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Environmental fate & pathways

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General introduction to chapters 4, 5, 6 and 7.


Under Regulation 793/93/CEE, an extensive risk assessment on Zinc and 5 zinc compounds (ZnO, ZnCl2, ZnSO4, Zn orthophosphate and zinc distearate) has been recently prepared by the Dutch authorities for the EU. The risk assessment report (RAR) on these 6 zinc substances has been recently published (ECB 2008).

Since these RARs were the result of intensive discussions between all stakeholders, and were approved by experts from all the member states; since they provide a recent review of the available evidence on zinc and zinc compounds (the file was closed in September 2006), they will be used as the main reference for this chemical safety report.

In this chemical safety report, the information, data and conclusions of the RAR will be summarised, focusing on the principles applied, the assumptions made and the conclusions. Where available and relevant, new information and data will be included and discussed.


General remarks on the chapter on environmental fate properties.


Zinc is a natural element, which is essential for all living organisms. It occurs in the metallic state, or as zinc compound, with one valency state (Zn++). All environmental concentration data are expressed as “Zn”, while toxicity is caused by the Zn++ion. For this reason, the sections on human toxicity and ecotoxicity are applicable to all zinc compounds, from which zinc ions are released into the environment. Some zinc compounds have however very low solubility and will therefore not release zinc ions; this strongly decreases their potential (eco-)toxicity. As a consequence, distinction for toxicity and ecotoxicity is being made between zinc compounds, as a function of their solubility (see chapters 5 and 7).

For checking the potential of metal substances to release ions in the environment, a specific test, the transformation/dissolution (T/D) test is used. For metallic zinc and some of the zinc compounds, this test has been performed. If applicable, the results of such T/D test are discussed in section 4.6. (data in IUCLID section 5.6.).

The issue of degradation (section 4.1.) is not applicable to inorganic compounds. However, the speciation of zinc in the environment compartments is relevant and is discussed under section 4.2.

When zinc ions are formed in the environment, they will further interact with the environmental matrix and biota. As such, the concentration of zinc ions that is available to organisms, the bioavailable fraction, will depend on processes like dissolution, absorption, precipitation, complexation, inclusion into (soil) matrix, etc. These processes are defining the fate of zinc in the environment and, ultimately, its ecotoxic potential. This has been recognised e.g. in the guidance to the CLP regulation 1272/2008 (metals annex):“Environmental transformation of one species of a metal to another species of the same does not constitute degradation as applied to organic compounds and may increase or decrease the availability and bioavailability of the toxic species. However as a result of naturally occurring geochemical processes metal ions can partition from the water column. Data on water column residence time, the processes involved at the water – sediment interface (i. e. deposition and re-mobilisation) are fairly extensive, but have not been integrated into a meaningful database. Nevertheless, using the principles and assumptions discussed above in Section IV.1, it maybe possible to incorporate this approach into classification.“


In the water, the bioavailability of zinc through interaction with components of the water and biota has been studied in detail in the zinc RA (ECB 2008). This has resulted in an approach for quantifying zinc bioavailability into risk assessment. The ultimate fate of zinc in water (in the water column) is assessed via the “unit world model”, that can quantify the “removal from the water column” of the zinc species. As such, it is shown that zinc (ions) brought into water will be rapidly removed from the water column (>70% removal within 28days). This phenomenon is described in section 4.6. (data in IUCLID 5.6), and is considered for classification.


In sediment, zinc binds to the sulphide fraction to form insoluble ZnS. As such, zinc is not bioavailable anymore to organisms. This has been discussed in the EU RA (ECB 2008), and has resulted in an approach for quantifying zinc bioavailability into risk assessment. Based on experimental data, a default conservative bioavailability factor of 0.5 was proposed in the RA. This approach can be refined when the relevant data on sulphide and Zn in sediment are available. Due to the insolubility of the ZnS (K=9.2 x 10-25) zinc will be sequestered in the (anaerobioc) sediments, and the re-mobilisation of zinc ions into the water column will be prevented. This is also quantified in the unit world model, see section 4.6.


In soil, short-term interaction of zinc ions upon spiking, and long term interactions (“ageing”) have been extensively discussed in the zinc RA (ECB 2008). This has resulted in an approach for quantifying zinc bioavailability into risk assessment. Based on experimental data, a general ageing factor of 3 was derived in the RA; according to soil type, the bio-availability of zinc can be further determined, when the relevant data on e.g. pH, CEC are available.

Specific remarks on the fate of ZnO nano: see old CSR, to review

Sources and main transfer pathways of ZnO nanoparticles (ZnO-NP) in the environment.

Main potential sources of ZnO-NP are: industrial emissions to water, and emissions resulting from the use of ZnO-NP-containing consumer products to (mainly) household waste waters. 

Industrial emissions of ZnO-NP are limited, since most processes are in ultra-clean environment and often dry, without waste waters. In this section, we will focus on the potential emissions resulting from the use of ZnO-NP in consumer products, e.g. cosmetics, paints, mastics/sealants and other household products. The use of these consumer product may lead to emission in the household waste waters. These are collected in sewage treatment plants (STP), where the ZnO-NP translate to the sludge fraction. When these sludges are used for fertiliser or biosolid in agriculture, they are a potential source of ZnO-NP to agricultural soil.

The household wastewater – sewage sludge pathway has been identified as a potentially main pathway for environmental exposure to ZnO-NPs. Lombi et al.(2012) studied the transformation processes of 3 ZnO-NP during anaerobic digestion processes of wastewater, as well as postprocessing (composting/stockpiling). It was observed that all Zn forms (“native” zinc and ZnO-NP-Zn) were rapidly converted to sulphides in all treatments. The composting process resulted in a shift of the speciation towards Fe oxy/hydroxides, and Zn complexed by phosphate. Still the speciation of zinc was the same in all treatments. The authors concluded that “at least for the materials tested, the risk assessment of ZnO-NP through this exposure pathway can rely on the significant knowledge already available in regard to other “conventional” forms of zinc present in sewage sludge” (Lombi et al.2012).


Fate and behaviour of ZnO nanoparticles (ZnO-NP) in the environment.

Schultz et al.(2015) considered the following range of physical and chemical transformations that (ZnO-)NP undergo in the environment:

·      Aggregation[1](we will consider “homo-aggregation, i.e. between ZnO-NP only)

·      Dissolution

·      Interaction with natural organic matter

·      Sulfidation phosphidization

Schultz et al.(2015) in general concluded that exposures to pristine, well-dispersed nanomaterials will occur only rarely if ever in nature. It is e.g. well recognised that NP that enter wastewater facilities will tend to aggregate with other mineral and organic components of the wastewater, resulting in associations with other solids and will not remain as dispersed nanosize suspensions (Batley et al.2013). The latter conclusion also leads to the problem of measuring environmental levels of NP in nature. Such measurements are indeed confounded by the co-existence in the environment of natural nanoparticles and colloids.

Upon release in the aquatic (or terrestrial) environment, ZnO-NPs readily aggregate and/or dissolve into soluble Zn2+ ions. Both processes are working in opposite directions: higher aggregation rate results in lower dissolution rate, and vice versa (Petosa et al.2011, Majedi et al.2014). Still, under environmentally relevant conditions, ZnO-NPs finally dissolve quickly, e.g. complete dissolution has been observed within 48hrs in Daphnia magna test setup (Adam et al.2014, 2015); after very rapid aggregation of the NPs (within 1-2 hrs, aggregates of 1,8µm – 1,9 µm size at pH 7.6 and 6.1, were observed respectively) Odzak et al.2014 observed rapid dissolution up to > 50%, that did not change afterwards.

In water, both the aggregation and dissolution of ZnO-NP are influenced by many different factors. As main factors are identified:

-       ionic strength and composition of the medium

High ionic strength promotes aggregation because it reduces inter-particle repulsion caused by a reduction in surface charges (Thwala 2013). Increased solution pH, H(PO4)2, Ca2+and Mg2+concentration reduced the dissolution of Zn2+ions from NP (Li et al.2013). Odzak et al.(2017) observed in natural waters (river Rhine, Lakes Greifen and Luzern) a rapid aggregation of NPs that was higher at higher pH and ionic strength; the difference however quickly disappeared and after 1 day only relatively few Zn-particulates were still present. The decrease of particulate Zn could be due to sedimentation or to dissolution; in the waters with low ionic strength, a large fraction of the NP was dissolved, while under conditions of high ionic strength, particle sedimentation occurred more efficiently (Odzak et al.2017). Majedi et al.(2014) concluded that the concentration of organic acids and pH were the most significant factors influencing aggregation and dissolution of NP in water. Stabilised NP do aggregate and precipitate notably when the pH is near their zero point of charge pHzpc (Peng et al.2016).

More specifically, it was shown that organic phosphates make NP dissolve more rapidly than inorganic Phosphate. Also in the presence of phosphate, decreasing pH results in higher dissolution of NP (Feng et al.2016). Polyphosphates (P2, P3, P6) increased the dissolution of NP, while orthophosphate (P1) decreased it (Wan et al.2017). These results may be especially relevant in waste waters carrying these P-forms.

-       presence of Humic Acids

Humic acids decreased the toxicity of NP to algae because they prevented the adhesion of the NP on organisms due to increased electrostatic repulsion (Tang et al.2015). The “coating” of NP with Humic Acids also seemed to stabilise the NP (Akhil and Khan 2016). The HA adsorbed ZnO-NPs were highly stable. HA reduced the photocatalytic activity of ZnO and at the same time increased the photostability of ZnO (Akhil et al.2015). In studying the interactions between NP and humic acids it is important to apply relevant environmental concentrations of both the ZnO and the HA. This is not always the case, leading to irrelevant results.

-       sulfidation

At ambient temperatures, sulfidation (ZnO -> Zn2+ -> ZnS) occurs rapidly when sufficient sulphide is present, like in sediments. Ma et al.(2013) observed complete sulfidation within 5 days, and the formation of ZnS NP forms. Also in sewage treatment plants, ZnO-NP converted largely to ZnS. (Brunetti et al.2015). This phenomenon is important in the environment, because Zn2+ions are bound into very stable and insoluble ZnS, which has very low bioavailability for uptake by organisms.

-       Surface treatment

Generally speaking, the coating for ZnO is never a full coating and as much as it does modify some characteristics of the pristine material, zinc oxide NP properties are still there. Surface treatment of NP influences their stability. Highest dissolution is observed with non-coated particles (Merdzan et al.2014). Surface treatment can influence the stability of the NP form: polymer-coated NP showed less aggregation and higher transport potential (Petosa et al.2011).

-       Surfactants

The presence of surfactants may alter the behaviour and fate of NP in the natural water environment. Experiments with an anionic sodium dodecyl sulfate (SDS) and a non-ionic nonylphenol ethoxylate (NP-9) showed effects on the aggregation and sedimentation of a 20nm ZnO-NP in different types of water. As the surfactant concentration increased from 0 to 0.030%, SDS reduced the zeta potential of the ZnO-NP, thus indicating lower stability, while NP-9 did not affect the zeta potential. SDS and NP-9 reduced significantly the aggregate size of ZnO-NP. Surfactants were found to reduce the aggregation and sedimentation of ZnO-NP in six natural water matrices in different degrees. The authors calculated that with the presence of 0.030% SDS in tap water, maximum reduction rates of aggregate size and sedimentation were recorded as 69.54% and 26.69%, respectively (Li et al.2017).

-       temperature

Increasing temperature may reduce the surface charge of NP, but the effect is considered less important than e.g. the effect of ionic strength (Majedi et al.2014).

-       light

Interactions of light with the dissolution of nanoparticles in natural waters is complex. Indirectly, UV light can cause the decomposition of dissolved organic carbon that stabilises NP under natural conditions (Odzak et al 2017). The same authors observed the major dissolution of ZnO NP under different light regimes and low pH, but a tendency for aggregation at higher pH and ionic strength.

In summary, influencing factors on NP aggregation and dissolution are:

·      Increasing ionic strength, pH, Ca2+, Mg2+, ortho-P: more aggregation

·      polyP, lowering pH, ionic strength, increases dissolution

·      humic acids stabilise NP, thus anticipated less dissolution and more aggregation

·      sulfidation: Zn2+dissolved from NP will in presence of S (e.g. in sediment) be bound into stable (and non-bioavailable) ZnS 

·      coatings, surfactants: influence stability and therefore aggregation/dissolution in different ways

·      light, temperature: some weaker influences


Waste water treatment plants

As indicated above, NP are transformed in STP in the same way as other forms of zinc entering the STP (Lombi et al.2012). Rapid transformation to ZnS was followed by shift in speciation to Fe-oxy/hydroxides and phosphates in the postprocessing of the sludges (Lombi et al.(2012). This was confirmed by Ma et al.(2013).

In bioreactor simulations, Musee et al.(2014) observed that NP had high affinities for the sewage sludge and aggregated under typical wastewater conditions. The zinc translated mainly to the sludge fraction, leading to removal of the NP from the wastewater by biosorption and biosolid settling mechanisms (Musee et al.2014).


In soils, the application of sewage sludge is the main potential input source of NP. Much like in the water, the characteristics of the NP (e.g.; size, shape, surface charge), and the receiving soil (pH, ionic strength, organic matter...) will influence the physical and chemical processes resulting in NP dissolution, and/or aggregation. The main difference between the water environment and the soil is the presence of the soil’s solid phase. Soil components like clay particles and humic molecules will influence the association of NP with the solid phase; Humic molecules may desorb into the pore water and stabilise the NP there. Still, pH, and ionic strength are the main factors defining the sorption of NP to the soil. Higher ionic strength decreases the repulsive forces between particles and between particles and the soil, resulting in increased aggregation and sorption (Tourinho et al.2012).

Trapping of particles in soil pores (pore straining) is another specific, physical soil process, influencing retention of particles in soils. The importance of this process depends on particle size and pore size distribution. In general, mobility of NP s in soil decreases with increasing particle size, because larger particles are more likely to be retained by pore straining (Darlington et al.2009). Logically, aggregation tends to increase the importance of pore straining.


[1]Aggregation is defined as the association of primary particles by strong binding, whereas agglomeration is defined as association by weak bonding caused by Vander Waals forces. However, in most papers, the term aggregation is used for situations where possibly only agglomeration has occurred. We will refer further to “aggregation”.