07/04/2018

Industry Scientific input to AMEC Report on Primary

Microplastics

Executive summary

An appropriate definition of what constitutes intentionally added Microplastics (MPs) as well as a

science-driven and evidence-based approach are essential elements to build an effective EU

Strategy for plastics, and define the related policy actions.

The definition should cover materials that contribute to aquatic litter and the potential associated

risks, which relate mainly to physical effects in aquatic life forms based on todays’ scientific

knowledge.

The aim of this paper is to summarise the current state of science on the environmental fate and

effects of intentionally added potential micorplastics from cosmetic and personal care products

(CPCPs) and household care products (HCP) and to demonstrate these should not be classified as

PBT substances.

Substances of very high concern (SVHC) are materials that may pose serious impacts on human

health and/or in the environment. For environmental concerns, such substances are persistent,

bioaccumulative and toxic (i.e., PBT). While the vast majority of microplastics (i.e., excluding

biodegradable plastics) have environmentally persistent properties, there is sufficient information

to cast considerable doubt regarding their potential to bioaccumulate or resulting in relevant

ecotoxicities.

Cosmetics Europe and A.I.S.E. recognize the global concern regarding increasing pollution from

plastics in the oceans. The contribution of primary MPs potentially deriving from CPCPs and HCP is

very low (e.g. estimated contribution in the North Sea marine environment of 0.1% – 1.5% for

PCPC, while for HCP it is estimated at 0.02% of the worldwide primary microplastics estimated to

be released in the environment per year by Eunomia). Microplastics used in rinse-off products may

be released in the environment down-the-drain into the aquatic and marine environment. In

North America, Europe and Australia releases are generally subject to waste water treatment via

wastewater treatment plants (WWTPs), in which particles are removed via a physico-/chemical

process with a high degree of efficiency.

Environmental exposure of MPs was estimated by AMEC using the EUSES model in a very

conservative manner. However, EUSES was not developed for modelling particles and simulation

of MPs by a high logKow lead the estimate to be biased toward settled sludge and sediment. There

are no linear relationships of logKow with particle behaviour. Hence, it is highly recommended that

exposure modelling take into account the state-of-the-science modelling tools.

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Regarding effects, the primary role of MPs in adversely affecting aquatic biota is via their non-

nutritive contributions to biologically useful energy and physical effects such as an inflammatory

response. However, nearly all published studies report adverse effects observed at much greater

concentrations than those observed in the environment, and are, therefore unrealistic, especially

when assessing physical effects in comparison to naturally occurring particulate matter. The

consequence is that observed physical effects are environmentally irrelevant. No adverse effects

have been reported neither for soil-dwelling organisms nor for sediment-dwelling organisms at

concentrations found in the environment, although we do acknowledge that published data on

effects in soil and sediments is less abundant than for aquatic organisms. Further research to

better determine sediment and soil exposure levels and potential effects on such organisms is

recommended to close this knowledge gap.

Questions have been raised in the literature regarding the role of MPs in serving as vectors of

exacerbating the uptake of POPs (persistent organic pollutants) by aquatic organisms. State-of-

the-art literature that have investigated the relationships of lab-, field-measured and modelled

concentrations in media and biota have clearly indicated that it is highly doubtful that MPs can

serve as vectors. Hence, they cannot exacerbate the bioconcentration and bioaccumulation of

chemicals known to be either B or vB.

The AMEC report estimates the potential environmental concentrations of additives from

microplastics. Cosmetics Europe and A.I.S.E. do not challenge the fact that additive leaching can

occur from Microplastics but, considering the amount of plastic littering as a whole, the

dimensions of additive leaching from primary MP appear disproportionate in the AMEC

considerations.

Finally, when establishing whether potential microplastic ingredients pose an unacceptable

environmental risk it is essential to compare the concentration of microplastic in the environment

with the concentrations that cause an observed adverse effect. It is acknowledged that

microplastics are found in the environment, and could satisfy persistence criteria. It is however

clear that there is sufficient data regarding their lack of bioaccumulation, insufficiency to serve as

vectors for PBTs/POPs casts great doubt the applicability of considering microplastics as PBTs.

There is sufficient information to negate any B and T criteria. Consequently, microplastics cannot

be considered as substances of equivalent concern to PBTs or vPvBs.

We do believe that it is possible to establish a PNEC for MPs in both the aquatic, sediment and soil

compartments given that adverse effects are only observed at very high concentrations and

recommend research is undertaken to collate the necessary data for such toxicity thresholds to be

defined.

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1. Overview

An appropriate definition of what constitutes intentionally added Microplastics (MPs) as well as a

science-driven and evidence-based approach are essential elements to build an effective EU

Strategy for plastics, and define the related policy actions.

The definition should cover materials that contribute to aquatic litter and the potential associated

risks, which relate mainly to physical effects in aquatic life forms based on todays’ scientific

knowledge.

Substances of very high concern (SVHC) are materials that may pose serious impacts on human

health and/or in the environment. For environmental concerns, such substances are persistent,

bioaccumulative and toxic (i.e., PBT). Other substances that have equivalent concern have an

endocrine disrupting potential (https://echa.europa.eu/chemicals-in-our-life/which-chemicals-are-

of-concern/svhc ). According to REACH, substances that are considered PBTs or very persistent

and very bioaccumulative (vPvB) cannot have a “safe” concentration. This stems from the

properties of persistence and biomagnification. Even very low concentrations of a PBT/vPvB have

the potential to accumulate to sufficient levels in the consuming portions of the food web to illicit

toxicity. Therefore, a separate assessment is required (see Article 14(3)(d);

https://echa.europa.eu/management-of-pbt-vpvb-substances ; and

https://echa.europa.eu/documents/10162/13632/information_requirements_r11_en.pdf/a8cce2

3f-a65a-46d2-ac68-92fee1f9e54f ). While the vast majority of microplastics (i.e., excluding

biodegradable plastics) have environmentally persistent properties, there is sufficient information

to cast considerable doubts regarding their potential to bioaccumulate, resulting in a high toxicity

profile.

Cosmetics Europe and A.I.S.E. believe that the science regarding the impact of microplastics in the

environment is often used selectively and misrepresented. We would therefore like to highlight

several key areas of science related to the environmental risk assessment of potential

microplastics ingredients . Each area of science will be considered separately.

For the present document, we have defined plastic as:

synthetic water insoluble polymers that are moulded, extruded or physically manipulated into

various, solid forms which retain their defined shapes in their intended applications during their

use and disposal. Additional information on the definition are provided in the framework of the

Cosmetics Europe and A.I.S.E. reply to the ECHA’s call for evidence.

In conclusion, we elaborate on microplastics as being considered as substances of equivalent

concern (i.e., as PBTs/vPvB).

2. Relative Source Attribution

Cosmetics Europe and A.I.S.E. recognize the global concern regarding increasing pollution from

plastics in the oceans. However, regarding microplastics, it is noted that focusing the concern to

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primary microplastics potentially found in CPCPs and HCPs, as opposed to secondary microplastics,

is disproportional and scientifically not sound (Essel et al. 2015, Duis and Coors, 2016). Several

source attribution exercises have been conducted. For example, studies conducted by Eunomia,

the Technical University of Denmark and the Norwegian Environment Agency indicate that

cosmetics and personal care products account for an extremely small portion of primary

microplastic litter (Lassen et al. 2015, Sundt et al. 2015, EUNOMIA 2016). With regard detergents

AMEC estimated a potential total amount of about 0.02% of total release of Microplastics per year

(EUNOMIA, 2016). When secondary sources are considered (microplastics originating from the

fragmentation of larger plastic items by use, waste management or in the environment), the

contribution of CPCPs and HCPs to marine litter can be shown to be even smaller. For instance,

Gouin et al, 2015, estimated a CPCP contribution of primary microplastics to the North Sea marine

environment of 0.1% – 1.5%; Eunomia 2018 study considered by far HCPs as the smallest potential

contributor among known sources for MPs

Additionally, industry surveys conducted by Cosmetics Europe demonstrate that scoping studies

(commissioned by the European Commission and UNEP, for example) overestimated the

contribution of CPCPs to microplastic debris. Moreover, survey data indicate that leave-on plastic

ingredients, which are less likely to reach the marine environment than rinse-off plastic

ingredients, are used at low levels compared to rinse-off products (4-5 times less usage). Survey

data also show that voluntary industry restrictions have significantly reduced the use of plastic

ingredients in rinse-off products. A Cosmetics Europe membership survey found a significant

(82%) reduction in the use of plastic microbeads for exfoliating and cleansing purposes in wash-off

cosmetic and personal care products when comparing use in 2012 with use in 2015.

Despite the steady global increase of plastic debris in the oceans a case study on historical fish

samples from the Baltic Sea over three decades indicated no increase in plastic contamination in

biota (Beer et al. 2017).

Although there are multiple studies that have indicated the presence of microplastics in the

environment, none of these studies provide evidence for the origin of the particles (i.e. primary vs

secondary). In addition, many studies employed only visual separation and identification

techniques which are limited to a maximum size of 0.5-1mm and may also fail to discriminate

between plastic and inorganic particles (e.g. Hidalgo-Ruz et al. 2012, Lenz et al 2015). These

shortcomings lead to high uncertainties of microparticle exposure assessments because it may

have resulted in false positive and/or false negative results (Duis and Coors 2016, Ivleva et al.

2017).

Notwithstanding the above-mentioned limits of source apportionment and analytics there is no

doubt that microplastics are present in the environment. Budimir et al (2018) reported the

frequency of microplastics found in the gut of planktivorous fish collected within the Baltic sea

were small. Specifically, 1.8%, 0.9% and no particles were found in herring, sprat and three-

spined stickleback, respectively. However, in the presence of suspended particulate matter the

uptake of particles appear to be a common natural event and any associated effects occur likewise

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with natural material (Michel et al. 2014). The presence of microplastics in biota is due to the

ingestion of particles with food. Filter feeders are simultaneously ingesting synthetic and natural

microparticles and fish may easily confuse small microplastic particles with food. So far, laboratory

and field studies reported that over 160 different marine species ingest MP, including

invertebrates, reptiles, fish, birds and mammals (Lusher 2015). It is possible that many of these

observations, however, may have been tainted by poor sampling and analytical methods. For

example, Hermsen et al (2017) sampled 400 fish from the North Sea using a strict quality

assurance method and observed only 2 particles in one fish (sprat) in the intestine. Such strict

approaches for understanding concentrations in biota have not, in general, been utilized (e.g.

Lusher 2015, EFSA CONTAM, 2016). Hence, available records of microplastic occurrence in biota

should be interpreted with caution.

Generally, larger molecules are not bioavailable to organisms and it has not come to our attention

that microplastics have been found in non-gut tissues of vertebrates like fish or birds, while in

lower animals like bivalves, crustacean and gastropods microplastics were reported to reach the

digestive glands and other associated organs (e.g. hepatopancreas). It is hypothesized that

passage of large molecular weight molecules across biological membranes is unlikely (e.g.

Opperhuizen et al., 1987). Dimitrov et al. (2002) observed a drop in bioconcentration of chemicals

at a maximum cross-sectional diameter of about 1.5 nm and interpreted this as an indication of a

switch of the mechanism of uptake of chemicals into cells above this threshold from passive

diffusion to facilitated diffusion or active transport (e.g. by pinocytosis). The latter is rather

unlikely in the intestinal of higher animals. Further to that ECETOC (2005) concluded that the

potential of chemicals to bioaccumulate is likely driven by a combination of molecular mass, its

size and its octanol solubility. However, evidence suggests that once the molecular weight is in the

region of 700 - 1,100, depending on other factors, a reduced BCF may be expected. This finding is

well established today in many regulatory frameworks like e.g. REACH (R7c Guideline) and US EPA

(64:60194-60204) and such substances are therefore not considered bioaccumulative. Hence,

because of the size of microplastics it can generally be assumed that such particles are likely to be

excreted along with natural faeces.

3. Environmental Exposure via Wastewater Treatment Plants (WWTPs)

Potential MPs ingredients from PCPCs and HCPs may be released into the environment either

down-the-drain (rinse-off) or via solid wastes (leave-on). With regard PCPCs there is also some

evidence for leave-on products to be released after showering down-the drain. MPs that are

released down-the-drain into the aquatic and marine environment are subject to waste water

treatment via wastewater treatment plants (WWTPs) in many places around the world. Based on

the present knowledge, microplastic particles are not readily biodegradable and mineralization of

plastics appears to be a very slow process (e.g. Duis and Coors, 2016). Removal of such particles

during sewage treatment is therefore not a biological but rather a physico-/chemical process.

Several studies conducted in the US and Europe conclude that WWTPs remove solids and plastic

particles with a high degree of efficiency via skimming with fats and oils in the primary treatment

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phase and settling of activated sludge in the secondary treatment phase (e.g. Carr et al. 2016,

Mason et al 2016, Mintenig et al. 2017). An extensive study on 10 waste water treatment plants

conducted by the Danish Environmental Protection Agency covering the 26% of the Danish

wastewaters concluded that WWTPs removed >99% of microplastic (Vollertsen and Hansen,

2017). This efficiency has also been shown by earlier studies (e.g. Magnusson and Norèn, 2014,

Murphy et al 2016).

A recent study by Besseling et al. (2017) also investigated the fate of microplastic discharged to

aquatic environment via WWTP effluent. Figure 1 of the study illustrated that particle size and

density governs the rate at which particles are sedimented and that some particle sizes (such as

those used for exfoliation in skin care products, i.e., >100 um) may be fully sedimented within ~10

km. Hence, it is unlikely that microplastics emanating from the vast majority of WWTPs within the

EU will reach river termini and the marine environment.

Since microplastics are efficiently removed via WWTPs processes, another potential source is

runoff from lands fertilized with WWTP sludge. No adverse effects have been reported for soil-

dwelling organisms at concentrations found in the environment. This is not surprising since soil

organisms, such as earthworms, are adapted to engulf and process solids. The same can be said

for sediment-dwelling organisms. In ecotoxicology, a rule of thumb is that soil- and sediment-

dwelling organisms are more tolerant of substances than their water-column-dwelling

counterparts.

4. Environmental Exposure Modelling of Plastic Particles

First introduced in 1997 the EUSES model represents a quantitative assessment tool of the risks

posed by chemical substances that has been reworked in its second version by regulators in the EU

for the estimation of risks of single substances to man and the environment (Lijzen and Rikken,

2004). A tiered standard exposure model is embedded into EUSES, that enables estimations of

substance distribution and concentrations in several environmental compartments (“box model”).

The AMEC report used the EUSES model to estimate environmental concentrations of

microplastics. Van de Ment & Traas (2014) identified that EUSES has limitations to insoluble and

solid nano-particles and that it was not originally developed for this purpose. These limitations

apply similarly to microparticles such as microplastics. By applying the EUSES model, Amec was

already aware of these limitations and pronounced this in the respective chapter. However, the

outcome and quantitative dimensions of their assessment were presented in such a way that they

are interpreted as evidence for realistic exposure levels.

The following paragraphs present a review of the EUSES modelling approach used in Amec (2017)

and underlines the assumptions where CE and A.I.S.E. consider uncertainties have been

introduced to the modelling results. Alternative modelling approaches for particles in the

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environment are also proposed. Please consider that the following example relates to CPCPs;

however, similar consideration can be applied to HCPs.

Firstly, CPCP emissions have been over estimated as follows:

• 100% of leave on products are assumed to go down the drain. The real figure may be

significantly lower, as it is known a considerable amount of these products are removed

via the use of wipes and solid waste.

• The microplastics tonnage for both rinse off and leave on professional and consumer

use products has been assumed to be 8200 Tonnes/year across the EU as a model

input. This figure was a worst case estimate by Eunomia (2016) and is not based on

industry data. The Amec report states this figure is subject to high uncertainty.

Industry data provided by Cosmetics Europe, meanwhile, gives a maximum tonnage of

714 for rinse off and at most 1120 for leave on products. The combined CPCP tonnage

should therefore be at most 1834 tonnes/year which is 4.5 times lower than the above

tonnage.

• Per EUSES default the regional usage of CPCP (professional and consumer use) is

assumed to be 10% of all EU usage occurs in the region while the regional population is

only 2.6% of the EU population. This is very high; a typical approach would be to

assume 2.6% usage in the region, in line with the population fraction, which would give

a regional usage of 221 tonnes per year instead of 820 tonnes per year. Alternatively,

Price et al. (2010) presented an approach that couples regional population density and

country-specific usage statistics for a range of home and personal care products. From

these analyses they indicated that the above figures are indeed overly conservative.

From the inclusion of three personal care product categories, a realistic fraction of

personal care products for the default regional tonnage was determined by be 5.4% of

the total usage. If we apply this percentage to the more realistic Cosmetics Europe

tonnage value of 1834 tonnes/year, the regional usage would fall to 99 tonnes/year,

eight times lower than the value modelled by Amec.

• In order to support and to aid registrants of cosmetic ingredients to apply more realistic

exposure parameters, CE developed SPERCs (Specific Environmental Release

Categories). SPERCs exist for the wide-spread use of personal care products and are

recommended for a more realistic environmental exposure modelling of cosmetic

products (https://www.cosmeticseurope.eu/files/2314/9077/4925/2012-10-

25_CE_SPERC8a.1a-c-_Cosmetics_WDU__Determinants.pdf). While Amec used SPERCs

in the formulation the SPERCs were not consulted for wide spread uses.

• The local WWTP influent rate has been derived from the assumed 10% regional usage

and an additional safety factor of 4 is applied to give a WWTP influent rate of 4.5

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kg/day. This safety factor has been introduced by the TGD (2003) for wide spread uses

to account for local variations. For consumer products, it has long been acknowledged

that this approach would over-conservatively describe reality. Fox et al (2002) verified,

by a thorough analysis of boron freights in waste water treatment plants throughout

the EU, that this local variation cannot be maintained for consumer products. Hence, in

the TDG (B-tables) and REACH, this factor has been removed for personal and home

care products. If instead a 5.4% usage is applied without the safety factor, this gives a

WWTP influent rate of 0.6 kg/day, even while using the over estimated 8200 tonnage

value. Scaling by the safety factor would give an influent rate of 2.4 kg/day which is still

1.8 times lower than the 4.5 kg/d figure assumed in the EUSES modelling. If instead

we use the Cosmetics Europe tonnage value of 1834 tonnes/year to this calculation we

derive a WWTP influent rate of only 0.13 kg/d (which becomes 0.54 kg/d if we apply

the safety factor).

In conclusion, the WWTP influent rate applied in the EUSES model of 4.5 kg/d for PCCP

(professional and consumer use) falls to 0.13kg/d if more realistic assumptions are applied.

The associated predicted environmental concentrations due to the local WWTP discharge

assuming the worst-case emissions value assumed by Amec vs the more realistic value of

0.13kg/d is shown in Table 1. A significant reduction in concentrations are shown with the

maximum concentration, for freshwater sediment, falling from 290,000 to 8378 particles per

kg DW sediment. The surface water and marine water concentrations are now well below

values in the literature where physical effects on organisms have been observed.

Table 1: Particle based predicted environmental concentrations for 92% removal in WWTP as reported in Amec (2017) of worst case assumed emissions of 4.5 kg/d vs more realistic emissions estimate of 0.13kg/d, representative of the CE-based SPERC.

Particles per litre Particles per DW sediment

Professional and consumer use emissions value for WWTP influent

Surface water Marine water Freshwater sediment

Marine sediment

4.5 kg/d 4.3 0.43 290,000 29,000

0.13 kg/d 0.12 0.012 8,378 838

The EUSES model has been used to predict PECs of microplastics in both freshwater, marine water,

freshwater sediment and marine sediment and agricultural soil compartments. Based on the

physico-/chemical parameter applied, removal fractions between 53-92% were derived. However,

9

studies in WWTPs indicate that the actual removal rates during t sewage treatment can reach 99%

and higher (e.g. Vollertsen and Hansen, 2017). The fraction removed appear to be also dependent

on the form and size of the particles (Mintenig et al 2017). The realistic concentrations in surface

waters of primary microplastics after sewage treatment are therefore likely to be more than a

factor 8 lower, while in regions where sludge is applied to fields concentrations in agricultural soil

may be higher.

The freshwater concentrations assume a WWTP discharge to freshwater dilution and represent

concentrations immediately downstream of the discharge. Marine concentrations applied to a

WWTP discharge also represent concentrations in the immediate vicinity of the discharge. The

concentrations are, therefore, considered worst case.

The model also predicted highest concentrations due to emissions from down the drain

microplastics (including from CPCP) in freshwater sediments. Since the EUSES model does not

have the functionality to simulate the physics of particle deposition to river sediments or transport

in the river, the model assumptions had to be altered. In this case, to mimic materials that are

highly sorptive and that may settle with biological solids emanating from WWTPs, a high logKow

was assumed. While this assumption represents a pragmatic approach to estimating PECs in

sediments, it does not accurately represent sedimentation processes, such as modelled by the

NanoDUFLOW model (Besseling et al., 2017). The NanoDUFLOW model describes the transport

and sediment deposition of microplastics more accurately by taking into account the physics of

particle aggregation, together with instream advection which transports the particles down the

river. The model showed that the transport and sedimentation of MPs varies considerably with

particle size. Smaller particles (less than 100µm in diameter) are carried much further downstream

and undergo successive settling and resuspension in the water column. Particles larger than

100µm, meanwhile, were predicted to all be deposited within 1km of the discharge.

Given the likely range of particle size of MPs in WWTP effluent discharges, this alternative

modelling approach suggests that MPs will be transported and deposited at different downstream

locations leading to a much greater distribution down the length of a water body than

simplistically assumed in the Amec (2017) modelling study.

A key attribute within EUSES that drives regional concentration estimates is SimpleBox

(https://www.rivm.nl/en/Topics/S/Soil_and_water/SimpleBox). Recently, Meesters et al (2014)

modified SimpleBox to accommodate nanoparticles. Importantly, the distribution of these

materials in water, sediment and soils are altered considerably via the same processes identified

by Besseling et al (2017), aggregation and agglomeration. Hence, in order to appropriately

estimate environmental concentrations at the regional and local levels, it is imperative that state

of the art methods that include particle-particle interaction and their potential roles in fate and

transport by considered. The AMEC approach for modelling microplastics is inappropriate.

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While we have not identified a comparative marine modelling study, the fact that transport due to

tidal and drift processes has been neglected in the model suggests microplastics will in fact be

more widely distributed at lower concentrations that predicted by the EUSES model.

In summary, the EUSES modelling approach does not represent the state of the art and has over

predicted environmental concentrations for the following two main reasons:

• Emissions are over estimated using very conservative model assumptions and lack of

realism. Modelled PECs could, therefore, be at least two orders of magnitude too high.

• Deposition in the freshwater sediment is over estimated by neglecting aggregation with

suspended solids coupled with hydrodynamic transport processes which results in MPs

being more widely distributed along the receiving river or coastal area rather than settling

in the immediate vicinity of the WWTP discharge (as more accurately modelled via

NanoDUFLOW).

5. Environmental Hazard of Plastic Particles

In concept, the primary role of microplastics in adversely affecting aquatic biota is via their non-

nutritive contributions to biologically useful energy and physical effects such as an inflammatory

response (OSPAR Commission, 2017; Scherer et al., 2018). As a result, an organism’s homeostasis,

growth and reproduction may suffer. Figure 13 (shown below) from the OSPAR Commission

(2017) report aptly describes an adverse outcome pathway that can be tested for microplastics.

Since the end result corresponds to classic ecotoxicity endpoints, it is entirely possible to derive a

PNEC based on appropriate laboratory and/or field studies.

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However, in a recent evaluation of ecotoxicity studies with microplastics, Connors et al. (2017)

reviewed 71 published manuscripts for the applicability in assessing effects thresholds useful for

risk assessment. The authors concluded that few studies adequately characterized the test

materials and evaluated concentrations or mass loadings to derive adverse effects concentrations.

Indeed, many published studies report effects observed at much higher concentrations than those

observed in the environment, and are therefore unrealistic, especially when assessing physical

effects. The consequence is that observed physical effects may be overestimated (Beer et al.,

2017; Burton et al., 2017; Koelmans et al., 2017; Scherer et al., 2018).

In addition, natural particulate matter has been shown to also cause adverse effects similar to

microplastic (e.g. Kirk 1992, Capper 2006, Gordon and Palmer 2015, Ogonowski et al. 2016;

Scherer et al., 2018). When reporting on effects against microplastic particles the majority of

studies fail to indicate natural particles as positive controls (e.g. Cole et al; 2015, Rehse et al 2016,

Sussarellu et al, 2016). Vice versa, in studies where the natural habitat was considered in the test

design the effects were usually not significant, either at unrealistic high concentrations or could

not be attributed to the presence of microparticles alone (e.g. Green 2016; Ogonowski et al.

2016).

Derivation of PNECs or water quality criteria for solids is not unprecedented. For instance, Smit et

al (2008) created species sensitivity distributions for two types of suspended clays (barite,

bentonite; see Figure 1). The hazard concentrations affecting the 5th percentile of organisms were

17.9 and 7.6 mg/L for barite and bentonite, respectively. These can serve as a suitable benchmark

for comparing PNECs based solely on data from lab studies with microplastics.

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Figure 1. Species sensitivity distribution constructed for barite and bentonite (Smit et al, 2008).

Many of the cited studies are following non-standardized test and report on physiological effects

of which their environmental relevance is not fully understood (e.g. von Moos et al. 2012, Oliveira

et al. 2013, Wright et al 2013, Rochman et al. 2014, Cole et al. 2015, van Cauwenberghe et al

2015, Paul-Pont et al. 2016, Lu et al. 2016, Sussarellu et al 2016; Connors et al., 2017). Connors et

al (2017) provided a list of 9 factors that should be considered when evaluating the quality of

ecotoxicity studies with microplastics. When tests were sufficient for establishing effect values on

standardized ecotoxicity endpoints (mortality, growth, reproduction), the concentrations required

to cause adverse effects were typically several orders of magnitude greater than concentrations

monitored in the environment (Booth et al., 2013; Lenz et al., 2015; Raimondo et al., 2007;

Rochman et al., 2013; Sjollema et al., 2016). Additionally, many studies measure concentrations

based on mass (e.g., mg/L) or surface area (number/km2). These units add large uncertainty to

actual organism exposures to these diverse particles (Burton et al., 2017), and very few studies

discuss or verify if the microplastic test solutions remain stable and homogenous. Experiments

with microplastics often lack a thorough characterization (e.g., polymer type, charge, size). There

is therefore a need for additional toxicity testing at environmentally relevant concentrations and

specific research to better assess effects of microplastics on living organisms.

Despite these deficits, it is possible to obtain sufficient acute toxicity information to derive a PNEC

of MP. To enable this exercise, it was assumed that toxicity will not be affected by polymer

identity, shape, charge, functionality, or density. A tentative PNEC is derived below for

microplastics less than 100 µm in size (largest size in any one dimension) using published data,

summarized in the table 2. Data are available for all three taxa groups – fish, invertebrates

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(daphnids) and algae. However, acute data (EC50) are only available for daphnids, whereas

chronic data (NOEC) are available for algae and fish. According to REACH Guidance R.10-4

(Assessment factors to derive PNECaquatic), an assessment factor of 1000 requires an acute data

point for algae, daphnid and fish. An assessment factor of 4 is used to convert the algal NOEC into

an EC50, whereas an assessment factor of 10 is used to derive an LC50 for fish from the chronic

NOEC. Based on this acute dataset, the PNEC ranged from 0.057 – 0.072 mg/L, considering the

two different size ranges (1-4 um and <50 um) and affected taxa (Daphnia magna and Salmo

trutta), respectively. Considering the availability of the chronic data point for Salmo trutta, an

assessment factor of 100 is used to derive a final PNEC of 0.072 mg/L, or 1 particle/L.

Table 2: Summary of published toxicity endpoints for microplastics less than 100 µm in size

Particle Size

(um)c

Endp

oint

Effect Species

Group

Species Effect concentration Citation

particles/mL

mg/L

Acute

PE 1-4 (1) mortality

EC50 Invert Daphnia magna

110,000,000

57.43a {Rehse, 2016 #24}

PE 90-106 (100)

mortality

EC50 Invert Daphnia magna

>800 >400a {Rehse, 2016 #24}

PS 6 growth

NOEC Algae Dunaliella tertiotlecta

2,100,000a

250a Sjollema et al. 2016. Aq Tox.

EC50* 8,400,000

1000

PS <50 (cryomil

led)

mortality

LC50**

Fish (embryo) OECD 212

Salmo trutta f. fario

1000 72 Schmieg et al. 2017, SETAC Europe poster (TH197)

Chronic

PS <50 (cryomil

led)

mortality

NOEC Fish (embryo) OECD 212

Salmo trutta f. fario

100a 7.2b Schmieg et al. 2017, SETAC Europe poster (TH197)

aReported concentration. Other value was calculated based on particle size and density.

bAssumes all particles are 50 µm in size

*EC50 estimated from NOEC value based on an AF of 4 from Sjollema et al, 2016

**LC50 estimated from NOEC value from on an AF of 10 from Schmieg et al., 2017

14

Measured EC50 values for Daphnia magna for microplastics were approximately factors of 3-7.5

above the HC5 for bentonite and barite clays, respectively. For both clay types, zooplankton were

the most sensitive taxa. Hence, there is sufficient data to suggest that zooplankton, Daphnia, may

be the most sensitive taxa. Given the rather close range of the limited data available for

microplastic ecotoxicity, it appears that treating microplastics as solids may be appropriate. If so,

it seems prudent that the use of water quality criteria for suspended solids may serve as an initial

assessment to the potency of these solid contaminants in aquatic systems.

The diagram from OSPAR (2017) describes a pathway for ecotoxicity due non-nutritive-based

effects. Current standard acute toxicity tests do not include feeding and, therefore, cannot directly

measure non-nutritive-based effects. In order to design studies to address these concerns, ratios

of food items vs microplastics need to be considered. For example, Burton (2017) describes the

concentrations of algae vs. the concentration of microplastics in Lake Erie (the most microplastic

contaminated water ever recorded) as the ratio of 1 to 3 microplastic particles per 300 to 700

liters vs. ~10,000 to 10 million algae per liter. To design a PNEC for non-nutritive aspects of

microplastics and other particles will require a design that incorporates food vs. microplastic

encounter rates that overlaps ratios found in the environment.

6. Bioaccumulation

Chemicals that are considered bioaccumulative indicate a bioconcentration factor (BCF) of greater

than 2000. Chemicals that are very bioaccumulative (vB) have a BCF value of more than 5000. For

such chemicals, diet is the primary source for accumulation within the biota. The bioaccumulation

classification is a chemically-based classification system whereby partitioning into biological lipids

his linearly related to their octanol-water partition coefficient (logKow) until the materials are too

large to efficiently move across cellular membranes (e.g., >logKow of 7) and into storage

fats/lipids. These materials may biomagnify up the food chain, potentially harming ecological

consumers, including humans. Due to the solid nature of microplastics, there is considerable doubt

whether such materials should fall under a chemically-based classification system. According to a

recent summary publication by OSPAR (2017), fish sampled from the North and Baltic Seas showed

“no signs of bioaccumulation nor biomagnification as microplastics were found within the

intestines, from which most contents are egested”. Hence, microplastics should be considered as

solids that do not conform to regulatory and classification processes meant for materials with a

high logKow.

7. Trojan Horse Effect (Microplastics as Vectors of POPs or PBTs)

Published studies provide little proof that plastics are responsible for observed contamination of

organisms by persistent organic pollutants (POPs) and many of the laboratory studies that have

employed environmentally unrealistic test gradients or media , hence can only provide crude

15

assessments of potential adverse effects (Ziccardi et al., 2016). Due to their particulate nature, the

most relevant exposure pathway for environmental organisms is via ingestion. However, this is

dependent upon particle size and the size(s) of the organism. Even so, accumulation to sufficiently

high levels to transfer their contents to secondary consumers is based more on potential than

evidence. For examples, 2 key references provided within the AMEC report (Setala et al., 2014;

Farrell and Nelson 2013) utilize exposure concentrations that are perhaps relevant for industrial-

scale spills than monitored concentrations found in marine and freshwaters, including municipal

wastewaters. Ingestion and transfer of microplastics (10 um diameter polystyrene fluorescent

sphere) within the planktonic food web was based upon exposures 1000, 2000 and 10,000

microspheres/mL (Setala et al., 2014)). Since most ecotoxicity values are expressed per liter, these

exposures translate into: 1 million/L, 2 million/L, and 10 million/L. While these excessive

exposures clearly indicated that there is potential trophic transfer, when compared to municipal

effluent concentrations (e.g., <0.1 to 5 particles/liter) their use for defending biomagnification is

highly dubious. Similarly, Farrell and Nelson (2013) illustrated that mussels (Mytilus edulis)

exposed to ~411 million (0.5 um particles) per 400 mL and then allowed to serve as prey for the

crab (Carcinus maenas (L.)) did illustrate transfer of microplastics from filter feeders to primary

consumers – however, such exposure levels are not indicative of reality. Simply said, the 411

million particles/400 mL translates into 1.03 billion particles/liter. Hence, such a crude experiment

should be considered with great caution and not adopted as an outright verification of

biomagnification. It is not surprising that a transfer under such artificial laboratory conditions can

be shown (Browne et al., 2008; Chua et al., 2014; Wardrop et al., 2016). However, these results

need to be put in context. An important caveat for particulate-based accumulation vs. chemical is

that the primary loss mechanism is egestion instead of a combination of egestion and

biotransformation. Several authors have observed less transfer from plastics than from other

more abundant and naturally occurring particles (e.g. sediment), indicating that the transfer of

contaminants from plastic is not significant (Beckingham and Ghosh, 2017; Browne et al., 2008).

Additionally, Herzke et al. (2016) indicated in a study conducted in Norway on Northern Fulmars

that bioaccumulation of POPs was not proportional with quantity of plastic ingested, an

observation that contradicts the hypothesis that plastic acted as a carrier of POPs or PBTs.

The physical/chemical properties of plastics are not conducive for both desorption and transfer to

biota (Gouin et al., 2011; Herzke et al., 2016; Koelmans et al., 2013). Gouin et al (2011) was the

first to model the potential contributions of microplastics as vectors of PBTs into organisms that

feed on microplastics. Their thermodynamically-based modelling clearly showed that for materials

up to a logKow 6.5, the likelihood of the vector concept was low. Even so, they outlined data

needs to fully test this hypothesis, including investigations of gut contents and egestion rates of

various microplastics.

Such suggestions were tested by Besseling et al (2013) via bioaccumulation experiments with

lugworms exposed to PCBs in sediments and as sorbed onto microplastics. While a

bioaccumulative increase in 3 PCB congeners was observed, it should be noted that the

relationship of the contaminants with microplastics in sediments with sediment-dwelling

16

organisms will be affected by chemical equilibria. That is, contaminants can move between

microplastics, sediments, water and organisms. Addressing this issue, Koelmans et al (2013)

specifically modelled the potential of microplastics as vectors of POPs. Their analysis confirmed

conclusions by Gouin et al (2011), indicating that microplastics do not increase the uptake of

persistent, lipophilic organics beyond that of other factors found in sediments. The lack of

increased biomagnification of PBTs in aquatic biota collected in the field verify the modelling of

Gouin and Koelmans.

Further if this were not the case, then tissue observations of POPs and PBTs (e.g., DDT, PCBs)

should have increased significantly across all geographies and water bodies that received micro-

and macroplastics. Such observations have not been observed with declining or plateauing trends

being the norm (Braune and Mallory, 2017; Campillo et al., 2017; McGoldrick and Murphy, 2016;

Riget et al., 2016; West et al., 2017).

A critical review and reinterpretation of all the empirical research regarding this hypothesis was

conducted by Koelmans et al (2016). When the data was reinterpreted to test the same

hypothesis, all data became consistent in demonstrating microplastic ingestion would not likely

increase internal body exposure, hence increasing the risk of PBTs/POPs in the marine

environment. Hence, the state of the science does not support microplastics to serve as a unique

hazard by exacerbating the uptake (i.e., vector) of PBTs and/or POPs into biota (OSPAR

Commission 2017).

In addition, there is a paucity of evidence suggesting that microplastics can biomagnify, as verified

by the OSPAR Commission (2017). Given the unrealistic laboratory experiments that do illustrate

potential for trophic transfer, but were insufficient to provide quantitative evidence verified via

field monitored organisms, microplastics as particles should no longer be considered as carriers of

substances with either B or vB characteristics.

8. Environmental Risk of potential Microplastics Ingredients

When establishing whether potential microplastics ingredients pose an unacceptable

environmental risk it is essential to compare the concentration of microplastics in the environment

with the concentrations that cause an observed adverse effect. However, in both, the exposure

and the effects assessment there currently exist a high level of uncertainty and ambiguity to

enable a final conclusion of a real risk arising from the presence of plastic microparticles in the

aquatic environment.

For example, a high uncertainty is presented in particle counting (Filella 2015) and lack of

standardization on reporting and analytics (Ivleva, et al. 2017). In addition, unsuitable exposure

models and approaches have been applied to estimate environmental concentrations (e.g. EUSES).

We suggest the use of SimpleBox4.0-nano and NanoDUFlow as two applicable modelling

approaches to estimate which compartments microplastics will partition and the rate of settling

due to particulate-associated properties, such as aggregation and agglomeration. Such inclusion

17

will enable the ability to assess microparticles and microplastics with diverse phys/chem

properties.

Ambiguity of a risk assessment on microplastic is also given by the fact that many non-

standardized adverse effects are reported in the literature. They are generally not interpreted in

separation of naturally occurring particles and their environmental relevance is not yet

understood. Therefore, Koelmans et al. (2017) called for newly defined parameters in the risk

assessment of particulate plastic material. The authors suggest a new expert knowledge elicitation

that assesses ecologically relevant parameters and known particle- and species specific adverse

effects to better understand and develop a realistic risk assessment framework rationale.

9. Migration of additives from microplastic

The AMEC report estimates the potential environmental concentrations of additives from

microplastics. The calculations are based on general additive contents ranging from 0.1-1% and a

microplastic removal rate of 53% during sewage treatment. Although the AMEC report

acknowledges that further physico-chemical properties (e.g. the LogKow, vapor pressure,

temperature) can significantly influence the flux (OECD 2009) this was not considered in the

concentration estimations. Such simplistic deterministic diffusion models may lead to a significant

overestimation of real diffusion coefficients (e.g. Pocas et al., 2008, Welle, F. 2013).

Cosmetics Europe and A.I.S.E. do not challenge the fact that additive leaching can occur from

microplastics. It is acknowledged, that most additives are not chemically bound to the polymer

and, hence, are principally prone to migration. However, similar to the source apportionment of

primary microplastic vs. secondary microplastic, the dimensions of additive leaching appear

disproportionate in the AMEC considerations. Secondary microplastics must be similarly

considered as potential sources since much of the plastic litter entering the oceans consists of thin

foils and plastic bags that undergo migration of additives in similar ways as assumed for primary

microplastics. In addition, Cosmetics Europe and A.I.S.E. consider the generic presentation on

additive leaching too simplistic. There is a variety of different chemistries used in polymers that

are broadly categorized to functional additives, colorants, fillers or reinforcements (Hahladakis et

al 2018). Many of these additives can be considered inert (e.g. inorganic fillers like CaCO3, clay,

mica, glimmer) or immobile (e.g. reinforcements like glass or carbon fibres) (OECD, 2009). Some

functional additives like plasticizers, antioxidants, heat stabilizers, flame retardants or slip agents

were reviewed to be the most common additives recovered from the environment. Among them

are phthalates, BPA, nonylphenols and brominated flame retardants (BFR) (Hermabessiere et al.,

2017).

Common molecular weights of plastics additives were estimated to be in the range of 200-2000

g/mol. A thorough understanding of additive migration is derived from additives used in plastics

with food contact (Hahladakis et al 2018). High temperatures and non-polar solvents tend to

18

facilitate migration of many additives from plastic polymers (e.g. Tawfik and Huyghebaert, 1998,

Xu et al, 2010). According to Hermabessiere et al (2017) the most common additives indicated

LogPow between 1.6 and 11.2 (median ~4.6). Due to their vast difference in lipophilicity they may,

therefore, not strictly follow the simplistic picture as reported by Amec. It appears difficult to

conclude whether a particular additive has a higher migration potential than another as it is

dependent also on the polymeric form that surrounds it. Therefore, it must be concluded that a

differentiated view must be applied to each additive when considering its migration into the

environment.

Whether or not a plastic additive is taken up by organisms has been investigated in laboratory

experiments. It appears that depending on the physico-chemical properties of the plastic additive

they may or may not have influence on organisms especially at higher plastic concentrations in the

g/l range (e.g. Lithner 2009, 2012). Experiments with food grade polyethylene (PE) plastic bags for

example indicated that not only the chemistry of the plastic material (here PE) but the

manufacturer also plays a role (Hamlin et al. 2015).

Aside the leaching from plastic polymers the presence of plastic additives in the environment can

result from multiple factors, such as e.g. industrial and municipal wastewater, atmospheric

deposition, runoff and river transport resulting from application of sewage sludge in agriculture.

Because supplementation with additives fulfil different and specific functions in each polymer,

each of the substances needs to be assessed separately. Additives are already addressed under

common regulations like the REACH Regulation. In addition, some of have been restricted as

chemicals of concern, e.g. Hexabromocyclododecane (HBCB) under the Stockholm Convention

(2016) or are under suspicion, e.g. some Phthalic acid esters (PAH) are suspected endocrine

disruptors (e.g. Oehlmann et al 2009). Hence, regulatory pathways to address and regulate risks

from additives in polymers exist already.

In summary, Cosmetics Europe and A.I.S.E. conclusions on plastic additives are as follows:

i) Compared to other plastic sources, additives from primary microplastics potentially

used as ingredients are of marginal impact to the environment (cf.see chapter 2).

Other sources of the respective chemicals need to be considered thoroughly for

environmental risk assessments.

ii) Estimations of additive migration from plastic particles must consider a substance

specific analysis in order to conclude on its environmental risk potential

individually. A generic exposure assumption as delivered by the AMEC report is

insufficient.

iii) Regulations for additives in plastics exist and should cover environmental risk

management sufficiently (e.g. REACH)

19

10. Are microplastics substances of equivalent concern?

Microplastics are found in the environment. While it is acknowledged that microplastics could

satisfy persistence criteria, it is quite clear that there is sufficient data regarding their lack of

bioaccumulation, insufficiency to serve as vectors for PBTs/POPs and the agreement with a solids-

based PNEC assessment - to cast great doubt the applicability of considering microplastics as PBTs.

That is, there is sufficient information to negate any B and T criteria. Consequently, microplastics

cannot be considered as substances of equivalent concern.

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