption occurs when substances within the media attach to the
surface of AgNP or other metallic ENM making them either mobile or
stationary (Hartmann et al., 2014). Primarily, adsorption will be to natural organic matter (NOM) as this is the likely dominant substance present in the water column (Table 1). Components of NOM include HA,
fulvic acids and humin derived from the decomposition of plant and animal matter (Wang et al., 2015c). The adsorption of NOM could influence the behaviour of the particles as the surface properties will be
modified to some extent which may affect the attachment efficiency
(Schaumann et al., 2015). In the presence of NOM, aggregation can be
hindered causing particles to persist (Bae et al., 2013). Desorption
could also occur due to forces exerted on the particles i.e. shear forces
due to turbulence resulting in the removal of associated material,
changing the surface characteristics of the particle (Schaumann et al.,
2015).
NOM concentrations in natural waters can be as high as ~50 mg/L or
more, indicating a high potential for colloidal interactions (Liu et al.,
2013). NOM is negatively charged in natural water due to attached surface groups, although the complex structure of NOM results in varied
colloidal efficiencies associated with surface charge and density
(Wang et al., 2015a, 2015b).
5.5. Partitioning of nanomaterials
ENMs do not partition within environmental media in the same
manner as ‘traditional’ contaminants as they are not expected to reach
thermodynamic equilibrium: attachment to different surfaces will be a
function of time rather than concentration. Therefore they will not
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
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D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
7
form a true solution of low molecular-weight substances and will remain in a state of instability (Praetorius et al., 2012). Cornelis (2015)
also found that partitioning coefficients (Kd values) are not appropriate
fate descriptors for ENMs as an equilibrium assumption is not valid.
redox reactions is dependent on environmental chemistry and conditions (Nowack et al., 2012).
5.6. Sedimentation
Dissolution results in the continuous reduction of a particle surface
diameter with a final endpoint of particle extinction. Dissolution rates
are particle (size, structure and shape) and media dependent, influenced by the pH, solubility constant, chemical speciation and available
surface area, with dissolution leading to the complete or partial depletion of the solid phase (Baker et al., 2014; Peretyazhko et al., 2014;
Wagner et al., 2014). A dissolving particle surface whereby surface
ions are released into the media is a significant process for certain metallic NPs such as Ag and copper oxide (CuO), while for other particles
such as gold (Au), TiO2 and carbon based NPs this process is not applicable (OECD, 2014). Surfactant coatings, ligands or other surface molecules resulting from colloidal interactions with NOM and other
anthropogenic constituents will hinder efficiency of dissolution
(Wagner et al., 2014). Dissolution rates for metallic ENMs in natural waters are difficult to measure as most metals are not readily soluble in
water (Quik et al., 2011). Dissolution rate can be estimated for a substance using the Noyes-Whitney equation (Education, 2016) (Eq. (4)).
Sedimentation is a key mechanism for the removal of ENMs. Increasing particle size and density caused by colloidal interactions resulting in
the formation of aggregates which are subjected to settlement due to
gravity (Hartmann et al., 2014). Gravitational force can be represented
through Stoke's law, which calculates the settling velocity of particles
in a solution based on their radius and density in comparison to the
fluid viscosity and density (Hartmann et al., 2014). Effectively, the larger
the particle the more rapidly it will settle in comparison to smaller more
dispersed particles. The sediments are likely to contain the highest concentrations of ENMs and so are a crucial compartment in environmental
fate modelling (Koelmans et al., 2015). Dale et al. (2015a, 2015b) observed slow rates of penetration to deeper sediments resulting in high
concentrations of Ag accumulating in surface sediments. Accumulations
in surface sediments may present a pathway for uptake by benthic organisms (Boxall et al., 2007; Klaine et al., 2008).
An estimation of sedimentation rates were derived by Quik et al.
(2014) using Eq. (3).
−ðVs
þkdis Þt
h
C t ¼ ðC 0 −C ns Þe
þ C ns
ð3Þ
where: Cns is non-settling concentration [Cns [g/L], vs is sedimentation
rate [m d−1], h is sedimentation length [m], kdis is dissolution rate and
t is time. A sedimentation rate for PVP coated AgNPs was estimated to
be 9.98 E−3 (m d−1) in river water. A study on the sedimentation potential of Ag and CeO2 found that the particles had weak binding potential to sediment constituents and could be potentially more mobile (Van
Koetsem et al., 2015).
6. Chemical and biological processes likely to influence fate and behaviour of metallic ENMs
Chemical characteristics of a water body will influence both chemical and biological processes which could have a significant influence
on the fate and behaviour of ENMs in aquatic media. These processes
are discussed in the next section with summary studies in Table 2.
6.1. Photochemical
Ultra violet (UV) solar radiation can induce changes to particle surface properties through the absorption of photons or the decomposition
of surfactant coatings (Hartmann et al., 2014). Silver NPs are known to
form reactive oxygen species (ROS) and induce oxidative stress in
cells (Schaumann et al., 2015), while TiO2 nanoparticle exposure to
UV light also resulted in the formation of ROS (Feckler et al., 2015;
Dalai et al., 2013). Sunlight irradiation of PVP coated AgNPs was
shown to impact greater on larger particles than smaller particles making them more susceptible to photo transformation (Yin et al., 2015).
Photo oxidation of PVP coated AgNPs has been observed during light irradiation experiments resulting in the release of Ag+ ions, indicating
the potential for increased toxicity of PVP stabilized AgNPs in waterways (Yu et al., 2014).
6.2. Redox reactions
Silver NPs can be destabilized under relevant environmental conditions resulting in the release of Ag+ (Reidy et al., 2013). Released ions
are free to reduce other chemical species (gain of electron). Exchanges
of ions during redox reactions are primarily associated with increased
toxicity of metallic ENMs (Hartmann et al., 2014). The efficiency of
6.3. Dissolution
dM
D
¼ A ðCs −Cb Þ
dT
d
ð4Þ
where: dM = solute dissolution rate (kg/s−1), m = mass of dissolved
material (kg), dT = time, A = surface area of the solute particle (m2),
D = diffusion coefficient (m/s−1), d = thickness of the concentration
gradient (m), Cs = particle surface concentration (kg/L), Cb = concentration in the bulk solvent/solution (kg/L).
Liu and Hurt (2010) developed and proposed an empirical kinetic
law for the dissolution of ion-free, Cit stabilized nAg under environmentally relevant conditions. They suggest that Cit stabilized nAg will not
persist as a particle but will be converted to ionic form through dissolution processes in the presence of dissolved oxygen. The rate of dissolution in this study was slow with a range of 6–125 days to completion
indicating that the particles in natural environments may persist and
be subjected to other interactions within the water column. Odzak
et al. (2014) indicated in their study that initial dissolution was observed but NPs became stable from the 5th day to the end of their
study at 19 days also indicating slow rates of dissolution. Thio et al.
(2012) found similar findings with Ag Cit coated and PVP coated
AgNPs in freshwater with ~1% dissolved after 30 days, while progressive
dissolution was observed over the 30 days for particles in natural sea
water. Baalousha et al. (2016) showed the importance of NP concentration for the fate and effects of ENMs during dissolution and aggregation.
The study suggests that more environmentally relevant nAg concentrations (i.e. low ppb) result in higher rates of dissolution and deceased aggregate size as compared to ‘traditional’ laboratory concentrations (i.e.
low ppm) suggesting a higher mobility potential for nAg than previously assumed.
6.4. Precipitation and speciation
Precipitation is a reversal of the dissolution/solubility process
resulting in the formation of dissolved metal ions agglomerating to reform a solid phase which may then settle out of solution (Hartmann
et al., 2014; Wagner et al., 2014). Precipitation processes are unlikely
to occur in natural systems due the presence of high levels of bio and
geo colloids which will drive the formation of heteroaggregates
(Dwivedi et al., 2015).
Speciation is the transformation/modification of particle surfaces
through interactions with ions and molecules, resulting in particle
functionalization and toxicity characteristics potentially different to its
original form (Barton et al., 2015a, 2015b). Metallic ENMs with the
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
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D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
same core chemical make-up but differing speciation may undergo
different environmental fate and behaviour processes, with surface
characteristics limiting attachment efficiency to various surfaces and
bioavailability to organisms.
6.5. Bio-transformations
Bio-transformation can result in the biodegradation or biomodification of an ENM. Although it has been hypothesized that these
processes could occur there is scarce data on the implication of surface
coatings and the effect they may have on these processes (Hartmann
et al., 2014). Aided by microorganisms, biodegradation resulting in the
disintegration of a material is mainly relevant for carbon based materials, while metallic ENMs are inorganic and therefore not susceptible
to the same process (ECHA, 2014; Hartmann et al., 2014). However, filter feeding organisms may actively transform metallic ENM surfaces by
digesting surface coatings exposing pristine particles to environmental
processes (Quik et al., 2011). Microbial organisms have been known
to reduce Ag+ and form AgNPs potentially reducing the toxic effects of
Ag ions (Law et al., 2008). Furthermore, bio-modification could occur
upon ingestion by an organism which could alter the original surface
properties and physical characteristics of the particle (Hartmann et al.,
2014). Attachment of extra-cellular material released from an organism
could potentially bind to the surface of particles resulting in modifications to surface characteristics and properties (Hartmann et al., 2014).
Evidence from existing literature, as referred above, indicates the
role and significance each natural attenuation process takes in potentially influencing metallic and non – metallic ENM fate and behaviour
in the natural aquatic environment (Tables 1–3). It should be noted
that, these processes do not happen in isolation or in structured sequence as the natural environment is dynamic and complex which
adds greater complexity when trying to assess risk factors. From Table
3, evidence suggests that hetero-aggregation, dissolution and sedimentation are likely to dominate potential removal of AgNPs and other metallic ENMs in the natural aquatic environment. Hence, any proposed
risk framework needs to ensure adequate capture of these processes
in the natural environment.
7. Environmental concerns of ENMs
Unnatural forms of materials entering in to the environment may
present toxicity concerns as it is unclear how organisms will interact
and can accommodate these materials. Extensive toxicity studies have
been conducted to determine the potential risks posed to the environment from ENMs (Buzea et al., 2007; Griffitt et al., 2008; Croteau et al.,
2014; Gondikas and Morris, 2015). While biota have adaptability
mechanisms developed through evolutionary traits to survive in the
presence of nano-scaled particles formed naturally. It remains unknown
if these adaptions can also deal with ENMs considering their engineered
properties. In relation to AgNPs, the release of surface Ag+ during dissolution is a concern for aquatic organisms as Ag+ have been associated
with cellular interference and enhanced by reactive oxygen species
(ROS) (Navarro et al., 2015). Exposure assessments to varying concentrations of ENMs have been conducted on a diverse range of test organisms for determination of lethal dose concentrations (LD50) and
effective concentrations (EC50) (Blaser et al., 2008). While these studies
provide best estimates for toxicity levels of concern they are performed
with unmodified particles at concentrations unlikely to be present in
the natural environment. Klaine et al. (2008) and Garner and Keller
(2014) reviewed fate and toxicity studies of ENMs while McGillicuddy
et al. (2017) presents a current review of the sources, detection and toxicity effects of AgNPs in the aquatic environment.
8. Nanometrology
Measuring of particles on the nanoscale is relatively complicated in
natural environment settings. Simply detecting, quantifying and characterizing metallic ENMs in complex media presents numerous challenges
for assessors in assessing fate, mobility and toxicity in aquatic environments. Natural waters and the omni presence of multiple naturally occurring colloidal materials, could potentially interact with ENMs and
alter their physical, chemical or biological state. There are many tools
and methods to analyse primary ENMs in relatively simple media,
their application to more complex media is still underdeveloped
(Montano et al., 2014). Current tools for detection and characterization
of ENMs are not adaptable for in-situ monitoring. Another obstacle to
monitoring in-situ is that prior to analysis, samples must be prepared
which could potentially result in contamination of samples. In-situ sampling is also challenging and problematic as perturbation during sampling can result in disturbances to the sample. These sampling
challenges can affect the true results of analysis and potentially result
in under or over estimation of NP concentrations. Presented in Table 4
are a variety of analytical techniques that are used in nanometrology
to characterize and detect nanoscaled materials. No single analytical
method exists that can adequately detect, quantify or characterize
ENMs, although a combination of analytical methods may provide
nanometrology data of ENMs in the natural environment.
9. Modelling nanomaterial environmental fate in aquatic systems
Nanomaterials are traditionally assumed, perhaps erroneously, to
undergo similar fate and transport processes to existing, traditional
Table 3
Summary of environmental processes and their likely impact on the fate and behaviour of Ag in aquatic systems.
Reference
Environmental process
Mechanism
Effect
Mobility Bio-availability Significance in natural
aquatic
environment
Dispersion
Agglomeration
Homo-aggregation
Brownian motion
Particle collision potential
Particle collision
↑
↓
↓
↑
↑
↑
⁎⁎
⁎
⁎
Römer et al. (2011)
António et al. (2015)
Yang et al. (2014)
Hetero-aggregation
Particle collision
Dilution
Cluster formation
Irreversible particle
formation
Interaction with background
colloids
Surface Ag+ release
↓
↓
⁎⁎⁎
Quik et al. (2011)
⁎⁎⁎
Dobias and Bernier-Latmani
(2013)
Hartmann et al. (2014)
Hartmann et al. (2014)
Dissolution
Solid nAg dissolves
Sedimentation
Sorption
(adsorption/absorption)
Gravitational settling
Binding with NOM or solid
surfaces
Removal from water column
Surface transformation
+
(Ag ) ↑
(nAg) ↓
↓
↓
+
(Ag ) ↑
(nAg) ↓
↓
↑
⁎⁎⁎
⁎⁎
↑ increase mobility or bio-availability.
↓ decrease mobility or bio-availability.
⁎⁎⁎ Highly significant.
⁎⁎ Moderately significant.
⁎ Mildly significant.
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
9
Table 4
Analytical methods for nanomaterial identification, quantification and detection.
Analytical method
Acronym
Transmission electron microscope/energy dispersive X-ray
Scanning electron microscope/energy dispersive X-ray
Nuclear magnetic resonance
Single particle-inductive coupled plasma-mass
spectrometry
Inductive coupled plasma-mass spectrometry
Inductive coupled plasma-spectrometry
Ultraviolet–vis spectroscopy
Dynamic light scattering
Flow-flow field fractionation
Sedimentation field flow fractionation
Nanoparticle tracking analysis
Zeta potential
Hydrodynamic chromotography
Size exclusion chromotography
Cross-flow ultra-filtration
Centrifugation
TEM/EDX *
SEM/EDX
NMR
SP-ICP-MS *
X-ray absorption spectroscopy
ICP-MS
ICP-OES
UV–vis
DLS
FL-FFF
Sed-FFF
NTA
Particle size (1–100
nm
range)
Particle Surface
Shape Elemental
Reference
conc.
charge/groups
composition
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
*
HDC
SEC
CFUF
*
*
*
*
*
*
*
*
XAS
*
Brunetti et al. (2015)
Schaumann et al. (2015)
Montano et al. (2014)
Baalousha et al. (2016)
Mitrano et al. (2014)
Quik et al. (2011)
Schaumann et al. (2015)
Weinberg et al. (2011)
António et al. (2015)
Montano et al. (2014)
Quik et al. (2015)
Montano et al. (2014)
Schaumann et al. (2015)
Montano et al. (2014)
Montano et al. (2014)
Van Koetsem et al.
(2015)
Brunetti et al. (2015)
* Indicates detection.
contaminants and therefore existing fate and transport models may be
adapted to assess these materials. There is considerable uncertainty
and variability associated with the fate and transport of potentially reactive materials such as nAg as the ENM core and surface coatings can be
subjected to chemical or biological degradation and transformation.
Fate and transport models for NMs typically assume steady state environments which can be an impediment to environmental realism
when modelling and predicting results of environmental processes
(Quik et al., 2014). Various modelling approaches are detailed in
Table 5, including software tools and bespoke modelling approaches
(i.e. no formal name, tailored by authors for a specific purpose).
Material flow analysis (MFA) has been used to model NMs by a number of researchers (Nowack and Mueller, 2008; Walser and Gottschalk,
2014). While MFA can provide insight into the possible life cycle of a
material, these models have no spatial detail, instead averaging mass
over an entire environmental compartment (Dale et al., 2015a,
2015b). Life cycle assessment by MFA modelling of NMs identified key
areas of study for estimation of nanoenabled products being released
to the environment (WWTPs, bio-solids, landfills, and incinerators).
These starting points provided estimates of PEC in the absence of NM
production volumes and release data. However, MFAs lack of spatial resolution and renders MFA inappropriate as a tool for PEC's and therefore
providing little practical relevance for regulatory purposes (Dale et al.,
2015a, 2015b).
Dale et al. (2015a, 2015b) recently explored the current state of the
science in nanomaterial fate and behaviour modelling in aquatic environments. They discussed the strengths and weaknesses of current
fate modelling approaches while identifying crucial gaps in modelling
such as the lack of environmental realism and the role of surface coatings and NOM. Dale et al. (2015a, 2015b) observed high ENM spatiotemporal variability in natural river systems, indicating that current
steady state models may not be representative for assessing exposure
in watersheds and thus may misrepresent risk. This highlights spatiotemporal variability as a critical factor in future risk assessment models.
Therezien et al. (2014) modelled homo/hetero-aggregation of NPs and
larger background particles over a number of size classes and
Table 5
Summarizing modelling approaches for ENMs in the environment.
Model
Method/processes
Material flow
analysis
(MFA)
Model quantities of ENMs released to the environment using a life-cycle
perspective for nano-Ag, nano-TiO2 and Carbon nanotubes (CNT). The risk
assessment covered water, air and soil compartments in Switzerland. Calculations of PEC were derived based on substance flow analysis.
NANO DUFLOW Spatially explicit modelling of homo and hetero-aggregation, dissolution,
degradation, Sedimentation and re-suspension in a natural river system.
The model outputs analysed included concentrations of free, homo and
hetero-aggregated ENPs in the water column and sediments over a distance
of 40 km from source. Monte Carlo probabilistic modelling was employed to
account for uncertainty.
Bespoke
Predicted spatial concentration profiles of nano TiO2, where TiO2 NPs were
grouped together into clusters. Similarity between cluster groups were
analysed and grouped and using cluster analysis these groups were then
used as representative groups for spatial concentration profiles.
Simplebox4nano Aggregation, attachment and dissolution. Mass balance equations linking all
(SB4N)
concentration and processes using first order rate constants.
Bespoke
Bespoke
Estimation of TiO2, Ag and CeO2 ENMs released from nano-enabled products
through WWTPs to Irish surface waters. Predicting releases through usage,
disposal and treatment processes.
Modelling flows and concentrations of nine ENMs. Probabilistic flow
modelling based on probabilistic production and use estimates
incorporating Monte Carlo simulations for probability distributions.
Main findings
Reference
Expected concentrations of CNT and nano-Ag probably pose
little risk based on available data. Nano-TiO2 may pose a risk to
aquatic life.
Nowack and
Mueller
(2008)
Sedimentation ‘hot spots’ and ENP speciation can be predicted Quik et al.
as a function of place and time. Outputs for Ag and CeO2 were
(2015)
similar indicating hetero-aggregation drives the fate of ENMs in
surface waters.
Hetero-aggregation influences fate of ENMs. Future modelling
should be directed to water characteristics in regions near
expected ENM emission sources as fate processes maybe readily
predicted in these locations.
SB4N predicted environmental concentrations maybe useful as
background concentrations in environmental risk assessment.
Semi-quantitative methodology for environmental exposure
assessment for surface waters.
Highest concentrations were predicted for carbon black and
photo stable TiO2 because of their production volumes. Ag is
expected to be in low concentrations in environmental media.
Sani-Kast
et al.
(2015)
Meesters
et al.
(2014)
O'Brien and
Cummins
(2010)
Gottschalk
et al.
(2015)
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
10
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
concentrations. The conclusion of their study indicated that background
particle interaction with NPs was size dependent, with larger background particles colliding less frequently with NPs. This size dependent
factor could result in NPs persisting in water.
Meesters et al. (2014) used a SimpleBox4Nano (SB4N) model to predict the transport and concentrations of ENMs in different environmental compartments and the transformational processes that influence
ENM fate and behaviour. First order kinetic rates for aggregation and attachment processes were used to track colloidal behaviour of NMs, with
hetero-aggregation based on collision and attachment efficiencies. The
SB4N model output suggests that released ENMs can be assumed to be
freely dispersed, hetero-aggregated or increase in mass over time.
10. Human health concerns
Cell line and biological organism studies have been conducted with
ENMs to categorize the risk they may pose to human health. One recent
comprehensive review covered nanoparticle induced mitochondrial
toxicity (Maurer and Meyer, 2016), while others have investigated molecular toxicity mechanisms (McShan et al., 2014) and DNA damage,
toxicity and functional impairment in human cell lines (Hackenberg
et al., 2011).
While current predictions for environmental ENM exposure levels
are unknown to impact on public health, in the absence of adequate
risk assessment models it is difficult to gauge the exact effect level without a significant degree of uncertainty, especially as the sector develops
commercially (Koelmans et al., 2015). Few actual measurements of
ENM environmental concentrations exist, modelling estimates in natural waters are in the range of nanogram/L (ng/L)–milligram/L (mg/L)
(Nowack and Mueller, 2008; Liu et al., 2014). Concentrations of nAg
have been estimated to range between 0.002 and 0.18 ng/L in
European surface waters (Dumont et al., 2015), while US waters were
estimated to have more elevated levels ranging between 40 and
320 ng/L (Blaser et al., 2008). A stand alone study by (Li et al., 2016) estimated AgNP concentrations to range between 2.0 and 8.6 ng/L when
measured at an outflow discharge pipe from a wastewater treatment
plant (WWTP) along the Isar river in Germany. Measurements taken
1.5 km from the discharge point indicated significant reduction in concentrations along the river with concentrations ranging from (0.0–
2.3 ng/L) suggesting that particles can be readily removed from the
water column or that they are so diluted that they become undetectable.
Nanomaterials such as AgNPs are also used in domestic filters and
water treatment processes for the removal of contaminants and bacteria
through the use of nano-enabled membranes (NEMs) potentially increasing the risk of inadvertent exposure to NMs (Matthews, 2015).
Nano-enabled membranes act as reactive sites for incoming contaminants and bacteria allowing for efficient inactivation, degradation or removal from the feed water. Assuming AgNPs are not fully removed from
treated water there is a likelihood for ingestion through consumption.
Ingestion of nAg and subsequent internal distribution throughout the
body could be dependent on the properties of the particles consumed,
with likely accumulation endpoints in the vital body organs of the
spleen, liver and kidneys. Studies on the distribution of nAg in mammals
have mainly focus on observed changes in blood chemistry, with
ingested coated and uncoated AgNPs absorbed through gastrointestinal
tract of rodents and into the blood vessels resulting in transport to the
vital organs (liver, kidneys and spleen) (Kovvuru et al., 2015).
Chronic exposure to Ag could result in the medical condition known
as Argyria, which is known to be caused by over exposure to elemental
Ag. Chronic exposure has resulted in blue-grey discolorations of skin
pigment and eyes in some patient cases (Bachler et al., 2013). While
there is no known effective medical cure or therapy for Argyria, its
long term health implications are still unknown except for the physiological impact of skin discoloration for the patient concerned (Bachler
et al., 2013).
An estimation of ENM exposure was modelled by O'Brien and
Cummins (2010) for population exposure through drinking waters for
a selection of metallic based ENMs (Eq. (5)).
Eannual exposure X¼W daily 365Aaccum Source
ð5Þ
where E is the annual exposure (ng yr−1) resulting from the consumption of water based on the value of X which is the specific ENM, Wdaily is
the daily water consumption (litres per person per day), A is the accumulated fraction of ingested dose and Source is the treated drinking
water likely to contain residual fractions of NM (ng) as estimated by
the author for each of the water schemes operating in Ireland. While
this study focused on drinking water consumption in Ireland, global
water consumption patterns vary widely, as do methods of water sourcing depending on the geographical location. Currently there are no standardized limits affiliated to nAg but the World Health Organization
(WHO) estimated for elemental Ag concentrations in US drinking waters to be in the range of 0–5 μg/L, and with water treated with Ag to
be N50 μg/L (WHO, 1996). In comparison, drinking water in Canada
was found to have concentrations in the range of 1–5 ng/L. The long
term health effects of Ag exposure are unclear but an acute lethal dose
of 10 g has been estimated (WHO, 1996). Nano Ag has been observed
to have an effect on cellular function with a sub-lethal exposure of 1.8
μg/L AgNPs resulting in endocrine disruption in test organism Xenopus
laevis tadpoles (Carew et al., 2015).
11. Quantification and risk assessment modelling
Existing risk assessment techniques have been assessed and
adapted, and new methods developed to quantify the likely exposure,
hazard and subsequent risk. Furthermore, when assessing exposure to
the environment it is also important to consider the presence of existing
chemicals from previous releases which can contribute to background
concentrations (Commission, E, 2003). There have been 3 key prerequisites proposed by the European Commission (EC) for the characterization and measurement of NMs for hazard and risk characterization
(ECHA, 2014):
1. An enforceable definition for NMs
2. Agreed data set of physic-chemical properties (e.g. size, surface area,
etc.)
3. Standardized methods for the quantification of these parameters.
The variable and complex nature of aquatic systems and the relatively small size and distribution of NPs present difficulties for their identification, quantification and distinction from natural colloids (Coll et al.,
2015). The dynamic and complex nature of a natural system creates uncertainties in understanding NP behaviour during their life cycle (ECHA,
2014). Environmental fate assessments can also be unreliable due to a
lack of harmonized testing proceedings (ECHA, 2014). Koelmans et al.
(2015) suggested that for environmental risk assessment a predictive
risk assessment (PRA) approach using read-across methods for known
fate and effects data on chemical should be used to determine PEC.
Production and potentially released concentrations have been estimated by Piccinno et al. (2012) for 10 different ENMs using European
and worldwide production surveys, focusing on the ENM distribution
to different product categories creating baseline data for life cycle analysis of nano enabled products. The results of the survey indicated that
TiO2 was the most abundantly produced ENM worldwide with up to
10,000 tons produced worldwide while Ag, quantum dots (QDs) and
fullerenes production in Europe was b10 tons each per annum.
Gottschalk et al. (2015) examined material flow data for nine selected
ENMs of concern to the Danish environment, with carbon black and
TiO2 predicted to result in the highest concentrations in natural aquatic
systems. Quik et al. (2015) presented a fate model (Nano DUFLOW)
using five size classes of ENMs and five classes of natural colloids to
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
model specific fate processes in a spatially explicit hydrological model
as a function of both place and time. The spatial profiles calculated for
cerium dioxide (CeO2) and Ag particles were very similar indicating
that similarities can be draw between the two materials when assessing
environmental risk. Concentrations of 5.1 ng/L Ag and 5.5 ng/L CeO2
where observed downstream of the test site. O'Brien and Cummins
(2010) provided estimates for environmental concentrations of TiO2,
Ag and CeO2 ENMs released from nano-enabled products through
WWTPs to Irish surface waters. While the study does not present a
quantitative assessment it presents a methodology from which environmental exposure risk maybe assessed.
12. Policy developments and legislation
Chemical substances entering European markets are evaluated by
the European Chemicals Agency (ECHA) through the Registration, Evaluation, Authorization and Restriction of Chemicals (REACH) tool. The
legislation provides for the safe manufacturing and importation of
chemicals for use within the European Union (Hansen et al., 2011;
Clément et al., 2013). The basis for the assessment of the ecological
and human hazard posed by a certain material is based on toxicity
tests performed on bulk (macro) materials with, as of yet, no direct provision for their nano-form.
The current REACH legislation does not specifically refer to NMs, although they are considered to be incorporated under their substance
definition under REACH Article 1(3) (ECHA, 2014). However, it may
be argued that under the regulations nano-scale materials are still
classed as bulk materials regardless of their dimensional characteristics
(Clément et al., 2013). The regulatory commissions second review on
NMs states that all substances either completely or partly comprising
nanomaterial form are equally covered by REACH legislation (ECHA,
2014). Therefore, close collaboration is needed between all stakeholders
and international organisations such as the Organisation for Economic
Cooperation and Development (OECD) in facilitation of knowledge
transfer in order to meet challenges of regulatory conformance (ECHA,
2014).
13. Knowledge and data gaps
Establishing and collating information systems relating to increasing
production, use and disposal of products and processes that utilize
ENMs is required to effectively quantify volumes for monitoring purposes (Hartmann et al., 2014). Such data is invaluable for predicting
baseline data on environmental concentrations for quantitative risk
assessment.
The biological, physical or chemical transformation of ENM cores
and surface coatings upon interaction with naturally occurring geo
and bio-macromolecules must be both qualitatively and quantitatively
determined (Hartmann et al., 2014). With that in mind significant challenges remain in differentiating natural from engineered nanoparticles
(ENPs) in natural media and ions from very small NPs (Schaumann
et al., 2015). Applicability of REACH testing strategies and standardized
testing procedures for hazard and toxicity assessment may therefore require the development of nano-specific testing for ENMs (ECHA, 2014).
Uncertainty around the transformational processes of ENPs in natural
waters could lead to the possibility of both over and underestimating
the toxic effects of ENMs on resident organisms (ECHA, 2014). This uncertainty highlights the need for behavioural studies on ENMs in natural
aquatic systems in order to insure proper regulatory controls are implemented (Park et al., 2014).
Current methods for estimating exposure and identifying hazard are
primarily based on the fate and effects on pristine NPs, however these
exposure estimates and toxic effects do not necessarily reflect the transformational processes that released manufactured ENMs undergo
through natural attenuation and colloidal ageing (Schaumann et al.,
2015; ECHA, 2014). While sediments are a likely sink for ENMs in the
11
aquatic environment the long term low dose effect to benthic organisms
and the potential chronic exposure and bioaccumulation need to be addressed (Schaumann et al., 2015; ECHA, 2014).
Studies on the “ageing” of ENMs in the environment and the continuous evolving modifications from interactions within complex media
are currently lacking (Reidy et al., 2013). Subsequently, knowledge
gaps remain for the understanding of the long term consequences for
biota and the environment as a whole.
Further development of analytical methods that don't compromise
the integrity of samples should be prioritized in order to validate results
of environmental samples. Current methods utilizing X-rays, electron
beams and so forth require prior sample preparation and could potentially impact on the integrity of the samples and result in the unwanted
introduction of artifacts. Hyper spectral imaging which has been used
extensively in geology and mineral exploration has been identified as
a promising method that is a relatively non-disruptive method for characterization of NPs in complex matrices when coupled with enhance
Dark field microscopy (Badireddy et al., 2012).
14. Conclusion
Considerable uncertainty still remains as to the fate and behaviour of
nAg and other metallic ENMs entering into aquatic environments
resulting in significant variability in exposure estimates. The identified
process of dissolution discussed in this review is controlled by a number
of variable factors. Therefore, rates of dissolution for soluble AgNPs and
other soluble metallic ENMs in complex media will be dependent on reliable measurements under realistic and relevant environmental conditions. The expected increase and utilisation of AgNPs and other metallic
ENMs will likely release either directly or indirectly these materials to
the various environmental compartments. Pathways to aquatic environment will accumulate these materials from other compartments
resulting in aquatic compartments identified as an environmental sink
for these materials. Monitoring of aquatic health of organisms
inhabiting the Limnetic and Benthic zones should indicate if accumulation of AgNPs and other metallic ENMs are having toxic effects.
Aquatic environments by their very nature are highly dynamic with
multiple processes taking place simultaneously. Degradation of natural
and anthropogenic materials present in the water system suggests
that it is almost certain volumes of ENMs entering aquatic compartments are in concentrations several orders of magnitude lower ubiquitous natural background colloids. Therefore, at current and predicted
ENM release rates, ENM and natural colloid association is likely resulting
in removal from the water column. Although, the unceasing volumes
entering the environment may unbalance this understand in the future.
The removal efficiency of AgNPs and other metallic ENMs is dependent
on the chemical conditions of the receiving environment, the types of
organic materials present and the physiochemical properties of the
ENM surface. Natural organic matter adsorption to particle surfaces
can inhibit the process of dissolution and subsequent ion release thereby limiting the toxic potential. Gravitational settling causes particles of
sufficient mass to migrate to the sediment compartment and being sequestered within the sediments. This should somewhat limit availability to aquatic organisms present in the water column. However,
accumulation within the sediment compartment may result in bioavailability to benthic organisms and a potential pathway to higher trophic levels within the food chain.
The physical properties of metallic ENMs with stabilized surface
coatings (PVP, Cit or PEG) may inhibit environmental transformation
potential. Particle fate will be driven by the potential for surface transformations, particle instability, UV destabilization, mobility and dissolution potential. While risk assessment modellers have long called for
studies on the transformation of ENMs under environmentally relevant
conditions, data has not been forthcoming, meaning models must currently be populated with highly uncertain or bridging data.
Please cite this article as: Shevlin, D., et al., Silver engineered nanoparticles in freshwater systems – Likely fate and behaviour through natural
attenuation processes, Sci Total Environ (2017), https://doi.org/10.1016/j.scitotenv.2017.10.123
12
D. Shevlin et al. / Science of the Total Environment xxx (2017) xxx–xxx
Human exposure to AgNPs and other metallic ENMs in aquatic systems would be expected to occur through the consumption of tainted
aquatic organisms or drinking water sourced from surface waters. During water treatment processes the characteristic size of NPs means that
they may not be adequately removed and therefore may persist in post
treated drinking water. Presence and persistence of ENMs in post treated drinking water raises health concerns as their small size means that
they have the ability to penetrate cellular membranes. AgNPs and Ag +
have been associated with interference with cellular functions or signaling which could potentially induce toxic effects. Therefore, there is a reliance on operational efficiency of drinking water treatment plants
(DWTP) to reduce/remove ENMs of concern to a tolerable level. This
an area that must be adequately assessed as there are currently no defined limits for ENMs in post drinking water standards, particularly as
it could result in a public health concern.
Greater co-operation is required between stakeholders (ENM manufacturers, industrial consumers and international regulators) in order to
quantify the risks of ENMs to the environment and the human health.
Until the hazards posed to environmental and human health by ENMs
is proven to be minimal/negligible, or their environmental fate and behaviour, is fully understood, metallic ENMs should be considered a potential risk to environment and human health and therefore handled
and regulated as such.
It should be noted that the nanotechnology sector is still in its relative infancy with much more research yet to be conducted. Considering
aquatic environments are invariably complex, considerable knowledge
deficits remain on fate and behaviour processes likely to impact on
ENMs. Therefore, further research and development of analytical
methods for the detection, characterization and risk of ENMs should
bridge the knowledge gaps in the future aiding risk assessment and
regulators.
Acknowledgements
The authors would like to acknowledge financial support from the
Irish Environmental Protection Agency for this work under the EPA Research Program 2014–2020 (DeTER Project No: 2014-HW-MS-1).
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