Environmental toxicology deals with the negative effects of the exposure to chemicals we regard as pollutants (or contaminants/toxicants). Environmental toxicants receive a lot of media attention, but many critical details are getting lost or are easily forgotten. The clip"You, Me, DDT" shows the discovery of the grandson of the works of his grandfather, the Swiss inventor of the insecticide DDT, Paul Hermann Müller, who received the Nobel Prize for Medicine for that in 1948 (see also Section 1.3). The clip "Stop the POPs" interviews (seemingly) common people about one of the most heavily regulated group of pollutants.
Organisms, including humans, have always been exposed to chemicals in the environment and rely on many of these chemicals as nutrients. Volcanoes, flooding of acid sulfur lakes, and forest fires have caused widespread contamination episodes. Organisms are also in many cases directly or indirectly involved in the fate and distribution of undesirable chemicals in the environment. Many naturally occurring chemicals are toxicants already (see also Section 1.3), think for example about:
local arsenic or mercury hotspots in the Earth’s crust, contaminating water pumps or rice irrigation fields;
plant-based defence chemicals such as alkaloids, morphine in poppy seeds, juglone from black walnut trees);
fungal toxins, such as mycotoxins threatening grain storage depots after harvests;
bacterial toxins, such as the botulinum toxin, a neurotoxic protein produced by the bacterium Clostridium botulinum which is the most acutely lethal toxin known at ~10 ng/kg body weight when inhaled;
phycotoxins, produced by algae, in mass algal blooms or those that may end up at dangerous levels in shell food;
zootoxins in animals, such as venom of snakes and defensive toxins on the skin of amphibians.
Human activities have had an enormous impact on the increased exposure to natural chemicals as a result of, for example, the mining and use of metals, salts and fossil fuels from geological resources. This is for example the case for many metals, nutrients such as nitrate, and organic chemicals present in fossil fuels. Additionally, the industrial synthesis and use of organic chemicals, and the disposal of wastes, have resulted in a wide variety of hazardous chemicals that had either never existed before, or at least not in the levels or chemical form that occur nowadays in our heavily polluted global system. These are typically organic chemicals that are referred to as anthropogenic (`due to humans in nature´) or xenobiotic (`foreign to organisms´) chemicals. In this chapter we aim to clarify the key properties and functionalities of the most common groups of pollutants as a result from human activities, and provide some background on how we can group them and understand their behaviour in the environment.
In the field of environmental toxicology, we are most often concerned about the effects of two distinct types of contaminants: metals and organic chemicals. In some cases other chemicals, such as radioactive elements, may also be important while we could also consider the ecological effects of highly elevated nutrient concentrations (eutrophication) as a form of environmental toxicology.
Metals
Metals and metalloids (elements intermediate between metals and non-metals) comprise the majority of the known elements. They are mined from minerals and used in an enormous variety of applications either in their elemental form or as chemicals with inorganic or organic elements or ions. Many metals occur as cations, but many processes influence the dissolved form of metals. Aluminium for example for example is only present under very acidic conditions as dissolved cation (Al3+), while at neutral pH the metal speciates into for example certain hydroxides (Al(OH3)0). Mercury as free ion is present at (Hg2+) but due to microbial transformation the highly toxic product methylmercury (CH3Hg+, or MeHg+) is formed. Mining and processing of metals together with disposal of metal-containing wastes are the main contributors to metal pollution although sometimes metals are introduced deliberately into the environment as biocides. The widely used pesticide copper sulfate in e.g. grape districts is an example (LINK on comparison to glyphosate here). More information on metals considered to be environmental pollutants is given in Section 2.2.1.
Organic chemicals
Organic chemicals are manufactured to be used in a wide variety of applications. These range from chemicals used as pesticides to industrial intermediates, fossil fuel related hydrocarbons, additives used to treat textiles and polymers, such as flame retardants and plasticisers, and household chemicals such as detergents, pharmaceuticals and cosmetics.
Organic chemicals that we regard as environmental pollutants include a huge variety of different structures and have a wide variety of properties that influence environmental distribution and toxicity. With such a wide variety of chemicals to deal with, it is useful to classify them into groups. Depending on our interest, we can base this classification on different aspects, for example on their chemical structure, their physical and chemical properties, the applications the chemicals are used in, or their effects on biological systems. These aspects are of course closely related to their chemical structure as this is the basis of the properties and effects of chemicals. An overview of different ways of classifying environmental contaminant (sometimes referred to as ecotoxicants) is shown in Tables 1A, 1B, and 1C.
Table 1A.Grouping options of organic contaminants with specific chemical structures
Term
Characteristics
Examples
Hydrocarbons
More CHx units: higher hydrophobicity/lipophilicity, and lower aqueous solubility
hexane
Polycyclic aromatic hydrocarbons
Combustion products. Flat structure
naphthalene, B[a]P
Halogenated hydrocarbons
H substituted by fluor, chlorine, bromide, iodine. Often relatively persistent
PCB, DDT, PBDE
Dioxins and furans
Combustion/industrial products, one or two oxygen atoms between two aromatic rings. Highly toxic.
TCDD, TCDF
Organometallics
Organic chemicals containing metals, used e.g. in anti-fouling paints
tributyltin
Organophosphate pesticides
Phosphate esters, often connecting two lipophilic groups. Act on nervous system
chlorpyrifos
Pyrethroids
Usually synthetic pesticides based on natural pyrethrum extracts
fenvalerate
Neonicotinoids
Synthetic insecticides with aromatic nitrogen, related to the alkaloid nicotine
imidacloprid
… Endless varieties / combinations and ...
...too many characteristics to list
...
Table 1B. Grouping options of organic contaminants with specific properties
Term
Characteristics
Examples
Persistent organic pollutants (POPs)
Bioaccumulative, end up even in remote Arctic systems
PCBs, PFOS
Persistent mobile organic chemicals (PMOCs)
Difficult to remove during drinking water production
PFBA, metformin
Ionogenic organic chemicals (IOCs)
Acids or bases, predominantly ionized under environmental pH
Prozac, MDMA, LAS
Substances of unknown or variable composition, complex reaction products or of biological materials (UVCB)
Multicomponent compositions of often analogue structures with wide ranging properties.
Oil based lubricants
Plastics
Chains of repetitive monomer structures. Wide ranging size/dimensions.
Polyethylene,
silicone, teflon
Nanoparticles (NP)
Mostly manufactured particles with >50% having dimensions ranging 1 - 100 nm.
Titanium dioxide (TiO2), fullerene
Table 1C. Grouping options of organic contaminants with specific usage
Term
Characteristics
Examples
Pesticides
Herbicides
Insecticides
Fungicides
Rodenticides
Biocides
Toxic to pests
Toxic to plants
Toxic to insects
Toxic to fungi
Toxic to rodents
Toxic to many species
DDT
atrazine, glyphosate
Chlorpyrifos, parathion
Phenyl mercury acetate
Hydrogen cyanide
Benzalkonium
Pharmaceuticals
Specifically bioactive chemicals with often (un)known side effects. Many bases.
Produced in large volumes by chemical industry for a wide array of products and processes
phenol
Fuel products
Flammable chemicals
kerosene
Refrigerants and propellants
Small chemicals with specific boiling points
freon-22
Cosmetics/personal care products
Wide varieties of specific ingredients of formulations that render specific properties of a product
sunscreen, parabenes
Detergents and surfactants
Long hydrophobic hydrocarbon tails and polar/ionic headgroups
Sodium lauryl sulfate (SLS), benzalkonium
Food and Feed Additives
To preserve flavor or enhance its taste, appearance, or other qualities
“E-numbers”, acetic acid = E260 in EU, additive 260 in other countries
Chapter 2 mostly discusses groups of chemicals in separate modules according to the specific environmental properties in Table 1B (Section 2.2) and specific applications in Table 1C (Section 2.3), according to which certain regulations apply in most cases. The property classifications can be based on (often interrelated) properties such as solubility (in water), hydrophobicity (tendency to leave the water), surface activity (tendency to accumulate at surfaces of two phases, such as for “surfactants”), polarity, neutral or ionic chemicals and reactivity. Other classifications very important for environmental toxicology are based on environmental behaviour or effects, such as persistency (“P” increasing problems with increased emissions), bioaccumulation potential (“B”, up-concentration in food chains), or type of specific toxic effects (“T”). The influence of specific chemical structures such as in Table 1A is further clarified in the current introductory chapter in order to better understand the basic chemical terminology.
Structures of organic chemicals and functional groups
Hydrocarbons and polycyclic aromatic hydrocarbons
As the name suggest, hydrocarbons contain only carbon and hydrogen atoms and can therefore be considered to be the simplest group of organic molecules. Nevertheless, this group covers a wide variety of aliphatic, cycloaliphatic and aromatic structures (see Figure 1 for some examples) and also a wide range of properties. What this group shares is a low solubility in water with the larger molecules being extremely insoluble and accumulating strongly in organic media such as soil organic matter.
Figure 1.Examples of hydrocarbons.
As a result of the ability of carbon to form strong bonds with itself and other atoms to form structures containing long chains or rings of carbon atoms there is a huge and increasing number (millions) of organic chemicals known. Chemicals containing only carbon and hydrogen are known as hydrocarbons. Aliphatic molecules consist of chains of carbon atoms as either straight or branched chains. Molecules containing multiple carbon-carbon bonds (C=C) are known as unsaturated molecules and can be converted to saturated molecules by addition of hydrogen.
Cyclic alkanes consist of rings or carbon atoms. These may also be unsaturated and a special class of these is known as aromatic hydrocarbons, for example benzene in Figure 1. The specific electronic structure in aromatic molecules such as benzene makes them much more stable than other hydrocarbons. Multiple aromatic rings linked together make perfectly flat molecules, such as pyrene in Figure 1, that can be polarized to some extent because of the shared electron rings. In larger sheets, these polycyclic aromatic molecules also make up the basic graphite structure in pencils, and also typically represent the strongly adsorbing surfaces of black carbon phases such as soot and activated carbon.
The structures of organic chemicals help to determine their properties as behaviour in the environment. At least as important in this regard, however, is the presence of functional groups. These are additional atoms or chemical groups that are present in the molecule that have characteristic chemical effects such as increasing or decreasing solubility in water, giving the chemical acidic or basic properties or other forms of chemical reactivity. The common functional groups are shown in Table 2.
Table 2. Common Functional Groups, where R are carbon backbone or hydrogen units
Halogenated hydrocarbons: first generation pesticides
The first organic chemical recognised as an environmental pollutant was the insecticide DDT (see clip 1 at the start of this chapter). It later became clear that other organochlorine pesticides such as lindane and dieldrin (Table 3) were also widely distributed in the environment. This was also the case for polychlorinated biphenyls (PCBs) and other organochlorinated industrial chemicals. These chemicals all share a number of undesirable properties such as environmental persistence, very low solubility in water and high level of accumulation in biota to potentially toxic levels. Many organochlorines can be viewed as hydrocarbons in which hydrogen atoms have been replaced by chlorine. This makes them even less soluble than the corresponding hydrocarbon due to the large size of chlorine atoms. In addition, chlorination also makes the molecules more chemically stable and therefore contributes to their environmental persistence. Other organochlorines contain additional functional groups, such as the ether bridges in PCDDs and PCDFs (better known as dioxins and dibenzofurans) and ester groups in the 2,4-D and 2,4,5-T herbicides. Many organochlorines were applied very successfully in huge quantities as pesticides for decades before their negative effects such as persistence and accumulation in biota became apparent. It is therefore no coincidence that the initial set of Persistent Organic Pollutants (POPs) identified in the Stockholm Treaty (see below) as chemicals that should be banned were all organochlorines, as shown in Table 3.
As well as chlorine, other halogens such as bromine and fluorine are used in important groups of environmental contaminants. Organobromines are best known as flame retardants and have been applied in large quantities to improve the fire safety of plastics and textiles. They share many of the same undesirable properties of organochlorines and several classes have now been taken out of production. Organofluorines are another important class of halogenated chemicals, and part of the well-known group of ozone depleting CFCs (Section 2.3.6). In particular, per-and polyfluoralkyl substances are widely used as fire-stable surfactants in fire-fighting foams, as grease and water resistant coatings and in the production of fluoropolymers such as Teflon. Organofluorines are much more water soluble and much less bioaccumulative than organochlorines and organobromines but are extremely persistent in the environment.
The recognition of these organochlorines as harmful environmental contaminants eventually resulted in measures to restrict their manufacture and use in the Stockholm Convention on Persistent Organic Pollutants signed in 2001 to eliminate or restrict the production and use of persistent organic pollutants (POPs). This initial list of POPs has been subsequently augmented with other harmful halogenated organic pollutants up to a total of 29 chemicals, which are either to be eliminated, restricted, or required measured to reduce unintentional releases. POPs are further discussed in section 2.2.4.
Table 3. Key persistent organic pollutants, also named POPs – the Dirty Dozen
Additional POPs to eliminate include: chlordecone, lindane (hexachlorocyclohexane), pentachlorobenzene, endosulfan, chlorinated naphthalenes, hexachlorobutadiene, tetrabromodiphenylether, and pentabromodiphenylether decabromodiphenyl ether (BDEs).
Alternatives for the organochlorine pesticides: effective functional groups
Since the signing of the Stockholm Convention, organochlorine pesticides have been replaced in most countries by more modern pesticide types such as the organophosphorus and carbamate insecticides. These compounds are less persistent in the environment, but still could pose elevated risks to environments surrounding the agricultural sites, and increased levels on food produced on these agricultural sites. The very toxic organophosphorus neurotoxicant parathion has been in use since the 1940s, and has the typical two lipophilic side chains on two esters (ethyl units), as well as a polar unit. Parathion has caused hundreds of fatal and non-fatal intoxications worldwide and as a result it is banned or restricted in 23 countries. The relatively comparable organophosphate structure of diazinon has been widely used for general-purpose gardening and indoor pest control since the 1970s, but residential use was banned in the U.S. in 2004. In Californian agriculture however, 35000 kg diazinon was used in 2012. The carbamate based insecticide carbaryl is toxic to target insects, and also non-target insects such as bees, but is detoxified and eliminated rapidly in vertebrates, and not secreted in milk. Although illegal in 7 countries, carbaryl is the third-most-used insecticide in the U.S., approved for more than 100 crops. In 2012, 52000 kg carbaryl was used in California, while this was 3 times more in 2000. Neonicotinoid insecticides, with the typical aromatic ring containing nitrogen, form a third generation of pesticide structures. Imidacloprid is currently the most widely used insecticide worldwide, but as of 2018 banned in the EU, along with two other neonicotinoids clothianidin and thiamethoxam.
Figure 2.Some examples of second and third generation replacements of organochlorine pesticides.
Relatively simple and (very) complex pollutants
As well as the pesticides discussed above, many other chemicals are brought into the environment inadvertently during their manufacture, distribution and use and the range of chemicals recognised as problematic environmental contaminants has expanded enormously. These include fossil fuel-related hydrocarbons, surfactants, pigments, biocides and chemicals used as pharmaceuticals and personal care products (PPCPs). Figure 3 gives an illustrative overview of the major routes by which PPCPs, but also many other anthropogenic contaminants other than pesticides, are released into the environment. Particularly wastewater treatment systems form the main entry point for many industrial and household products.
Figure 3.Emission pathways to soil and water for pharmaceuticals and personal care products. Adapted from Boxall et al. (2012) by Evelin Karsten-Meessen.
The wide variety of contaminant structures does not mean that most chemicals have become increasingly more complex. For risk assessment, molecular properties such as water solubility, volatility and lipophilicity are often estimated based on quantitative structure-property relationships (Section 3.4.3). With increasingly complex structures, such property-estimations based on the molecular structure become more uncertain.
The antibiotic erythromycin for example (Figure 4), is a very complex chemical structure (C37H67NO13) that has 13 functional units along with a 14 member ring. In addition, the tertiary nitrogen group is an amine base group that can give the molecule a positive charge upon protonation, depending on the environmental pH. Erythromycin is on the World Health Organization's List of Essential Medicines (the most effective and safe medicines needed in a health system), and therefore widely used. Continuous emissions in waste streams pose a potential threat to many ecosystems, but many environmentally and toxicologically relevant properties are scarcely studied, and poorly estimated.
Figure 4.Relatively simple or (very) complex chemicals? Glyphosate speciation profile (chemicalize.org) for the 4 dominant glyphosate different species. Glyphosate in soil (pH4-8) predominantly occurs with 3 charged groups (net charge -1), partly with 4 charged groups (-2). The soap SLS is always negatively charged, GHB predominantly negative (pKa 4.7), amphetamine predominantly positive (pKa 9.9), erythromycin predominantly positive (pKa 8.9)
There are also many contaminants or toxicants with a seemingly simple structure. Many surfactants are simple linear long chain hydrocarbons with a polar or charged headgroup (Figure 4). The illicit drug amphetamine has only a benzene ring and an amine unit, the illicit drug GHB only an alcohol and a carboxylic acid, the herbicide glyphosate only 16 atoms. Still, these 4 chemical examples also have acidic or basic units that often result in predominantly charged organic molecules, which also strongly influences their environmental and toxicological behaviour (see sections on PMOCs and Ionogenic Organic Compounds). In case of glyphosate, the chemical has 4 differently charged forms depending on the pH of the environment. At common pH of 7-9, glyphosate has all charged groups predominantly ionized, making it very difficult to derive calculations on environmental properties.
References
Boxall, A. B. A., Rudd, M. A., Brooks, B. W., Caldwell, D. J., Choi, K., Hickmann, S., Innes, E., Ostapyk, K., Staveley, J. P., Verslycke, T., Ankley, G. T., Beazley, K. F., Belanger, S. E., Berninger, J. P., Carriquiriborde, P., Coors, A., DeLeo, P. C., Dyer, S. D., Ericson, J. F., Gagné, F., Giesy, J. P., Gouin, T., Hallstrom, L., Karlsson, M. V., Larsson, D. G. J., Lazorchak, J. M., Mastrocco, F., McLaughlin, A., McMaster, M. E., Meyerhoff, R. D., Moore, R., Parrott, J. L., Snape, J. R., Murray-Smith, R., Servos, M. R., Sibley, P. K., Oliver Straub, J., Szabo, N. D., Topp, E., Tetreault, G. R., Trudeau, V. L., Van der Kraak, G. (2012). Pharmaceuticals and personal care products in the environment: what are the big questions? Environmental Health Perspectives 120, 1221-1229.
2.2. Pollutants with specific properties
2.2.1. Metals and metalloids
Author: Kees van Gestel
Reviewers: John Parsons, Jose Alvarez Rogel
Learning objectives:
You should be able to:
describe the difference between metals and metalloids
describe a classification using different binding affinities of metals to macromolecules and infer on its importance for their toxicity and/or bioaccumulation
mention important sources of metal pollution
Keywords: Heavy metals, Metalloids, Rare earth elements, Essential elements
Introduction
The majority of the elements in the periodic table consists of metals: Figure 1.
Figure 1.Periodic table of elements, with the most important elements for Environmental Toxicology shown. The shaded elements are metals, the partially shaded elements are metalloids. Bold lettered metals area heavy metals (specific density > 5 g/cm3). Elements shown within bold lines (and in italics) are essential elements. The Lanthanides and Actinides together are the rare earth elements (REEs). (Source: Steven Droge).
The distinction between metals and heavy metals (density relative to water < or >5 g cm-3) is not very meaningful for such a heterogeneous group of elements with rather different biological and chemical properties. The rare earth elements (REEs), lanthanides and actinides, have, for example, a high density or specific weight but are usually not considered heavy metals because of their rather different chemical behaviour. Metalloids have both metallic and non-metallic properties or are nonmetallic elements that can combine with a metal to produce an alloy. Figure 1 shows the periodic table of elements, indicating the groups of (heavy) metals, metalloids and rare earth elements.
Also indicated in Figure 1 are the elements that are known to be essential to life and include besides C, H, O and N, the major essential elements Ca, P, K, Mg, Na, Cl and S, the trace elements Fe, I, Cu, Mn, Zn, Co, Mo, Se, Cr, Ni, V, Si, As and B (the latter only for plants) and some elements that may support physiological functions at ultra-trace levels (Li, Al, F and Sn) (Walker et al., 2012).
Chemical and physical properties
Except for mercury, most pure metals are solid at room temperature. In general, metals are good electrical and thermal conductors having high luster and malleability. Upon heating, metals readily emit electrons. These descriptors of metals, however, are not very helpful when having to deal with elements that do not exist prominently in the pure elemental state, but rather are present as metal compounds, complexes, and ions at fairly low environmental concentrations.
More useful are characteristics that influence metal transport between environmental compartments and their interaction with abiotic and biotic components of the environment. The speciation, the chemical form in which an element occurs (e.g., oxidized, free ion or complexed to inorganic or organic molecules), determines its transport and interaction in the environment (see Section on Metal Speciation). Chemical bonding is determined by outer orbital electron behavior, with metals tending to lose electrons when reacting with nonmetals. In many normal biological reactions, metals are cofactors within coenzymes (e.g. in vitamins) and can act as electron acceptors and donors during oxidation and reduction reactions (Newman, 2015).
Nieboer and Richardson (1980) proposed a classification, based on the equilibrium constant for the formation of metal complexes. They distinguished:
Class A-metals: acting as hard Lewis acids (electron acceptors) with high affinity for oxygen-containing groups in macromolecules, such as carboxyl and alcohol groups. Al, Ba, Be, Ca, K, Li, Mg, Na and Sr belong to this group;
Class B-metals: acting as soft Lewis acids with high affinity for nitrogen- and sulphur-containing groups in macromolecules, such as amino and sulphydryl groups. This group includes Ag, Au, Bi, Hg, Pd, Pt and Tl.
In addition, an intermediate or borderline group is defined, in which the type A or B characteristics are less pronounced. As, Cd, Co, Cr, Cu, Fe, Mn, Ni, Pb, Sb, Sn, Ti, V, and Zn belong to this group.
This classification of metals is highly relevant for the transport across cell membranes, the intercellular storage in granules and the induction of metal-binding proteins as well as for their behaviour in the environment in general.
Occurrence
(Heavy) metals and rare earth elements are diffusely distributed over the Earth, but at some places certain elemental combinations are highly concentrated (in metal ores). Despite this diffuse distribution, differences in background metal concentrations in soils can be large, depending on the type and origin of rock or sediment (Table 1).
Table 1. Background concentrations (mg/kg dry weight) of (heavy) metals and metalloids in crust material and median and maximum concentrations in different top soils across the world. Derived from Kabata-Pendias and Mukherjee (2007) and Alloway (2013).
In general, volcanic rock (e.g. basalt) contains high and sedimented rock (e.g. limestone) low metal levels. But there is no relation between metal concentrations in the Earth's crust and the elemental requirements of organisms.
Emissions of metals
Upon weathering of stone formations and ores, elements are released and enter local, regional and global biogeochemical cycles. Depending on their water solubility and on soil properties and vegetation, metals may be transported through the environment and deposited or precipitated at places close to or far away from their source.
Volcanoes take account of the largest natural input of metals to the environment but the concentrations of these metals in the soil are rarely elevated to toxic levels due to the massive dilution which takes place in the atmosphere. Permanently active volcanoes may be an important local source of (metal) pollution.
A special case is arsenic, that may occur as a natural element of soils. At some places, As levels are fairly high, particularly in ground water. High As-groundwater areas are found in Argentina, Chile, Mexico, China and Hungary, and also in Bangladesh, India (West Bengal), Cambodia, Laos and Vietnam. In the latter countries, especially in the Bengal Basin, millions of wells have been dug to provide safe drinking water. Irrigation pumping leads to an inflow of oxygen and organic carbon, which causes a mobilisation of arsenic normally bound to ferric oxyhydroxides in these soils. As a result in many wells dissolved As concentrations are exceeding the World Health Organisation (WHO) guideline value of 10 µg/L for drinking water.
Important anthropogenic sources of metals in the environment include:
Metal mining, which may also lead to an enormous physical disturbance of the environment (destruction of ecosystems).
Metal smelting.
Use of metals in domestic and industrial products, and the discharge of domestic waste and sewage.
Metal-containing pesticides, e.g. 'Bordeaux Mixture (copper sulphate with lime (Ca(OH)2), used as a fungicide in viniculture, hop-growing and fruit-culture, and metal-containing fungicides, such as organo-tin compounds.
The use of metals and especially of REEs in microelectronics.
Energy-producing industries burning coal and oil, and producing metal-containing fly ash.
Transport of energy and traffic making use of electricity, giving rise to corrosion of electric wires and pylons.
Non-metal industries, e.g. leather (chromium) and cement production (thallium).
, using Tetra Ethyl Lead (TEL) as anti-knocking agent in petrol (nowadays banned in most countries) and the use of catalysts in cars (platinum, palladium).
Anthropogenic releases of many metals, such as Pb, Zn, Cd and Cu, are estimated to be between one and three orders of magnitude higher than natural fluxes (Depledge et al. 1998). An estimated amount of up to 50,000 tonnes of mercury are released naturally per year as a result of degassing from the Earth's crust, but human activities account for even larger emissions (Walker et al. 2012).
References
Alloway, B.J. (2013). Heavy Metals in Soils. Trace Metals and Metalloids in Soils and their Bioavailability. Third Edition. Environmental Pollution, Volume 22, Springer, Dordrecht.
Depledge, M.H., Weeks, J.M., Bjerregaard, P. (1998). Heavy metals. In: Calow, P. (Ed.). Handbook of Ecotoxicology. Blackwell Science, Oxford, pp. 543-569.
Kabata-Pendias, A., Mukherjee, A.B. (2007). Trace Elements from Soil to Human. Springer Verlag, Berlin.
Newman, M.C. (2015). Fundamentals of Ecotoxicology. The Science of Pollution. Fourth Edition. CRC Press, Taylor & Francis Group. Boca Raton.
Nieboer, E., Richardson, D.H.S. (1990). The replacement of the nodescript term 'Heavy metals' by a biologically and chemically significant classification of metal ions. Environmental Pollution (Ser. B) 1, 3-26.
Walker, C.H., Hopkin, S.P., Sibly, R.M., Peakall, D.B. (2012). Principles of Ecotoxicology, Fourth Edition. CRC Press Taylor & Francis Group, London.
2.2.2. Radioactive compounds
Authors: Nathalie Vanhoudt, Nele Horemans
Reviewer: Robin de Kruijff
Learning objectives:
You should be able to:
describe the process of radioactive decay
describe the different types of radiation and their interaction with matter
explain the difference between natural and artificial radionuclides and give examples
Naturally occurring radionuclides are omnipresent in the environment and exposure to radiation is unequivocally related to life on Earth. Every day we are exposed to cosmic radiation, radon exhalation from the soil and radioactive potassium naturally present in our bodies. Moreover, radionuclides and ionising radiation are successfully applied in many domains such as nuclear medicine, research applications, energy production, food preservation and other industrial activities. To be able to positively apply radionuclides or ionising radiation and to evaluate the impact on man and environment in case of a contamination scenario, it is important to understand the process of radioactive decay, the different types of radiation and radionuclides and how ionising radiation interacts with matter.
Radioactive decay
Radioactivity is the phenomenon of spontaneous disintegration or decay of unstable atomic nuclei to form energetically more stable ones (Krane, 1988). Within this process, particles (e.g. protons, neutrons) and/or radiation (photons) can be emitted. This radioactive decay is irreversible and after one or more transformations, a stable, non-radioactive atom is formed.
Radioactive decay is considered a stochastic phenomenon as it is impossible to predict when any given atom will disintegrate. However, the probability per unit time that an unstable nucleus can decay is described by the disintegration or decay constant λ [s-1]. The fact that this probability is constant, forms the basic assumption of the statistical theory of radioactive decay.
Radioactive decay follows an exponential function (Eq. 1, Figure 1) with N0 the number of nuclei at time 0, N(t) the remaining nuclei at time t and λ the decay constant [s-1].
\(N(t)=N_0 e^{-λt}\) Eq. 1
The decay constant λ is specific for every radionuclide and the half-life t1/2 of a radionuclide can be derived from this constant (Eq. 2).
\(t_{1/2} = {ln 2 \over λ}\) Eq. 2
The specific half-life of a radionuclide gives the time that is necessary for half of the nuclei to decay. Half-lives can vary between fractions of seconds to many billions of years depending on the radionuclide of interest. For example 238U and 232Th are two primordial radionuclides with half-lives of 4.468 × 109 y and 1.405 × 1010 y, respectively. 137Cs on the other hand is an important radionuclide released during the Chernobyl and Fukushima nuclear power plant accidents and has a half-life of 30.17 y. While other shorter-lived radionuclides released during these accidents (e.g. 131I with a half-life of 8 days) have already decayed, 137Cs has the most substantial long-term impact on terrestrial ecosystems and human health owing to its relatively long half-life and high release rate (Onda et al., 2020).
Figure 1. Illustration of exponential decay and half-life t1/2.
The activity A of a radioactive material is defined by the rate at which decay occurs in the sample and is determined by the amount of radioactive nuclei present at time t and the decay constant λ (Eq. 3). As such, the activity of A sample is a continuously decreasing value following the same exponential curve as presented in Fig. 1.
\(A(t)=λN(t)\) Eq. 3
The SI-unit to express activity is Becquerel [Bq], equal to one disintegration per second. In a sample with an activity of 100 Bq, it is expected that 100 radioactive disintegrations will occur every second. Also the older non-SI unit Curie (Ci) is still often used to express activity, with 1 Ci being equal to 3.7 × 1010 Bq.
One way by which an unstable nucleus will strive towards a more stable state is by emitting particles and as such creating a new nucleus. During this process, α-particles, protons, neutrons, β--particles and β+-particles can be emitted by the nucleus.
For example, during alpha decay, an α-particle, which is a stable configuration of two protons and two neutrons (4He nucleus), is emitted, resulting in a new nucleus with an atomic number Z that is two units lower (2 protons) and a mass number A that is 4 units lower (2 protons + 2 neutrons) (Figure 2).
Figure 2. Alpha decay.
During beta decay, the nucleus can correct an imbalance between neutrons and protons by transforming one of its nucleons (i.e. converting a neutron into a proton or vice versa). This process can occur in different ways that all involve an extra charged particle (beta particle or electron) to conserve electric charge (Krane, 1988). During β--decay, a neutron is converted into a proton with emission of a highly energetic negatively charged electron (β--particle) and an antineutrino (Figure 3). During β+-decay, conversion of a proton into a neutron is accompanied by emission of a positively charged electron (positron or β+-particle) and a neutrino. In addition, the nucleus can also correct a proton excess by capturing an inner atomic electron to convert the proton into a neutron. This process is called electron capture.
Figure 3. Beta minus decay.
Although fission is usually considered as a process that is artificially induced (e.g. nuclear reactor), some heavy nuclei with an excess of neutrons naturally decay through fission, resulting in two lighter nuclei and a few neutrons. These new nuclei usually also further decay.
A second way by which a nucleus in its excited state will transform into a more stable state, is by emitting energy in the form of highly energetic electromagnetic radiation called photons. During this process, the original nucleus is maintained. Gamma decay is often a secondary process after alpha or beta decay as the nuclei often contain an excess amount of energy after transformation. As the energies of the emitted gamma rays are unique to each radionuclide, gamma ray energy spectra can be used to identify radionuclides in a sample.
Today, more than 4000 radionuclides are known and information regarding these radionuclides is compiled in a nuclide chart, which is a two dimensional representation of the nuclear and radioactive properties of all known atoms (Figure 4) (Sóti et al., 2019). In contrast to the periodic table, the nuclide chart arranges nuclides according to their number of neutrons (X-axis) and protons (Y-axis). This chart includes information on half-lives, mass numbers, decade modes, energies of emitted radiation, etc. Different colours are used to represent stable nuclei and specific modes of radioactive decay (e.g. alpha decay, beta decay, electron capture). Sóti et al. (2019) can be consulted for more information regarding the content and use of the nuclide chart and an interactive nuclide chart has been made available by the International Atomic Energy Agency (IAEA).
Naturally occurring radionuclides and artificial radionuclides
Naturally occurring radionuclides such as 238U, 232Th, 226Ra and 40K are omnipresent in the environment and high concentrations can often be found in certain geological materials such as igneous rocks and ores. For example, activity concentrations between 7 and 60 Bq kg-1 of 238U and between 70 and 1500 Bq kg-1 of 40K can be found in the most common rock types (IAEA, 2003). One group of naturally occurring radionuclides are the primordial radionuclides that were created before the formation of planet Earth and have long half-lives of billions of years. While some of these primordial radionuclides exist alone (e.g. 40K), others are the head of nuclear decay chains (e.g. 238U, 232Th and 235U). Through subsequent alpha and beta decay, these radionuclides decay until a stable Pb isotope is formed. Radionuclides such as 238U, 232Th, 226Ra, 210Pb, 210Po, with their own specific chemical and radiological properties, are part of these radioactive decay chains. Similar as for other elements, the chemical form in which these radionuclides occur will determine their behaviour and fate in the environment and finally their possible risk to humans and other biota.
The three radioactive decay chains and the primordial radionuclide 40K contribute most to the external background radiation humans are exposed to. Within the 238U and 232Th radioactive decay chains, two isotopes of the noble gas Rn (222Rn and 220Rn, respectively) are formed. In contrast to the other decay products, this noble gas has the potential to migrate through the pores of rocks towards the soil surface. Through this process, radioactive Rn can be released into the atmosphere resulting in an average activity concentration of 1-10 Bq m-3 in air, although this value is highly dependent on the soil type and composition. Although 222,220Rn itself is inert it can decay to other alpha and beta emitters that can attach to tissues. Especially when inhaled, the decay products of 222,220Rn can cause internal lung irradiation.
In addition, several industries (e.g. metal mining and milling, the phosphate industry, oil and gas industry) are involved in the exploitation of natural resources that contain naturally occurring radionuclides. These activities will result in enhanced concentrations of radionuclides in products, by-products and residues that can lead to elevated (or more bioavailable) radionuclide levels in the environment posing a risk to human and ecosystem health (IAEA, 2003).
Besides the primordial radionuclides and radionuclides that are part of the 238U, 232Th or 235U radioactive decay chains, some radionuclides are continuously formed in the atmosphere through interaction with cosmic radiation. For example, 14C is continuously produced in the atmosphere through interaction of thermal neutrons with nitrogen (14N(n,p)14C).
Artificial radionuclides are those radionuclides that are artificially generated, for example in nuclear power plants, particle accelerators and radionuclide generators. These radionuclides can be generated for different purposes such as energy production, medical applications and research activities.
In the last century, nuclear weapon production and testing, improper waste management, nuclear energy production and related accidents have contributed to the spread of a large array of anthropogenic radionuclides in the environment, including 3H, 14C, 90Sr, 99Tc, 129I, 137Cs, 237Np, 241Am and several U and Pu isotopes (Hu, 2010). Although a wide range of radionuclides were released during the Chernobyl and Fukushima nuclear power plant accidents, most of them had half-lives of hours, days and weeks resulting in a rapid decline of radionuclide activity concentrations (IAEA, 2006, 2020). After the initial release period, 137Cs remained the most important radionuclide causing enhanced long-term exposure risk for humans and biota (IAEA, 2006, 2020). Nonetheless, compared to nuclear weapon production and testing, nuclear accidents contribute only for a small fraction to the environmental contamination (Hu, 2010). Recent maps on the 137Cs atmospheric fallout from global nuclear weapon testing and the Chernobyl accident in European topsoils are presented by Meusburger et al. (2020).
Interaction of ionising radiation with matter
Ionising radiation has the potential to react with atoms and molecules in matter and cause directly or indirectly ionisations, excitations and radicals, which will result in damage to organisms. Although ionising radiation can originate from radioactive decay, it can also be artificially generated (e.g. X-rays) or come from cosmic radiation.
Directly ionising radiation consists of charged particles such as alpha or beta particles with sufficient kinetic energy to cause ionisations. When colliding with electrons, these particles can transfer part of their energy resulting in ionisations. Alpha particles have usually a high energy, typically around 5 MeV. Due to their relatively high mass, high kinetic energy and their charge, they have a high ionising potential. When interacting with matter, an alpha particle follows, due to its high mass, a relatively straight and short path along which ionisations and excitations occur (Figure 5). During each interaction, a small amount of the particle’s energy is transferred until it is finally stopped. This will result in a lot of damage in a small area, hence its high ionising potential. As their penetration depth is low, the alpha particle can be stopped by a few centimetres of air or a sheet of paper (Figure 5). This means that alpha particles cannot penetrate the skin resulting in low hazard in case of external irradiation. On the other hand, when present inside the body (in case of internal contamination), much more damage can be induced due to its high ionising capacity and the lack of a shielding barrier. Their property to deposit all their energy in a very small area makes alpha emitters perfectly suited for the local treatment of tumour cells. A targeting biomolecule to which an alpha emitter (or a radionuclide that decays into an alpha emitter) is chemically bound can be injected intravenously to spread through the body and accumulate in specific body tissues or cells where it locally irradiates tumour metastases.
Beta particles are high speed electrons or positrons emitted during radioactive decay. Due to their low mass, usually high kinetic energy and their charge, they have a lower ionising potential compared to alpha radiation but a higher penetration depth. In contrast to alpha particles, beta particles do not follow a linear path when interacting with matter. When colliding with other electrons, beta particles can change direction, resulting in a very irregular interaction pattern (Figure 5). In air, beta particles have a penetration potential from several decimetres up to a few meters while this is reduced to centimetres when interacting with solids or liquids. Care has to be taken when selecting the best shielding material as beta particles can also generate Bremsstrahlung, which is electromagnetic radiation produced when the beta particle is deflected in the electric field of an atomic nucleus (Figure 5). Materials with low atomic number such as Plexiglas or aluminium are preferred to minimize the additional risk of Bremsstrahlung production (Figure 5). As beta particles can penetrate the human tissue up to a few millimetres, it forms both an external and internal risk.
In the case of indirect ionising radiation such as gamma radiation, charged particles are first created through energy transfer from the radiation field to matter which will then cause ionisations. Not all types of electromagnetic radiation are considered ionising radiation. Only radiation with a short wavelength will have sufficient energy to induce ionisations such as gamma radiation, X-rays and high energy UV radiation. However, the interaction with matter is fundamentally different between charged and uncharged particles such as gamma radiation. While charged particles interact with many particles at the same time, resulting in a lot of ionisations, uncharged particles will mainly not interact with particles along their pathway through matter but there is a likelihood that they will interact. When they interact, ionisations are induced indirectly as energy is first transferred to release charged particles (such as electrons) that will in turn cause ionisations (Figure 5). As such, due to its lack of charge and minimal mass, gamma radiation has a high penetration potential, forming an important internal and external risk. Lead is a commonly used shielding material for gamma radiation (Figure 5). Nonetheless, the high penetration potential and the difference in interaction with tissues of different density, forms the basis to use X-rays in internal imaging techniques for medical and industrial purposes.
Figure 5. Illustration of the interaction of alpha, beta and gamma radiation with matter.
References
Hu, Q.-H., Weng, J.-Q., Wang, J.-S. (2010). Sources of anthropogenic radionuclides in the environment: a review. Journal of Environmental Radioactivity 101, 426-437. https://doi.org/10.1016/j.jenvrad.2008.08.004.
IAEA (2003). TRS 419 Extent of environmental contamination by naturally occurring radioactive material (NORM) and technological options for mitigation. International Atomic Energy Agency. Vienna, Austria.
IAEA (2006). Environmental consequences of the Chernobyl accident and their remediation: Twenty years of experience. International Atomic Energy Agency. Vienna, Austria.
IAEA (2020). TECDOC 1927 Environmental transfer of radionuclides in Japan following the accident at the Fukushima Daiichi Nuclear Power Plant. International Atomic Energy Agency. Vienna, Austria.
Krane, K. (1988) Introductory Nuclear Physics. John Wiley & Sons, Inc.
Meusburger, K., Evrard, O., Alewell, C., Borrelli, P., Cinelli, G., Ketterer, M., Mabit, L., Panagos, P., van Oost, K., Ballabio, C. (2020). Plutonium aided reconstruction of caesium atmospheric fallout in European topsoils. Scientific Reports 10:11858. https://doi.org/10.1038/s41598-020-68736-2.
Onda, Y., Taniguchi, K., Yoshimura, K., Kato, H., Takahashi, J., Wakiyama, Y., Coppin, F., Smith, H. (2020). Radionuclides from the Fukushima Daiichi Nuclear Power Plant in terrestrial systems. Nature Reviews Earth & Environment 1, 644-660. https://doi.org/10.1038/s43017-020-0099-x.
Sóti, Z., Magill, J., Dreher, R. (2019). Karlsruhe Nuclide Chart – New 10th edition 2018. EPJ Nuclear Sciences and Technologies 5, 6. https://doi.org/10.1051/epjn/2019004.
2.2.3. Industrial Chemicals
Authors: Steven Droge
Reviewer: Michael McLachlan
Leaning objectives:
You should be able to
discuss a history perspective on key chemical legislations around the world
look up registration dossiers yourself to obtain relevant ecotoxicological information
realize that complete dossiers are most urgent for high production tonnage substances and the most hazardous substances
understand why for some groups of chemicals already specific regulations were in place apart from common industrial substances.
Keywords: Chemical industry, tonnage, hazardous chemicals, REACH, regulation
Introduction
The chemical industry produces a wide variety of chemicals that find use in industrial process and as ingredients in day-to-day products for consumers. Instead of chemicals, ‘substances’ may be a more carefully worded description as it also includes complex mixtures, polymers and nanoparticles. Many substances are produced by globally distributed companies in very high volumes, ranging for example from 100 - 10,000 tonnes (1 tonne = 1000 kg) per year. Worldwide, governments have tried to control and assess chemical safety, as nicely summarized on the ChemHAT website. Australia for example, has the Industrial Chemicals (Notification and Assessment) Act 1989 (2013 version). Just like elsewhere in the world, in the European Union (EU) a variety of regulatory institutes at all levels of government used to perform safety assessments regarding the use of substances in products, and how these are emitted into waste streams. This changed dramatically in 2007.
On June 1st, 2007 (Figure 1), a new EU regulation went into force called REACH (official legislation documents C 1907/2006; about REACH; EU info on REACH). This law reversed the role of governments in chemical safety assessment, because it placed the burden of proof on companies that manufacture a chemical, import a chemical into the EU, or apply chemicals in their products. Within REACH companies must identify and manage the risks linked to the chemicals they manufacture and market in the EU. REACH stands for Registration, Evaluation, Authorisation and Restriction of Chemicals. China soon followed with the analogous “China REACH” in 2010, and then came South Korea in 2015 with “K-REACH”. The main focus in this module is on EU-REACH as the leading and well documented example. Other legislation regulating industrial chemicals can often be easily found online, e.g. via the ChemHAT link above.
Figure 1. Scheme for the registration phase of the REACH regulation for existing industrial chemicals of different tonnage bands and hazardous potential, as well as newly developed chemicals (“Non phase-in”) for the EU market. CMRs = chemicals that are proven carcinogenic, mutagenic or toxic to reproduction. R50/R53 labels indicate “Very toxic to aquatic organisms”/ “May cause long-term adverse effects in the aquatic environment”. Source: http://www.cirs-reach.com/REACH/REACH_Registration_Deadlines.html (with permission).
In REACH, each chemical is registered only once. Accordingly, companies must work together to prepare one dossier that demonstrates to the European Chemical Agency (ECHA) how chemicals can be safely used, and they must communicate the risk management measures to the users. ECHA, or any Member State, authorizes the dossiers, and can start a “restriction procedure” when they are concerned that a certain substance poses an unacceptable risk to human health or the environment. If the risks cannot be managed, authorities can restrict the use of substances in different ways. In the long run, the most hazardous substances should be substituted with less dangerous ones.
So which chemicals have been registered in the past decade (2008-2018) in REACH?
In principle, REACH applies to all chemical ‘substances’ in the EU zone. This includes metals, such as “iron” and “chromium”, organic chemicals such as “methanol” and “fatty acids” and “ethyl-4-(8-chloro-5,6-dihydro-11H-benzo[5,6]cyclohepta[1,2-b]pyridin-11-ylidene)piperidine-1-carboxylate (see Box 1)”, and (nano)particles like “zink oxide” and “silicon dioxide”, and polymers. Discover for example the registration dossier link in Box 1.
Box 1.Examples from the REACH dossiers
The REACH registration data base can be searched via LINK. Accept the disclaimer, and you are ready to search for chemicals based on name, CAS number, substance data, or use and exposure data.
Search for example for the name “ethyl 4-(8-chloro-5,6-dihydro-11H-benzo[5,6]cyclohepta[1,2-b]pyridin-11-ylidene)piperidine-1-carboxylate” and you find the link to the dossier of this substance with CAS 79794-75-5 as compiled by the registrant. This complex chemical name is better known as the antihistamine drug Loratadine, but this name does not show up in the dossier search!
Click on the name to get basic information on the compound. The hazard classification reads: “Warning! According to the classification provided by companies to ECHA in REACH registrations this substance is very toxic to aquatic life, is very toxic to aquatic life with long lasting effects, is suspected of causing cancer, causes serious eye irritation, is suspected of causing genetic defects, causes skin irritation, may cause an allergic skin reaction and may cause respiratory irritation.” This compound is “PBT” labeled based on limited available data (classifying as a combination of Persistent / Bioaccumulative / Toxic). However, the section [About this substance] reads: “for industrial use resulting in the manufacture of another substance (use of intermediates).” As an intermediate in a restricted process, many parts of the dossier did not have to be completed for REACH. As a medicinal product Loratadine is strictly regulated elsewhere. Scroll down to the REACH link for the registration dossier (.../21649) to find out more the different entries for this chemical.
If we do a search for [“ Bisphenol ”], we get a long list of optional chemicals, for example Bisphenol A (CAS 80-05-7) but also for example Bisphenol S if you scroll down further (CAS 80-09-1). If we look at the dossier of the first Bisphenol A entry, with tonnage “100 000 - 1 000 000 tonnes per annum”, you can find a long list of REACH information packages besides the dossier, as this chemical is hotly debated. The dossier for Bisphenol A was evaluated in 2013, and also this is available (look for the pdf in theDossier evaluation status). In this compliance check, the registrant is requested to submit additional rat and mouse toxicity data, along with statements of reasons. There is for example also a link to the [Restriction list (annex XVII)], which leads to a pdf called 66.pdf, which states an adopted restriction for this chemical within the REACH framework and the previous legislation, Directive 76/769/EEC: “Shall not be placed on the market in thermal paper in a concentration equal to or greater than 0,02 % by weight after 2 January 2020”.
Find your own chemical of interest to discover more on the transparancy of the chemical information on which risk assessment is based.
However, some groups of chemicals are (partly) exempt from REACH because they are covered by other legislation in the EU:
Active substances used in plant protection products (Section 2.3.1) and biocidal products (Section 2.3.2) are considered as already having been registered and assessed by institutes separate from ECHA. Biocides such as disinfectants and pest control products are per definition hazardous chemicals, but they are also very useful in many ways. The very strict and elaborate biocide laws aim to verify that the potential risk of harm associated with the intended emission scenarios is in balance with expected benefits.
Food and feedstuff additives (Section 2.3.9) have different legislation and authorisation laws to demonstrate (following a scientific evaluation) that the additive has no harmful effects on human and animal health or on the environment ( developed since 1988,schematic graph, Regulation (EC) No 1331/2008)
Medicinal products (Sections 2.3.3 and 2.3.4) have different legislation and authorisation laws to guarantee high standards of quality and safety of medicinal products, while promoting the good functioning of the internal market with measures that encourage innovation and competiveness (starting with Directive 65/65 in 1965,an overview since, a pdf of the 2001 EU legislation2001/83/EC)
“Waste” is not part of the REACH domain, but a product recovered from waste is not.
A detailed overview of European chemical safety guidelines related to chemicals with different application types is presented in Figure 2.
Figure 2.Scheme of societal sectors, their chemical uses, the dedicated policy frameworks for registration and authorization, and pathways to the aqueous environment direct or via industrial- or household effluent treatment plants (circle symbols). Redrawn from Van Wezel et al. (2017) by Evelin Karsten-Meessen.
Following pre-registration of the 145,297 chemicals most likely to require regulation, REACH came into force in 2008 in a stepwise process with different deadlines for different groups of chemicals. The first dossiers were to be completed by 2010 for the highest produced volume chemicals (>1000 tonnes/y) and the most hazardous chemicals (CMRs >1 tonne/y, and chemicals with known very high aquatic toxicity >100 tonnes/y). These groups potentially pose the greatest risk because of either their high emissions or their inherent toxicity. In 2013, registration dossiers for chemicals with a lower tonnage (100-1000 tonnes/y) were to be completed. By May 31 2018, all chemicals with a quantity of 1-100 tonnes/y chemicals on the EU market should have been registered. New chemicals will all be subject to the REACH procedures.
In 2018, 21.787 substances had been registered under REACH. A total of 14.262 companies were involved. In comparison, 15.500 substances were registered in 2016 (i.e., 6287 chemicals were added in the two following years). In 2018, 48% of all substance registrations had been done in Germany. For 24% of the registered substances a dossier was already available prior to REACH, 70% are “old chemicals” for which no registration had been done before REACH was initiated, and only 6% are newly developed substances that needed to be registered before manufacture or import could start.
There are multiple benefits of REACH regulation of industrial chemicals. Most data on chemicals entered in the registration process is publically available, creating transparency and improving customer awareness. If registered chemicals are classified as Substance of Very High Concern (SVHC) based on the chemical information in these dossiers and after agreement from research panels, alternatives that passed the same regulation can be suggested instead.
The necessity to add data on potential toxicity for so many chemicals has been combined with a strong focus on, and further development of, animal friendly testing methods. Read-across from related chemicals, weight of evidence approaches, and calculations based on chemical structures (QSAR) allow much experimental testing to be circumvented. In vitro studies are also used, but a 2017 REACH document (REACH alternatives to animal testing 2017, which followed up 2011 and 2014 reports) reports that 5.795 in vitro studies were used overall to determine endpoints for REACH, compared to 9,287 in vivo studies (ratio of 0.6). Clearly, many new animal tests have been performed under REACH to complete the dossiers on industrial chemicals. Prenatal developmental and repeated dose toxicity testing as well as extended one generation reproductive toxicity studies remain difficult to circumvent without animal use. However, the safe use of industrial chemicals must be ensured and demonstrated.
References:
Van Wezel, A.P., Ter Laak, T.L., Fischer, A., Bäuerlein, P.S., Munthee, J., Posthuma, L. (2017). Mitigation options for chemicals of emerging concern in surface waters; operationalising solutions-focused risk assessment . Environmental Science: Water Research 3, 403–414.
2.2.4. POPs
(draft)
Authors: Jacob de Boer
Reviewer:
Leaning objectives:
You should be able to
understand how POPs are defined
recognize chemical structures of POPs
gain knowledge on the purpose of the Stockholm Convention on POPs
Keywords: Persistence, bioaccumulation, long range transport, toxicity, analysis
Introduction
Chemicals are generally produced because they have a useful purpose. These purposes can vary widely, such as to protect crops by killing harmful insects or fungi, to protect materials against catching fire, to act as a medicine, to enable a proper packing of food materials, etc. Unfortunately, the properties which make a chemical attractive to use, often have a downside when it comes to environmental behavior and/or human health. A number of synthetic chemicals have properties that make them persistent organic pollutants (POPs). POPs are xenobiotic (foreign to the biosphere) chemicals that are persistent, bioaccumulative and toxic (‘PBT’) in low doses. In addition, they are transported over long distances. Criteria for these properties, which are used to define a chemical as a POP, were set by the United Nations (UN) Stockholm Convention, which was adopted in 2001 and entered into force in 2004 (Fiedler et al., 2019). These criteria are summarized in Table 1 (http://chm.pops.int). The objective of the Stockholm Convention is defined in article 1: “Mindful of the precautionary approach, to protect human health and the environment from the harmful impacts of persistent organic pollutants”. Initially, 12 chemicals (aldrin, chlordane, dieldrin, DDT, endrin, heptachlor, hexachlorobenzene (HCB), mirex, polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs) and toxaphene) were listed as POPs. Gradually the list was extended with new POPs, which appeared to fulfil the criteria. For some of the new chemicals exceptions were made for limited use, in case no suitable alternatives are available. For example, in the battle against malaria DDT can still be used to a limited extent for in-house spraying in Africa (Van den Berg, 2009). Until now all POPs are chemicals that contain carbon and halogen atoms. Some POPs, such as the PCDDs and PCDFs (together often short-named as dioxins) are not intentionally produced. They are formed and released unintentionally during thermal processes. PCDDs and PCDFs tended to be released by waste incinerators (Karasek and Dickson, 1987). The combination of elevated temperatures and the presence of chlorine from e.g. polyvinylchloride (PVC) led to the formation of the extremely toxic PCDDs and PCDFs. Stack emissions could contaminate entire areas around the incinerators with consequences for the quality of cow milk or local crop. Dioxins were first discovered after the Seveso (Italy) disaster (1976) when high quantities of dioxins were released after an explosion in a trichlorophenol factory (Mocarelli et al., 1991). Meanwhile, in many countries incinerators have been improved by changing the processes and installing appropriate filters.
Table 1. Stockholm Convention criteria for persistence, bioaccumulation, toxicity and long range transport of POPs.
Persistence
i
Evidence that the half-life of the chemical in water is greater than two months, or that its half-life in soil is greater than six months, or that its half-life in sediment is greater than six months; or
ii
Evidence that the chemical is otherwise sufficiently persistent to justify its
consideration within the scope of this Convention
Bioaccumulation
i
Evidence that the bio-concentration factor or bio-accumulation factor in aquatic species for the chemical is greater than 5,000 or, in the absence of such data, that the log Kow is greater than 5
ii
Evidence that a chemical presents other reasons for concern, such as high bioaccumulation in other species, high toxicity or ecotoxicity; or
iii
Monitoring data in biota indicating that the bio-accumulation potential of the
chemical is sufficient to justify its consideration within the scope of this Convention
Long range transport potential
i
Measured levels of the chemical in locations distant from the sources of its release
that are of potential concern
ii
Monitoring data showing that long-range environmental transport of the chemical,
with the potential for transfer to a receiving environment, may have occurred via air,
water or migratory species; or
iii
Environmental fate properties and/or model results that demonstrate that the chemical has a potential for long-range environmental transport through air, water or migratory species, with the potential for transfer to a receiving environment in locations distant from the sources of its release. For a chemical that migrates significantly through the air, its half-life in air should be greater than two days
Adverse effects
i
Evidence of adverse effects to human health or to the environment that justifies consideration of the chemical within the scope of this Convention; or
ii
Toxicity or ecotoxicity data that indicate the potential for damage to human health or
to the environment
Structures and use
Whereas all initial POPs were chlorinated chemicals and mainly pesticides, POPs that were added at a later stage also included brominated and fluorinated compounds and chemicals with a more industrial application. Brominated diphenylethers (PBDEs) and hexabromocyclododecane (HBCD) belong to the group of brominated flame retardants. These chemicals are being produced in high quantities. Many national legislations require the use of flame retardants in many materials, such as electric and electronic systems (TV, cell phones, computers), furniture and building materials. Although the PBDEs and HBCD have been banned in most countries, other brominated flame retardants are still being produced in annually growing volumes.
Figure 1. Structures of p,p’-DDT, 2,3,7,8-tetrachloro-p-dioxin, 2,4,2’,4’-tetrabromodiphenylether (a specific PBDE) and perfluoroctane sulfonic acid (PFOS).
Perfluorinated alkyl substances (PFASs) have many applications. Examples are Teflon production, use in fire-fighting foams, ski wax, as dirt and water repellant on outdoor clothes and carpets and many more. They are different from most other POPs because they are both lipophilic and hydrophilic due to a polar group present in most of the molecules. Examples of structures of a few POPs are given in Figure 1.
Persistence and bioaccumulation
The carbon-halogen bond is so strong that any type of degradation is unlikely to occur or will only occur on the long term and to a minor extent. Due to the size of the halogen atom, the strength of the halogen-carbon bond decreases from C-F to C-Cl, C-Br and C-I. In addition, these halogenated chemicals are lipophilic and, therefore, easily migrate to lipids, such as in living organisms. Because fish is a primary target, POPs enter the food chain in this way and biomagnification can occur (De Boer et al., 1998). High levels of POPs are, consequently, found in marine mammals (seals, whales, polar bears) and also in humans (Meironyte, 1999). Women may transfer a part of their POP load again to their children, the highest quantities to their firstborns.
Long range transport
Chemicals that migrate significantly through the air with a half-life in air greater than two days qualify for the POP criterion of long range transport. Many chemicals are indeed transported by air, often in different stages. Chemicals are emitted from a stack or evaporate from the soil in relatively warm areas and travel in the atmosphere toward cooler areas, condensing out again when the temperature drops. This process, repeated in ‘hops’, can carry them thousands of kilometers within days. This is called the ‘grasshopper effect’ (Gouin et al., 2004). It results in colder climate zones, in particular countries around the North Pole, receiving relatively high amounts of POPs.
Adverse environmental and health effects
There is very little doubt on the toxicity of POPs. Of course, the dose is always determining if a compound is causing an effect in the environment or in humans. POPs, however, are very toxic at very low doses. The Seveso disaster showed the high toxicity of dioxins for humans. Polybrominated biphenyls (PBBs) caused a high mortality in cattle when they were inadvertently fed with these chemicals (Fries and Kimbrough, 2008). Evidence of toxicity is often coming from laboratory studies with animals (in vivo) and more recently from in vitro studies. These studies are in particular important for the assessment of chronic toxicity. Many POPs are carcinogenic or act as endocrine disruptor.
Analysis
The analysis of POPs in environmental or human matrices is relatively complicated and costly. The compounds need to be isolated from the matrix by extraction. Subsequently, the extracts need to be cleaned from interfering compounds such as fat from organisms or sulphur in case of sediment or soil samples. Finally, due to the required sensitivity and selectivity, expensive instrumentation such as gas or liquid chromatography combined with mass spectrometry is needed for their analysis (Muir and Sverko, 2006). UN Environment is investing in large capacity building programs to train laboratories in developing countries in this type of analysis. According to the Stockholm Convention, countries shall manage stockpiles and wastes containing POPs in a manner protective of human health and the environment. POPs in wastes are not allowed to be reused or recycled. A global monitoring program has been installed to assess the effectiveness of the Convention.
Future
Much has to be done to achieve the original goals of eliminating the production and use of POPs and gradually reduce their spreading into the environment. A global treaty such as the Stockholm Convention with 182 countries involved is in a continuous challenge with procedures and political realities in countries, which hamper the achievement of perceived simple goals such as to eliminate the use of PCBs in 2025. The goals are, however, extremely important, as POPs are a global threat for current and future generations.
References
De Boer, J., Wester, P.G., Klamer, J.C., Lewis, W.E., Boon, J.P. (1998). Brominated flame retardants in sperm whales and other marine mammals - a new threat to ocean life? Nature 394, 28-29.
Fiedler, H., Kallenborn, R., de Boer, J., Sydnes, L.K. (2019). United Nations Environment Programme (UNEP): The Stockholm Convention - A Tool for the global regulation of persistent organic pollutants (POPs). Chem. Intern. 41, 4-11.
Fries, G.F., Kimbrough, R.D. (2008). The PBB episode in Michigan: An overall appraisal. CRC Critical Rev. Toxicol. 16, 105-156.
Gouin, T., Mackay, D., Jones, K.C., Harner, T., Meijer, S.N. (2004). Evidence for the “grasshopper” effect and fractionation during long-range atmospheric transport of organic contaminants. Environ. Pollut. 128, 139-148.
Karasek, F.W., Dickson, L.C. (1987). Model studies of polychlorinated dibenzo-p-dioxin formation during municipal refuse incineration. Science 237, 754-756.
Meironyte, D., Noren, K., Bergman, Å. (1999). Analysis of polybrominated diphenyl ethers in Swedish human milk. A time-related trend study, 1972-1997. J. Toxicol. Environ. Health Part A 58, 329-341.
Mocarelli, P., Needham, L.L., Marocchi, A., Patterson Jr., D.G., Brambilla, P., Gerthoux, P.M. (1991). Serum concentrations of 2,3,7,8‐tetrachlorodibenzo‐p‐dioxin and test results from selected residents of Seveso, Italy. J. Toxicol. Environ. Health 32, 357-366.
Muir, D.C.G., Sverko, E. (2006). Analytical methods for PCBs and organochlorine pesticides in environmental monitoring and surveillance: a critical appraisal. Anal. Bioanal. Chem. 386, 769-789.
Van den Berg H. (2009). Global status of DDT and its alternatives for use in vector control to prevent disease. Environ. Health Perspect. 117:1656–63.
2.2.5. Persistent Mobile Organic Chemicals (PMOCs)
Authors: Pim de Voogt
Reviewers: John Parsons, Hans Peter Arp
Leaning objectives:
You should be able to:
define a substance’s persistence
define a partition coefficient
understand the relationship between KOW, DOW, KD and mobility
understand the relationship between a substance’s mobility and persistence on the one hand and its potential for human exposure on the other
Keywords: Mobility, persistence, PMT
Introduction
Ecosystems and humans are protected against exposure to hazardous substances in several ways. These include treating our wastewater so that substances are prevented from entering receiving surface waters, and purification of source waters intended for drinking water production.
Currently, a majority of the drinking water produced in Europe is either not treated or treated by conventional technologies. The latter remove substances by degradation (physical, microbiological) or by sorption. However, chemicals that are difficult to break down and that can pass through soil layers, water catchments and riverbanks and cross natural and technological barriers may eventually reach the tap water. Typically, these chemicals are persistent and mobile.
Polarity
When the electrons in a molecule are unevenly divided over its surface, this results in an asymmetric distribution of charge, with positive and negative regions. Such molecules have electric dipoles (see Figure 1) and are polar, in contrast to molecules where the charge is evenly distributed thus resulting in the molecule being neutral or apolar. The ultimate form of polarity is when a permanent charge is present in a compound. Such chemicals are called ionogenic. We distinguish between cations (having a permanent positive charge, e.g. protonated bases, and quaternary amines) and anions (negatively charged ions, e.g. dissociated acids, and organosulfates). Ionic charges in molecules can be pH dependent (e.g. acids and bases). Most, and in particular small, polar and ionic chemicals are water soluble, in other words they have a strong affinity to water (often referred to as hydrophilic). Because water is one of the most polar liquids possible (a strong negative charge on the oxygen and two strong positive charges on each hydrogen), this means that for very polar organic molecules solvation by water is more favorable energetically then sorption to solid particles.
Chemicals that are nonpolar are inherently poorly water soluble and therefore tend to escape from the water compartment, resulting in their evaporation, or sorption to sediments and soils, or uptake and accumulation in organisms. It is therefore relatively easy to remove them from water during water treatment. In contrast, mobile organic chemicals, especially those that do not breakdown easily, pose a more serious threat for (drinking) water quality because they are much more difficult to remove. It should be noted that mobility and polarity can be thought of as a gradient, rather than a distinct category, with water being the most polar molecule, a large aliphatic wax being the most non-polar molecule, and all other organic molecules falling somewhere in the spectrum between.
In a recent study contaminants were analysed in Dutch water samples covering the journey from WWTP effluent to surface water to groundwater and then to drinking water. While the concentration level of total organic contaminants decreased by about 2 orders of magnitude from the WWTP effluents to the groundwater used for drinking water production, the hydrophilic contaminants (using chromatographic retention time as an indicator for hydrophilicity) in the WWTP effluents remained in the water throughout its passage to groundwater and into the drinking water (see Figure 2).
Figure 2.Average chromatographic retention time (tR) as a measure of average hydrophilicity of contaminants present in different water types. EFF, effluent; SW, surface water; GW, groundwater; DW, drinking water. Redrawn from Sjerps et al. (2016) by Wilma IJzerman.
Mobility and persistence
The mobility of chemicals in aquatic ecosystems is determined by their distribution between water and solid particles. The more the substance has an affinity for the solid phase the less mobile it will be. The distribution coefficient is known as KD, which expresses the ratio between the concentrations in the solid phase (soil, sediment, suspended particles), CS, and the dissolved phase at equilibrium, CW, i.e. KD = CS/CW. For neutral non-polar chemicals the distribution is almost entirely determined by the amount of organic carbon in the solid phase, fOC, and hence their distribution is usually expressed by KOC, the organic carbon-normalized KD (i.e. KOC = KD/ fOC). Unfortunately, there are relatively few reliable KD or KOC data available, in particular for polar chemicals. Instead, KOW is often used as a proxy for KOC. The n-octanol/water partition coefficient: KOW, is the equilibrium distribution coefficient of a chemical between n-octanol and water, KOW = Coctanol/Cwater. It's logarithmic value is often used as a proxy to express the polarity of a compound: a high log KOW means that the compound favors being in the octanol phase rather than in water, which is typically the case for a nonpolar compound. For ionizable chemicals we need to account for the pH dependency of KOW: at low pH an organic acid will become protonated (this in turn depends on its pKa value) and thus less polar. DOW is the pH-dependent KOW. It can be assumed that ions, whether cationic or anionic, will no longer dissolve into octanol but rather be retained in the water, because ions have much higher affinity for water than for octanol.
Accounting for this, for organic acids, the pH dependency of the DOW can be expressed as:
Therefore, as pH increases above the pKa, the smaller the DOW will get in the case of organic acids. In the case of basis, the opposite is true; the more the pH dips below the pKa of an organic base, the more cations form and the lower the Dow becomes.
However, one has to keep in mind that the assumption that the (log) KOW or DOW value inversely correlates with a compound’s aquatic mobility is, certainly, very simplistic. The behavior of an ionic solute will obviously also be determined by interactions i) with sites other than organic carbon, e.g. ionizable or ionic sites on soil and sediment particles, and ii) with other ions in solution.
The persistence of a compound is assessed in experimental tests by monitoring the rate of disappearance of the compound from the most relevant compartment. This is often done using standardized test protocols. In the European REACH legislation of chemicals, criteria have been established to qualify chemicals as persistent (P) and “very persistent”(vP) based on the outcomes of such tests. Table 1 presents the P and vP criteria used. Unfortunately, good-quality experimental data on half-lives are rare and obtaining such data requires both time-consuming and expensive testing.
Currently there is no certified definition of a compound’s mobility (M). Several possible compound properties have been proposed to characterize mobility, including a compound’s aqueous solubility or its KOC value. If (experimental) KOC values are not available, DOW values can be used as a proxy.
Table 1. P and vP criteria identical to Annex XIII to the REACH regulation (source: ECHA chapter R.11. Version 3.0, June 2017)
Persistent (P) in any of the following situations
Very persistent (vP) in any of the following situations
Freshwater
Half-life > 40 days
Half-life > 60 days
Marine water
Half-life > 60 days
Half-life > 60 days
Freshwater sediment
Half-life > 120 days
Half-life > 180 days
Marine sediment
Half-life > 180 days
Half-life > 180 days
Soil
Half-life > 120 days
Half-life > 180 days
Table 2.Proposed cut off values of compound properties proposed by the German Environmental Agency (UBA) to define substance mobility*
Mobile (M) if compound is P or vP and any of the following situations
very Mobile (vM) if compound is P or vP and any of the following situations
Lowest experimental log KOC
(at pH 4-9)
≤4.0
≤3.0
Log DOW
(at pH 4-9)
≤4.0 if no experimental
Log KOC data available
≤3.0 if no experimental
Log KOC data available
* note that the proposed criteria may change by the time of publication
Regulation and gaps in knowledge
The majority of chemicals for which international guidelines exist or that are identified as priority pollutants in existing regulations (e.g. EU Water Framework Directive and REACH), are nonpolar with log DOW values mostly above two (see Figure 3b). The German Ministry of Environment (UBA) has recently proposed to develop regulation for chemicals with P, M and toxic (T) properties (PMT substances) analogous to the existing PBT criteria used for regulation of chemicals in the EU. UBA proposed to use cut-off values of the KOC or DOW (if KOC data are not available) to define Mobile (M) or very mobile (vM) in conjunction with persistence criteria (see Table 2). Note that the KOC and DOW values have to be obtained from testing at an environmentally relevant pH range (pH 4-9).
Figure 3.Box and whisker plots of calculated logDOW values at pH 7.4 of: (a) contaminants in water analyzed by either GC-MS or LC-MS and of examples of mobile chemicals; (b) contaminants regulated by the Stockholm Convention; candidates of Substances of Very High Concern (SVHCs) according to REACH, Article 57 d−f; the list of priority substances according to the Water Framework Directive (WFD); and the so-called Watch List of the WFD. The whiskers point to the 10th and 90th percentile. Numbers in (a) refer to 1: Aminomethylphosphonic acid (AMPA), 2: Paraquat, 3: Cyanuric acid, 4: N,N-dimethylsulfamide (DMS), 5: Diquat, 6: 5-Fluorouracil, 7: Glyphosate, 8: Melamine, 9: Metformin, 10: Perfluoroacetic acid, 11: EDTA. Redrawn from Reemtsma et al. (2016) by Wilma IJzerman.
When we consider current analytical techniques used for monitoring contaminants in the environment, it can be readily seen that the scope of techniques most often used (gas chromatography, GC, and reversed-phase liquid chromatography, RPLC) do not overlap with what is required for chemicals having log DOW values typical of the most mobile chemicals, having a log Dow below zero (see Figure 3a). Consequently, there is limited information available on the occurrence and fate of these mobile chemicals in the environment. Nevertheless, some examples of persistent and mobile chemicals have been identified. These include highly polar pesticides and their transformation products, for instance glyphosate and aminomethylphosphonic acid (AMPA), short-chain perfluorinated carboxylates and sulfonates, quaternary ammonium chemicals such as diquat and paraquat and complexing agents such as EDTA. There are, however, likely to be many more chemicals that could be classified as PMOC and we can therefore conclude that there is a gap in the knowledge and regulation of persistent mobile organic chemicals.
References
Arp, H.P.H., Brown, T.N., Berger, U., Hale, S.E. (2017). Ranking REACH registered neutral, ionizable and ionic organic chemicals based on their aquatic persistency and mobility. Environmental Science: Processes Impacts 19, 939-955.
Reemtsma, T., Berger, U., Arp, H. P. H., Gallard, H., Knepper, T. P., Neumann, M., Benito Quintana, J., de Voogt, P. (2016). Mind the gap: persistent and mobile organic chemicals - water contaminants that slip through. Environmental Science & Technology 50, 10308-10315.
Sjerps, R.M.A., Vughs, D., van Leerdam, J.A., ter Laak, T.L., van Wezel, A.P. (2016). Data-driven prioritization of chemicals for various water types using suspect screening LC-HRMS. Water Research 93, 254-264.
2.2.6. Ionogenic organic chemicals
Authors: Steven Droge
Reviewer: John Parsons, Satoshi Endo
Leaning objectives:
You should be able to:
understand that IOCs are abundantly present in many types of chemical classes, but all share common features such as dissociation of a proton from a polar moiety (acids to form anions) and association of a proton onto a polar moiety (bases to form cations).
calculate the fraction of neutral/ionic species for each chemical at a given pKa and pH.
make some rough predictions about pKa values from the IOC molecular structure.
Ionogenic organic chemicals (IOCs) are widely used in industry and daily life, but also abundantly present as chemicals of emerging concern. For environmental risk assessment purposes, IOCs may be defined as organic acids, bases, and zwitterionic chemicals that under common environmental pH conditions exist for a large part as charged (ionic) species, with only modest fraction of neutral species. The environmentally relevant pH range could be argued to lie between 4 (acidic creeks, even lower in polluted streams from volcanic regions or mine drainage systems) to 9 (sewage treatment effluents). The environmental behaviour of IOC pollutants of concern are different from neutral chemicals of concern, because the aqueous pH controls the neutral fraction of dissolved IOCs and the ionic form is highly soluble and interacts partly via electrostatic interactions with environmental susbtrates. Note that the major fraction of an IOC can also be neutral in a certain environmental system, and in that case it is often the neutral form that dominates the chemical’s behavior.
Figure 1.A survey on the 1999 World Drug Index (WDI) database revealed that >63% of the 51,596 chemicals had acidic or basic functionalities. Shown in the circle diagram A is the distribution of different ionizable groups in this fraction of 63% drugs, and in bar diagram B the distribution in pKa values for basic drugs (CNS = drugs acting on central nervous system). Adapted from Manallack (2007) by Wilma IJzerman.
IOCs are common in many different types of pollutant classes. A random subset analysis of all EU (pre-)registered industrial chemicals indicated that large fractions of the total list of >100,000 chemicals are IOCs (51% neutral; 27% acids; 14% bases; 8% zwitterions/amphoterics). In another source, It has been estimated that >60% of all prescription drugs (Section 2.3.3) are IOCs (Manallack, 2007), and even higher fractions for illicit drugs (Section 2.3.4) (Figure 1). Well known examples are basic beta-blockers (e.g. propranolol), basic antidepressants (e.g. fluoxetine and sertraline), acidic non-steroidal anti-inflammatory drugs (NSAID such as diclofenac), and basic opioids (e.g. morphine, cocaine, heroin) and basic designer drugs (e.g. MDMA). The majority of surfactants and polyfluorinated chemicals (e.g. PFOS and GenX) are IOCs (Section 2.3.8), as well as wide variety of important pesticides (e.g. zwitterionic glyphosate) (Section 2.3.1) and (natural) toxins (Section 2.1) (e.g. peptide based multi-ionic cyanobacterial toxins).
Environmental behavior of IOCs
The release into the environment is specific for each of these types of IOCs with a different use, but in many cases happens via sewage treatment systems. If sorption to sewage sludge is very strong, application of sludge onto terrestrial (agricultural) systems is a key entry in many countries. However, many IOCs are rather hydrophilic and will mainly be present in wastewater effluent released into aquatic systems. As they are hydrophilic, they are considered rather mobile which allows for rapid transport through e.g. groundwater plumes, soil aquifers, and (drinking water) filtration steps. The distinction between (mostly) neutral chemicals and IOCs is important because the ionic molecular form generally behaves very differently from the corresponding neutral molecular form. For example, in many aspects ionic molecules are non-volatile compared to the corresponding neutral molecules, while neutral molecules are more hydrophobic than the corresponding ionic molecules. As a result of their lower “hydrophobicity”, ionic molecules often bind with lower affinity to soils and are therefore more mobile in the environment. The ionic forms bioaccumulate to a lower extent and can therefore be less toxic than the corresponding neutral form (though not necessarily). However, there are various important exceptions for these rules. For example, clay minerals sorb cationic IOCs fairly strongly via ion exchange mechanisms. Certain proteins (e.g. the blood serum protein albumin) tightly bind anionic chemicals because of cationic subdomains on specific (enzymatic) pockets, which allows for effective transport throughout our systems and over cell membranes.
Calculating and predicting the dissociation constant (pKa)
The critical chemical parameter describing the ability to ionize is the acid dissociation constant (pKa). The pKa defines at which pH 50% of the IOC is in either the neutral or ionic form by releasing an H+ from the neutral molecule acids (AH to anion A-), or accepting an H+ onto the neutral molecule base (B to cation BH+). The equilibrium between neutral acid and dissociated form can thus be defined as:
[ HA ] ↔ [H+] + [A- ] (eq.1)
where the chemical’s equilibrium speciation is defined as:
\(K_a = {[H^{+}] [A^-]\over [HA]}\) (eq.2)
which gives the pKa as:
\(pK_a = - log(K_a)\) (eq.3)
and as a function of pH, the ratio of the acid and anion is defined as:
\(pH = pK_a\ + log\ {[A^-]\over [HA]}\) for acids and \(pH = pK_a\ + log\ {[B]\over [BH^+]}\) for bases (eq.4)
Although the term pKb is also used to denote the base association constant, it is conventional we consider [BH+] as acid and use ‘pKa’ and other relationships for bases as well. The fraction of neutral species (fN) for simple IOCs (one acidic or basic site) can be readily calculated with a derivatization of the Henderson-Hasselbalch equation:
\(f_N = {1\over 1+10^{\alpha(-pH+pK_a)}}\) (eq.5)
in which α = 1 for bases, and -1 for acids.
in which α = 1 for bases, and -1 for acids. Using equation 5, Figure 2 presents a typical speciation profile for an acid (shown with pKa 5, so perhaps a carboxylic acid) and a base (shown with pKa 9, so perhaps a beta-blocker drug). Following the curve of equation 5, it is interesting to see some simple rules: if the pH is 1 unit lower than the pKa, the deprotonated species fraction is present at 10%. If the pH is 2 units lower than the pKa, the deminished species fraction is present at 1%, 3 units lower would give 0.1%, etc. From this, it is easy to make a good estimate for the protonation of a strong basic drug like MDMA (pKa reported 9.9-10.4) in blood (pH 7.4): with a maximum of 3 units lower pH, up to 99.9% of the MDMA will be in the protonated form, and only 0.1% neutral. For toxicological modeling studies, e.g. in terms of permeation through the blood-brain barrier membrane, this is highly relevant.
Figure 2.pH dependent speciation of an acid (with a pKa of 5) and a base (with a pKa of 9). As shown by the arrows, if pH = pKa, the IOCs exist 50% in the neutral form, and 50% in the ionic form. For the acid at pH = pKa -1 (pH 4), 90% is in the neutral form (AH), and 10% is in the negatively charged form (A-). Drawn by Steven Droge.
Boxes 1-3. Extended learning: calculating the dissociation constant for multiprotic chemicals:
see end of this module
Acidic IOCs:
Figure 3.Different types of anionic moieties: carboxylate (anionic form of carboxylic acid), sulfonate (anionic form of sulfonic acid), sulfate (anionic form of sulfuric acid), phenolate (anionic form of phenol). (Source: Steven Droge)
For example, the painkiller (or non-steroidal anti-inflammatory drug, or NSAID) diclofenac is a carboxylic acid with a pKa of 4.1. This means that at pH 4.1, 50% of the dissolved diclofenac is in the dissociated (anionic) form (so, (1 - fN) from equation 5). At pH 5.1 (1 unit higher than the pKa) this is roughly 90% (90.91% to be more precise, but simply remembering 90% helps), at pH 6.1 (2 units higher than the pKa) this is 99%. This stepwise increase in 50-90-99% with each pH unit works for all acids, and for bases the other way round. Test for yourself that at physiological pH of 7.4 (e.g. in blood) diclofenac is calculated to be 99.95% anionic.
Many carboxylic acids have a pKa in the range of 4-5, but the neighboring molecular groups can affect the pKa. Particularly electronegative atoms such as chlorine, fluorine, or oxygen may lower the pKa, as they reduce the forces holding the dissociating proton to the oxygen atom. For example, trichloroacetic acid (CCl3-COOH) has a pKa of 0.77, while acetic acid (CH3-COOH) has a pKa of 4.75. For the same reason, perfluorinated carboxylic acids have a strongly reduced acidic pKa compared to the analogous non-perfluorinated carboxylic acids.
Sulfate acids (see figure 3) are very strong acids, with a pKa <0. These acids almost always occur in their pure form as a salt, for example the common soap ingredient sodium dodecylsulfate (“SDS” or “SLS”, Na.C12H25-SO4). Other common detergents are sulfonates, such as linear alkylbenzenesulfonate (“LAS”, C10-14-(benzyl-SO3)), where the SO3 anionic moiety is attached to a benzene ring, which can be positioned to different carbon atoms of a long alkyl chain. Even at the lower environmental pH range of about 4 these soap chemicals are fully in the anionic form. Such very strong acids, but also many weaker acids, are thus often sold in pure form as salts with sodium, potassium, or ammonium, which causes them to have different names and CAS numbers (e.g. Na.C12H25-SO4 or K.C12H25-SO4) than the neutral form.
Many phenols have a pKa > 8, and are therefore mostly neutral under environmental pH. Electron-withdrawing groups on the aromatic ring of the phenol group, such as Cl, Br and I, can lower the dissociation constant. For example, the dinitrophenol-based pesticide dinoseb has a pKa of 4.6, and is thus mostly anionic in the aquatic environment. Note that a hydroxyl group (-OH) not connected to an aromatic ring, such as -OH of alcohols, can in most cases for risk assessment be considered permanently neutral.
To help interpret the differences in pKa between molecules, it sometimes helps to remember that in more acidic solutions, there are simply higher H+ concentrations, in a logarithmic manner on the pH scale. At pH 3, the concentration H+ protons in solution is 1 mM, while at pH 9 the H+ concentration is 1 nM (6 pH units equals 106 times lower concentrations). The affinity (“Ka”) of H+ to associate with a negatively charged molecular group, is so low for strong acids that even at very high dissolved H+ concentrations (low pH) only very few AH bonds (neutral acid fraction) are actually formed. In other words, for chemicals with a low Ka, even at low pH the neutral fraction is still low. For weak acids such as phenols, already at a very low dissolved H+ concentrations (high pH) many AH bonds (neutral acid fraction) are formed. So it can be reasoned that the affinity of common acidic groups to hold on to a proton is in the order:
Figure 4.Different types of cationic moieties: primary amine, secondary amine, tertiary amine, quaternary ammonium (permanently charged), pyridinium (cationic form of pyridine), alkylpyridinium (permanently charged). (Source: Steven Droge).
For bases, it is mostly a nitrogen atom that can accept a proton to form an organic cation, because of the lone electron pair in nitrogen. Neutral nitrogen atoms have opportunities for 3 bonds. A primary amine group has the nitrogen atom bonding to only 1 carbon atom (represented here as part of a molecular fragment R), and two additional bonds with hydrogen atoms. The remaining electron pair readily accepts another proton to form a cationic molecule [ R-NH3+ ]. Neutral secondary amines have one bond to hydrogen and two bonds to carbon atoms and can accept a proton to form [ R-NH2+-R’ ], whereas neutral tertiary amines have no bonds to hydrogen but only to carbon atoms and can form [ R-NH+-(R’)(R’’) ]. Of course, each R group may be the same (e.g. a methyl unit).
Many basic chemicals have complex functionalities that can influence the pKa of the nitrogen moiety. However, as shown in the examples of figure 5, as long as there are at least two carbon atoms between the amine and the polar molecular fragment (for example OH, but much stronger for =O), the pKa of the basic nitrogen groups in all three types of bases (primary, secondary and tertiary amines) is high, often above 9 (dissolved H+ concentration <10-9). So even at very low H+ concentrations, dissolved protons like to be associated to such amine groups. As a result, amines such as most beta-blockers and amphetamine based drugs are predominantly positively charged molecules (organic cationic amines) in the common environmental pH range of 4-9, as well as in the pH of most biotic tissues that are useful for toxicological assessments. As soon as a polar group with oxygen (e.g. ketones or hydroxyl groups) is connected to the second carbon away from the nitrogen (e.g. R-CH(OH)-CH2-NH2) the pKa is considerably lowered. Also nitrogen atoms as part of an aromatic ring, or connected to an aromatic ring, have much lower pKa’s: protons have rather low affinity to bind to these N-atoms and only start doing so if the proton concentration becomes relatively high (solution becomes more acidic).
Figure 5.Different proton dissociation constants for amine groups: pKa is influenced by other functional structures. (Source: Steven Droge)
Relevance of accounting for electrostatic interactions
Most classical pollutants, such as DDT, PCBs and dioxins, are neutral hydrophobic chemicals. On the other hand, most metals are almost always cationic species (e.g. Cd2+). Consequently, their environmental distribution and biological exposure are influenced by quite distinct processes. Obviously, predominantly charged IOCs behave somewhat in between these two extremes. The charged positive or negative groups cause a strong effect of electrostatic interactions between the IOC and environmental substrates (sorption or ligand/receptor binding). While also metals speciate into different forms, pH difference between environmental compartments can strongly influence the IOCs chemical fate and effect if ionizable group is relatively weak. An important difference to metals is that the nonionic molecular part still influences the IOC’s hydrophobicity even in charged state for several processes.
As will be discussed in other chapters regarding chemical processes (see Chapter 3), it needs to be taken into account for IOCs that many environmental substrates (DOC, soil organic matter, clay minerals) are mostly negatively charged in the common range of environmental pH, but also that proteins involved in biotic uptake-distribution-effects are rich in ionogenic peptides that are part of binding pockets and reactive centers.
References
Manallack, D.T. (2007). The pKa distribution of drugs: application to drug discovery. Perspectives in Medical Chemistry 1, 25-38.
Box 1. Extended learning: calculating the dissociation constant for multiprotic chemicals:
Several common inorganic acids are multiprotic: they have multiple protons that can dissociate.
Multiples species can occur at a certain pH, such as for phosphoric acid (H3PO4, H2PO4-, HPO42- , HPO42-), and carbonic acid (H2CO3, HCO3-, CO32-). It is important to realize that there are actually two micro-species of HCO3-, because two hydroxylgroups can dissociate: HO-C(=O)-OH
A polyprotc acid HnA can undergo n dissociations to form n+1 species. Each dissociation has a pKa.
But how to calculate the fraction of each species of multiprotic chemicals?
The charge of a polyprotic acid can be described as Hn-jAj-. A useful variable, v, can be defined for each general polyprotic acid:
The degree of dissociation of the acid (η) is equal to the ratio of the total charge (TC) to the total mol acid (TM). For a diprotic acid, a plot of η as a function of pH provides the dissociation curve:
You can set up such a calculation in MS Excel, with calculations of α0, α1, α2, at a range of different pH values ([H+] concentrations), for a given K1 and K2, and plot the speciation against pH.
More details are described by King et al. (1990) J. Chem. Educ. 67 (11), p. 932; DOI: 10.1021/ed067p932
Box 2. Example 1 for multiprotic chemicals: Carbonic acid
Let’s try carbonic acid (bicarbonate, or H2CO3) as an example first. H2CO3 is the product of carbon dioxide dissolved in water. In pure water/seawater the hydration equilibrium constant Kh = [H2CO3]/[CO2] ≈ 1.7×10−3 / ≈ 1.2×10−3 respectively, indicating that only 0.1% of dissolved CO2 equilibrates to H2CO3. The dissolved concentration of CO2 depends on atmospheric CO2 levels according to the air-water distribution coefficient (Henry constant kH = pCO2/[CO2]= 29.76 atm/(mol/L)). Because of the relevance of CO2 in e.g. ocean acidification, and gas exchange in our lungs, it is interesting to see how H2CO3 speciates depending on pH. As in the formula HnA, n= 2 for H2CO3.
Box 3. Example 2 for multiprotic chemicals: Zwitterions
Many organic pH buffers are zwitterionic chemicals, that contain both an acidic and a basic group.Norman Good and colleaguesdescribed a set of 20 of such buffers forbiochemicalandbiologicalresearch (see for example www.interchim.fr/ft/0/062000.pdf, or www.applichem.com/fileadmin/Broschueren/BioBuffer.pdf). Examples are MES, MOPS, HEPPS (FigureA). These buffers are selected to:
have a buffering pKa in the range of pH6-8 where most biochemical tests are performed;
be readily soluble in water,
be stable in test solutions, so resistant to (non)enzymatic degradation, not forming precipitates with salts
ideally be impermeable to cell membranes so that they don’t accumulate or reach active intracellular sites
readily available, reasonably cheap.
the zwitterionic chemicals with sulfate groups are actually always have the sulfate group charged, making it highly soluble and impermeable to cell membranes, while the amine group protonates between pH6-10, depending on neighbouring functional groups. The speciation of the amine groups in MES and MOPS simply follows the single pKa calculation of equations 1-5. In HEPPS, either of the two amines is protonated, the second pKa is 3, so the doubly charged molecules only occurs at much lower pH, but can still be used as a buffer.
Figure A. MES, MOPS and HEPPS in charged form.
A zwitterionic chemical with two apparent pKa values relatively close is p-amino-benzoic acid. If chemicals have not one ionisable group, but N ionisable groups that speciation in a relevant pH range, than the amount of possible species is 2N. So the zwitterion p-amino benzoic acid has 4 species, each with a separate pKa (pH where both species are present in equal concentrations).
Let’s formulate the benzyl group in p-amino benzoic acid as X, the neutral amine base as B, and the neutral carboxylic acid as AH, so that the fully neutral species is BXAH.
Compared to the carboxylic acid, we now have under the most acidic conditions (BHXAH)+, as \(α_0\), the neutral species BXAH and the zwitterionic intermediate (BHXA)0, as \(α_1\), and anionic species at most alkaline conditions (BXA)-, as \(α_2\). The calculation of the fraction of each species can be calculated according to similar rules as for carbonic acid if the two dissociation constants are known (K1 = 2.4, K2=4.88).
However, this does not inform us on the ratio between the zwitterionic form and the fully neutral form. To do this, the speciation constants of the 4 microspecies are required.
[BH+XAH] ↔ [BXAH] + [H+] for which the pk1 is calculated to be 2.72
K1=10^-2,72 = [BXAH]*[H+]/ [BH+XAH]
[BH+XAH]*10^-2.72 = [BXAH]*[H+]
which rearranges to [BXAH]= 10^-2.72 *[BH+XAH] /[ H+]
[BH+XAH] ↔ [BH+XA-] + [H+] for which the pk2 is calculated to be 3.93
K2=10^-3.93 = [BH+XA-]*[H+] / [BH+XAH]
[BH+XAH]*10^-3.93 = [BH+XA-]*[H+]
which rearranges to [BH+XA-] =10^-3.93*[BH+XAH]/[H+]
[BXAH] ↔ [BXA-] + [H+] for which the pk3 is calculated to be 4.74
K3= 10^-4,74 = ([BXA-] * [H+]) / [BXAH]
[BXAH]*10^-4,74 = [BXA-] * [H+])
[BH+XA-] ↔ [BXA-] + [H+] for which the pk4 is calculated to be 4.31
K4=10^-4,31 = [BXA-]*[H+] / [BH+XA-]
[BH+XA-]*10^-4,31 = [BXA-]*[H+]
So the ratio between zwitterionic form [BH+XA-] and neutral form [BXAH] equals to:
[BH+XA-] / [BXAH] = 10^-pK2 / 10^-pK1
[BH+XA-] / [BXAH] = 10^-3.93 / 10^-2.72 = 0.06: so only 6% zwitterionic vs 94% neutral species.
Explain how complex mixtures/UVCBs differ from well-defined substances.
Explain the challenges and uncertainties in the risk and hazard assessment of complex mixtures/UVCBs.
Key words: UVCB; Complex substances; Constituents
Introduction/type of substances
Beside substances that consist of a single chemical structure, there are also substances produced that contain multiple constituents each having its unique molecular structure. In general, three types of substances can be identified: 1) mono-constituent substances, 2) multi-constituent substances, and 3) UVCBs that are substances of Unknown or Variable composition, Complex reaction products or Biological materials (ECHA, 2017a). Mono-constituent substances contain one main constituent that makes up at least 80% of the substance (Figure 1A), whereas multi-constituent substances consist of several main constituents that are present at a concentration between 10% and 80% (Figure 1B). Potential other constituents within these substances are considered impurities (Figure 1A&B). These first two substance categories are sometimes also described as well-defined substances, as the composition is (or can be) well characterized. However, this section will specifically focus on the third category, the UVCB substances. UVCBs contain many different constituents of which some can be (partially) unknown and/or the exact composition can be variable or difficult to predict (Figure 1C). Principally, none of the constituents in a UVCB are considered as impurities. Although different terms are used to define UVCBs / complex chemical substances within various regulatory frameworks (Salvito et al., 2020), the term ‘UVCB’ will be used throughout this section to represent these various denominations.
Figure 1. Overview of the three main types of substances. A) mono-constituent substance. B) multi-constituent substance. C) UVCB substance. Pictures are derived from ECHA What is a substance? - ECHA (europa.eu).
Types and naming of UVCB substances
Several types of UVCB substances can be defined, including UVCBs that are synthesized or derived/refined from biological or mineral sources (ECHA, 2017a). Common types of UVCBs include:
Extracts from plant or animal products (e.g. ‘Lavender, Lavandula hybrida, ext.’ [CAS 91722-69-9; EC 294-470-6]).
Reaction products that are formed during a chemical reaction (e.g. ‘Reaction product of 1,3,4-thiadiazolidine-2,5-dithione, formaldehyde and phenol, heptyl derivs.’ [EC 939-460-0]).
Products that are derived from industrial processes (e.g. ‘Naphtha (petroleum), catalytic reformed’ [CAS 68955-35-1; EC 273-271-8]).
In general, the name of a UVCB substance is a combination of its source (e.g. name of the species for biological sources, or name of the starting material for non-biological sources) and the used process(es) (e.g. extraction, fractionation, etc.) (ECHA, 2017a). In addition, for some UVCB categories, specific nomenclature systems are developed that can also include a description of the general composition or characteristics (e.g. physicochemical properties, like boiling range). A specific nomenclature system is for instance developed for hydrocarbon solvents and oleochemicals, in which the nomenclature is based on the chemical composition (e.g. ‘Hydrocarbons, C9-C11, n-alkanes, isoalkanes, cyclics, < 2% aromatics [CAS 64742-48-9; EC 919-857-5]’) (OECD, 2014; OECD, 2015).
Challenges in the risk and hazard evaluation of UVCBs
It is difficult to fully characterize the chemical composition of UVCBs as they can contain a relative large number of constituents. Generally, it is technically challenging or impossible to identify, and thus to test, all individual constituents present in a UVCB. As a consequence, a significant fraction is often defined as ‘unknown’ or is only specified in general terms (ECHA, 2017a,b). Nevertheless, specific information down to the individual constituent level could be relevant for risk/hazard assessment as some constituents can already cause effects at low concentrations (see chapter 6). In addition to an ‘unknown’ fraction, UVCBs can also have a variable composition. The composition may for instance depend on fluctuations in the manufacturing processes and/or source material, including spatial and temporal variations. Although this variability may not affect the functionality of the UVCB substance, it could influence and warrant a new hazard assessment (ECHA, 2017a,b; Salvito et al., 2020). Obviously, these key characteristics of UVCBs (i.e. the compositional complexity), complicate their risk and hazard assessment.
As it is not possible to identify, isolate and assess all individual constituents, alternative assessment approaches are being developed to evaluate UVCBs, including whole-substance and constituent based approaches. Within whole-substance based approaches, the UVCB is used as test-item. Testing of the whole substance might be relevant when the UVCB consists of structurally very similar constituents that are expected to have comparable fate and effect properties (Figure 2). However, when the UVCB displays a wide range of physicochemical properties a constituent based approach is generally preferred for risk/hazard assessment purposes, as the results of whole substance testing can be very difficult to interpret. For instance, the results of whole substance testing typically provide a single profile for the whole UVCB, while the fate, behavior and effects between (groups of) constituents could differ significantly. Furthermore, result interpretation might be challenging, as it could be difficult to maintain stable dosing/exposure concentrations when constituents with varying physicochemical properties are combined (e.g. due to differences in sorption, evaporation, solubility etc.).
Within constituent-based approaches, generally one or a few representative constituents are selected and evaluated. The results for these constituents are subsequently extrapolated to the other constituents, and ultimately to the UVCB. The selection of representative constituents can be based on several aspects, including in silico predictions of fate and hazard properties, the relevance or availability of the constituents and the structural variability (ECHA, 2017b; Salvito et al., 2020). To support the generation and selection of representative constituents computational methodologies are being developed (Dimitrov et al., 2015; Kutsarova et al., 2019).
One of the best described constituent based approaches is the ‘fraction profiling approach’, which is also known as the hydrocarbon block method (ECHA, 2017b; King, 1996). This method is specifically developed for petroleum substances (although it may also be applied to other UVCBs) and is applied in several risk assessment and PBT (Persistent, Bioaccumulative, Toxicity) assessment approaches (CONCAWE, 2016; Wassenaar and Verbruggen, 2021). Within the hydrocarbon block method, the composition of a UVCB is conceptually divided in blocks/fractions of (structurally) similar constituents. The underlying assumption is that all constituents within a block have fairly similar properties and could be assessed as if it is ‘one constituent’. More details on the hydrocarbon block method are provided in section 2.3.5 Hydrocarbons.
In general, the choice of the assessment approach is dependent on the substance and may also be dependent on the data already available as well as the stage and the general purpose of the assessment (e.g. PBT-assessment, risk assessment, etc.). In some cases, a combination of varying approaches could be considered most efficient. For instance by using a whole substance as test item in combination with analytical measurements of individual constituents over time.
Figure 2. UVCB substances X and Y contain constituents with varying physicochemical properties (colors). For substance X a whole-substance based approach might be used, whereas for substance Y a constituent based approach is generally preferred. The different shapes represent constituents that could potentially be grouped according to other properties, such as mode of toxic action. The ‘?’ represents an ‘unknown’ fraction. This figure is adopted and modified from Salvito et al. (2020).
ECHA (2017a). Guidance for identification and naming of substances under REACH and CLP. European Chemicals Agency, Helsinki.
ECHA (2017b). Guidance on Information Requirements and Chemical Safety Assessment. Chapter R.11: PBT/vPvB assessment. European Chemicals Agency, Helsinki.
King, D.J., Lyne, R.L., Girling, A., Peterson, D.R., Stephenson, R., Short, D. (1996). Environmental risk assessment of petroleum substances: the hydrocarbon block method. CONCAWE report no. 96/52. Brussels.
Kutsarova, S.S., Yordanova, D.G., Karakolev, Y.H., Stoeva, S., Comber, M., Hughes, C.B., Vaiopoulou, E., Dimitrov, S.D., Mekenyan, O.G. (2019). UVCB substances II: Development of an endpoint-nonspecific procedure for selection of computationally generated representative constituents. Environmental Toxicology and Chemistry 38, 682-694. https://doi.org/10.1002/etc.4358.
OECD. (2015). OECD guidance for characterising hydrocarbon solvents for assessment purposes. Series on Testing and Assessment, No. 230. ENV/JM/MONO (2015)52. Organization for Economic Cooperation and Development, Paris.
OECD. (2014). OECD guidance for characterising oleochemical substances for assessment purposes. Series on Testing and Assessment, No. 193. ENV/JM/MONO(2014)6. Organization for Economic Cooperation and Development, Paris.
Salvito, D., Fernandez, M., Jenner, K., Lyon, D.Y., De Knecht, J., Mayer, P., MacLeod, M., Eisenreich, K., Leonards, P., Cesnaitis, R., León-Paumen, M., Embry, M., Déglin, S.E. (2020). Improving the environmental risk assessment of substances of unknown or variable composition, complex reaction products, or biological materials. Environmental Toxicology and Chemistry 39, 2097-2108. https://doi.org/10.1002/etc.4846.
Wassenaar, P.N.H., Verbruggen, E.M.J. (2021). Persistence, bioaccumulation and toxicity-assessment of petroleum UVCBs: A case study on alkylated three-ring PAHs. Chemosphere 276, 130113. https://doi.org/10.1016/j.chemosphere.2021.130113.
2.2.8. Plastics
(draft)
Author: Ansje Löhr
Reviewer: John Parsons
Leaning objectives:
You should be able to:
indicate the relevance of plastic for society
describe the main characteristics of (micro)plastics
describe the main ecological effects of (micro)plastics
Keywords: Plastic types, sources of plastics, primary and secondary microplastics, plastic degradation, effects of plastics
Introduction
Since its introduction in the 1950s, the amount of plastics in the environment has increased dramatically (Figure 1). A recent study by Jambeck et al. (2015) estimated that 192 coastal countries generated 275 million metric tonnes of plastic waste in 2010 of which around 8 million tons of land-based plastic waste ends up in the ocean every year. By UN Environment plastic pollution is seen as one of the largest environmental threats. If waste management does not change rapidly, another 33 billion tonnes of plastic will have accumulated around the planet by 2050. (Micro)plastics is widely recognized as a serious problem in the ocean, however, plastic pollution is also seen in terrestrial and freshwater systems.
Classification by size and morphology
Plastics are commonly divided into macroplastics and microplastics; the latter plastic particles are <5 mm in diameter (including nanoplastics). There are several ways to classify microplastics but the following two types are often used; primary microplastics and secondary microplastics. Primary microplastics have been made intentionally, like pellets or microbeads, secondary microplastics are fragmented parts of larger objects. Microplastics show a large variety in characteristics such as size, composition, weight, shape and color. These characteristics have an influence on the behaviour in the environment, like for instance, the dispersion in water and the uptake by organisms (Figure 2). Low-density particles float on water and are therefore more prone to advection than particles with a
higher density. Similarly, spheres are more likely to be taken up by organisms than fibers. The characteristics also affect the absorption of contaminants, adsorption of microbes, and potential toxicity.
Figure 1. Global plastic production and future trends (source: http://www.grida.no/resources/6923; 2019, GRID-Arendal & Maphoto/Riccardo Pravettoni).
Figure 2. Marine litter comes in all sizes. Large objects may be tens of metres in length, such as pieces of wrecked vessels, lost. (Source: http://www.grida.no/resources/6924; 2019; GRID-Arendal & Maphoto/Riccardo Pravettoni).
Classification by chemistry
Plastic is the term used to define a sub-category of the larger class of materials called polymers, usually synthesized from fossil fuels, although biomass and plastic waste can also be used as feedstock. Polymers are very large molecules that have characteristically long chain-like molecular architecture. There are many different types of plastics but the market is dominated by 6 classes of polymers: polyethylene (PE, high and low density), polypropylene (PP), polyvinyl chloride (PVC), polystyrene (PS, including expanded EPS), polyurethane (PUR) and polyethylene terephthalate (PET) (figure 3). In order to make materials flexible, transparent, durable, less flammable and long-lived, additives to polymers are used such as flame retardants (e.g. polybrominated diphenyl ethers), and plasticisers (e.g. phthalates). Some of these substances are known to be toxic to marine organisms and to humans.
Figure 3. Different types of plastics (source https://isustainrecycling.com/plastics-recycling/)
Biopolymers/ bioplastics
There is a lot of discussion on bioplastics as degradable plastics that these may still persist for a long time under marine conditions. Please watch this video by dr. Peter Kershaw.
Plastic degradation
Degradation of plastics takes place as soon as the plastic loses its original integrity and properties. There is a faster breaking up phase (degradation into microparticles) and a much slower mineralization phase (polymer chains being degraded to carbon dioxide). The degradation rate of plastics is determined by its polymer type, additive composition and environmental factors. Many commonly-used polymers are extremely resistant to biodegradation. Although plastics degrade in natural environments it is argued that no polymer can be efficiently biodegraded in a landfill site. Plastics in aquatic environments can be subject to in-situ degradation, e.g. by photodegradation or mechanical fragmentation but are in general very durable. As a result, plastics that are present in our oceans will degrade at a very slow pace, (Figure 4). So the majority of plastics produced today will persist in the environment for decades and probably for centuries, if not millennia.
Figure 4. “How long until it is gone” : the time required to degrade different materials. (source https://futurism.com/plastic-decomposition)
Plastics in the environment
Sources and pathways:
Plastics are found in terrestrial, freshwater, estuarine, coastal and marine environments, and even in very remote areas of the world and the deep-sea. Sources and pathways of marine litter are diverse and exact quantities and routes are not fully known. But there is a surge in interest to determine the exact quantities and types of plastic litter and pathways in the environment and most of the plastic in our oceans originates from land-based sources (Figure 5) but also from sea-based sources. Most PE and PP is used in (single-use) packaging products that have a short lifetime and end up soon as waste.
Figure 5. Overview of the major sources of primary microplastics and the generation of secondary microplastics (Source: http://www.grida.no/resources/6929; 2019; GRID-Arendal & Maphoto/Riccardo Pravettoni).
Primary microplastics in terrestrial environments mostly originate from the use of sewage sludge containing microplastics from personal care or household products. In agricultural soils the application of sewage sludge from municipal wastewater treatment plants to farmland is probably a major input, based on recent MP emission estimates in industrialized countries. Plastic pollution in terrestrial systems is also linked to the use of agricultural plastics, such as polytunnels and plastic mulches. Secondary microplastics originate from varying and diverse sources, for example from mismanaged waste either accidentally or intentionally.
Effects:
As plastics have become widespread and ubiquitous in the environment, they are present in a diversity of habitats and can impact organisms at different levels of biological organization, possibly leading to population, community and ecosystem effects. Entanglement is one of the most obvious and dramatic physical impacts of macroplastics, as it often leads to acute and chronic injury or death. In particular the higher taxa (mammals, reptiles, birds and fish) are affected, and it may be critical for the success of several endangered species. Because of similar size characteristics to food, plastics are both intentionally and unintentionally ingested by a wide range of species, such as invertebrates, fish, birds and mammals. Ingestion of the non-nutritional plastics can cause damage and/or obstruction of the digestive tract and may lead to decreased foraging due to false feelings of satiation, resulting in reduced energy reserves.
Microplastics and in particular nanoparticles that are small enough to be taken up and translocated into tissues, cells and body fluids can cause cellular toxicity and pathological changes due to particle toxicity. In addition, there are also chemical risks involved as plastics can be a source of hazardous chemicals. These chemicals can be part of the plastic itself (i.e. monomers and additives) and/or chemicals that are absorbed from the environment into the matrix or such as lead, cadmium, mercury, persistent organic pollutants (POPs) like PAHs, PCBs and dioxins. However, as this process depends on the fugacity gradients, there is a lot of uncertainty about the extent that transfer of pollutants does occur in the environment. Actually, when taking all exposure pathways into account, the transfer from (micro)plastics seems to be a minor pathway.
Watch the video on the research of Inneke Hantoro.
Finally, marine plastics may act as floating habitats for invasive species, including harmful algal blooms and pathogens, leading to spreading beyond their natural dispersal range and creating the risk of disrupting ecosystems of sensitive habitats.
2.2.9. Nanomaterials
Author: Martina Vijver
Reviewers: Kees van Gestel, Frank van Belleghem, Melanie Kah
Leaning objectives:
You should be able to:
explain the differences between nanomaterials and soluble chemicals.
describe nano-specific features and explain the difference between nanomaterials and particles with a larger size, as well as how they differ from molecules.
Engineering at the nanoscale (i.e. 10-9 m) brings the promise of radical technological development. Due to their unique properties, engineered nanomaterials (ENMs) have gained interest from industry and entered the global market. Potentials ascribed to nanotechnology are amongst all: stronger materials, more efficient carriers of energy, cleaner and more compact materials that allow for small yet complex products. Currently, nanomaterials are used in numerous products, although exact numbers are lacking. In 2014, the market was estimated to contain more than 13,000 nano-based products (Garner and Keller, 2014). There is a wide variety of products containing nanomaterials, ranging from sunscreens and paint, to textiles, medicines, electronics covering many sectors (Figure 1).
Figure 1. Applications in different sectors where engineered nanomaterials are used (source: http://www.enteknomaterials.com/wp-content/uploads/2016/08/nano-malzemeler-5.jpg).
Nanomaterials:
The European Commission in 2011 adopted a new definition of ‘nanomaterial’ reading ‘a natural,incidental or manufactured material containing particles, in an unbound state or as an agglomerate or as an aggregate 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’.
Nanomaterials occur naturally, think of minuscule small fine dust, colloids in the water column, volcanic ash, carbon black and colloids known as ocean spray. In paints the features of the colloids are used to obtain the pigment colors. From the year 2000 on, an exponential growth was seen in their synthesis due to the advanced technologies and imaging techniques needed to work on a nano-scale. First generation nanotechnologies (before 2005) generally refers to nanotechnology already on the market, either as individual nanomaterials, or as nanoparticles incorporated into other materials, such as films or composites. Surface engineering has opened the doors to the development of second and third generation ENMs. Second generation nanotechnologies (2005-2010) are characterised by nanoscale elements that serve as the functional structure, such as electronics featuring individual nanowires. From 2010 onward there has been more research and development of third generation nanotechnologies, which are characterised by their multi-scale architecture (i.e. involving macro-, meso-, micro- and nano-scales together) and three-dimensionality, for applications like biosensors or drug-delivery technologies modelled on biological templates. Self-assembling bottom-up techniques have been widely developed at industrial scale, to create, manipulate and integrate nanophases into more complex nanomaterials with new or improved technological features. Post 2015, the fourth generation ENMs are anticipated to utilise ‘molecular manufacturing’: achieving multi-functionality and control of function at a molecular level. Nowadays, virtually any material can be made on the nanoscale.
Figure 2.Relationship between particle diameter and the fraction of atoms at the surface. Drawn by Wilma Ijzerman.
Size does matter
Nanoscale materials have far larger surface areas than larger objects with similar masses. A simple thought experiment shows why nanoparticles have phenomenally high surface areas.
A solid cube of a material 1 cm on a side has 6 cm2 of surface area, about equal to one side of half a stick of gum. When the 1 cm3 is filled with micrometer-sized cubes — a trillion (1012) of them, each with a surface area of 6 square micrometers — the total surface area amounts to 6 m2.. As surface area per mass of a material increases, a greater proportion of the material comes into contact with surrounding materials (Figure 2). Small particles also give that there is a high proportion of surface atoms, high surface energy, spatial confinement and reduced imperfections (Figure 2). It results in the fact that ENMs are having an enlarged reactivity. compared to larger “bulk” materials. For instance, ENMs have the potency to transfer concentrated medication across the cell membranes of targeted tissues. By engineering nanomaterials, these properties can be harnessed to make valuable new products or processes.
ENMs are often designed to accomplish a particular purpose, taking advantage of the fact that materials at the nanoscale have different properties than their larger-scale counterparts.
ENMs and environmental processes
ENMs are described as a population of particles, and quantified by the particle size distribution (PSD). Nonetheless often a single value (e.g. average ± standard deviation) is reported and not the full PSD. When the particles are suspended in an exposure medium, the size distribution of the NPs is changing over time. After being emitted into aquatic environments, NPs are subject to a series of environmental processes. These processes include dissolution and aggregation (see Figure 3) and subsequent sedimentation. It is known that the behavior and fate of NPs are highly dependent on the water chemistry. In particular, environmental parameters like pH, concentration and type of salts (especially divalent cations) and natural organic matter (NOM) can strongly influence the behaviour of NPs in the environment. For example, pH can affect the aggregation and dissolution of metallic NPs by influencing the surface potential of the NPs (von der Kammer et al., 2010). The divalent cations Ca2+ and Mg2+ are able to efficiently compress the electrical double-layer of NPs and consequently enhance homo-aggregation and hetero-aggregation of NP (see Figure 2) and the cations will be related to bridging the electrostatic interactions. In surface water, aggregation processes most often lead to sedimentation and sometimes to floating aggregates (depending on the density).
Coating of ENMs will change the dynamics of these processes. As a result of these nano-specific features, ENMs form a suspension which is different from chemicals that dissolve and form a solution. These ENMs in suspension then follow different environmental fate and behaviour compared to solutions. For this reason the way the dosage of ENPs should be expressed is highly debated within the nano-safety community. Should this be on a mass basis, as is the case for molecules of conventional chemicals (e.g. mg/L), or is the particle number the preferred dose metric such as in colloidal science (e.g. number of particles/L, or relative surface-volume ratio) or multi-mixed dosimetry expression. How to express the dose for nanomaterials is a quest still debated within the scientific community (Verschoor et al., 2019).
Figure 3.Top: Schematic illustration how metallic nanoparticles (NPs) behave in an exposure matrix (redrawn from Nowak & Bucheli, 2007). Bottom: Aggregation is divided into two type: homo-aggregation (i.e., aggregation between nanoparticles) and hetero-aggregation (i.e., aggregation of nanoparticle and biomass). Drawn by Wilma IJzerman.
Classification of NMs
Although we learned from the text above that changing the form of a nanomaterial can produce a material with new properties (i.e. a new nanomaterial); often a group of materials developed is named after the main chemical component of the ENMs (e.g. nanoTiO2) that is available in different (nano)forms. Approaches to group ENMs have been presented below:
Classification by dimensionality / shape / morphology:
Shape-based classification is related to defining nanomaterials, and has been synopsized in the ISO terminology.
Classification by composition / chemistry:
This approach groups nanomaterials based on their chemical properties.
Classification by complexity / functionality:
The nanomaterials that are in routine use in products currently are likely to be displaced by nanomaterials designed to have multiple functionalities, so called 2nd-4th generation nano-materials.
Classification by biointerface:
A proposal relates to the hypothesis that nanomaterials acquire a biological identity upon contact with biofluids and living entities. Systems biology approaches will help identify the key impacts and nanoparticle interaction networks.
References
Garner, K.L., Keller, A.A. (2014). Emerging patterns for engineered nanomaterials in the environment: a review of fate and toxicity studies. Journal of Nanoparticle Research 16: 2503.
Nowack, B., Bucheli, T.D. (2007). Occurrence, behavior and effects of nanoparticles in the environment. Environmental Pollution 150, 5-22.
Verschoor, A.J., Harper, S., Delmaar, C.J.E., Park, M.V.D.Z., Sips, A.J.A.M., Vijver, M.G., Peijnenburg, W.J.G.M. (2019). Systematic selection of a dose metric for metal-based nanoparticles. NanoImpact 13, 70-75.
Von der Kammer, F., Ottofuelling, S., Hofmann, T. (2010). Assessment of the physico-chemical behavior of titanium dioxide nanoparticles in aquatic environments using multi-dimensional parameter testing. Environmental Pollution 158, 3472-3481.
2.3. Pollutants with specific use
2.3.1. Crop Protection Products
Author: Kees van Gestel
Reviewers: Steven Droge, Peter Dohmen
Leaning objectives:
You should be able to:
describe the role of crop protection products in agriculture
mention different types of pesticides and their different target groups.
distinguish and mention important chemical groups of pesticides
- related to the chemistry
- related to the mode of action
describe major components included in a commercial formulation of a crop protection product beside the active substance(s).
Keywords: Insecticides, Herbicides, Fungicides, Active substances, Formulations
Introduction
Crop protection products are used in agriculture. The principle target of agriculture is the provision of food. For this purpose, agriculture aims to reduce the competition by other (non-crop) plants and the loss of crop due to herbivores or diseases. An important tool to achieve this is the use of chemicals, such as crop protection products (CPP). Accordingly, CPPs are intentionally introduced into the environment and represent one of the largest sources of xenobiotic chemicals in the environment. These chemicals are by definition effective against the target organism, often already at fairly low doses, but may also be toxic to non-target organisms including humans. The use of pesticides, also named Crop Protection Products (CPP) or often also Plant Protection Products (PPP, the latter term may be misleading for herbicides which are intended to reduce all but the crop plants), is therefore strictly regulated in most countries. The main pesticides used in the largest volumes world-wide are herbicide all s, insecticides, and fungicides. As shown in Table 1, pesticides are used against a large number of diseases and plagues.
Table 1. Classification of pesticides according to what they are supposed to control
Pesticide type
Target
acaricides
against mites and spiders (incl. miticides)
algicides
against algae
althelmintics (vermicides)
against parasites
antibiotics
against bacteria and viruses (incl. bactericides)
bactericides
against bacteria
fungicides
against fungi
herbicides
against weeds
insecticides
against insects
miticides
against mites
molluscicides
against slugs and snails
nematicides
against nematodes
plant growth regulators
retard or accelerate the growth of plants
repellents
drive pests (e.g. insects, birds) away
rodenticides
against rodents
Formulations
A pesticidal product usually consists of one or more active substances, that are brought onto the market in a commercial formulation (spray powder, granulate, liquid product etc.). The formulation is used to facilitate practical handling and application of the chemical, but also to enhance its effect or its safety of use. The active substance may, for instance, be a solid chemical, while application requires it to be sprayed. Or the active substance degrades fast under the influence of sunlight and therefore has to be encapsulated. One of the most used types of formulation is a concentrated emulsion, which may be sprayed directly after dilution with water. In this formulation, the active substance is dissolved in an oily matrix and a detergent is added as emulsifier to make the oil miscible with water. In this way, the active substance becomes quickly available after spraying. In so-called slow-release formulations, the active substance is encapsulated in permeable microcapsules, from which it is slowly released. Another component of a formulation can be a synergist, which increases the efficacy of the active substance, for instance by blocking enzymes that metabolize the active substance. Here an overview of main formulation constituents:
Solvents: to ease handling and application of the active substances, they are usually dissolved. For a highly water soluble compound this solvent may just be water; however, most compounds have low water solubility and they are thus dissolved in organic solvents.
Emulsifier, detergents, dispersants: used to provide a homogeneous mixture of the active substance in the aqueous spray solution.
Carrier: solid formulations, such as wettable granules (WG). All wettable powder (WP) formulations often use inert materials such as clay (kaolinite) as carrier.
Wetting agent: they help providing a homogeneous film on the plant surface.
Adjuvant: may help to increase uptake of the active substance into the plant.
Minor constituents:
Antifreeze agent: to keep the formulation stable also in cold storage conditions.
Antifoam agent: some of the formulants may result in foaming during application, which is not wanted.
Preservative, biocide: to prevent biological degradation of the active substance or the formulants during storage.
There are numerous additional specific additional constituents for specific purposes such as colours, detterents, stickers, etc
Four types of nomenclature are used in case of pesticides:
1. The trade name, e.g. Calypso®, which is given by the manufacturer. The same active substance is often sold under more than one different trade names (accordingly, the use of trade names only is not a sufficient description of the test substance in scientific literature).
2. The code name, which is the "common" name of the active substance. Calypso® 480 SC, for example, is a concentrated suspension containing 480 g/L of the active substance thiacloprid.
3. The chemical name of the active substance. Thiacloprid is [3-[(6-chloropyridin-3-yl)methyl]-1,3-thiazolidin-2-ylidene]cyanamide.
4. The name of the chemical group to which the active substance belongs, in case of thiacloprid: neonicotinoids.
Chemical classes
Pesticides represent quite a number of different groups of chemicals. Pesticides include inorganic chemicals (like copper used as a fungicide), organic synthetic chemicals, and biologicals (organic natural compounds). Pesticides from the same chemical group may be used against different pest organisms, like the organotin compounds (see below). Some chemicals have a broad mode of action: many soil disinfectants, such as metam-sodium, kill nematodes, fungi, soil insects and weeds. Other pesticides are more selective, like neonicotinoids acting only on insects, or very selective, like the insect-growth regulator fenoxycarb, which is used against leaf-rollers without affecting its natural enemies. Selectivity of a pesticide also indicates to what extent non-target species may be affected upon its application (side-effects). Integrated pest management (IPM) aims at an as sustainable as possible crop protection system by combining biological agents (predators of the pest organism) using chemicals having a selective mode of action. Such systems are nowadays receiving increasing interest in different agricultural crops.
Some groups of pesticides that were used or still are widely used are presented in more detail. Their modes of action are discussed in Chapter 4.
Chlorinated hydrocarbons
Best known representative of this group is DDT (dichloro diphenyl trichloroethane; Figure 1), which was discovered in 1939 by the Swiss entomologist Paul Hermann Müller and seemed to be an ideal pesticide: it was effective, cheap and easy to produce and remained active for a long period of time. As a remedy against Malaria and other insect borne diseases, it has saved millions of human lives. However, the high persistency of DDT, its strong bioaccumulation and its effects on bird populations have triggered the search for alternatives and its ban in most Western countries. But in some developing countries, because of a lack of suitable alternatives for an effective control of malaria, DDT is still in use to kill malaria mosquitos.
Other representatives of chlorinated hydrocarbons are lindane, also called gamma-hexachlorocyclohexane (Figure 1), and the cyclodienes that include the "drins" (aldrin, dieldrin, endrin, See Section 2.1) and endosulfan (Figure 1). Because of their high persistence and bioaccumulative potential, most organochlorinated pesticides have been banned.
Volatile halogenated hydrocarbons were often used as soil disinfectant. These compounds were injected into the soil, and acted as a nematicide but also killed fungi, soil insects and weeds. An example is 1,3-dichloropropene (Figure 1).
Figure 1.Chemical structures of four different organochlorinated pesticides widely used in the past, from left to right: DDT: 1,1'-(2,2,2-trichloroethane-1,1-diyl)bis(4-chlorobenzene), lindane: gamma-1,2,3,4,5,6-hexachlorocyclohexane, endosulfan: 6,7,8,9,10,10-hexachloro-1,5,5a,6,9,9a-hexahydro- 6,9-methano-2,4,3-benzodioxathiepine-3-oxide, and 1,3-dichloropropene. (Source: Steven Droge)
Organophosphates
Organophosphates are esters of phosphoric acid and constitute important biological molecules such as nucleic acids (DNA) or ATP. Within the contents of pesticides this refers mainly to a group of organophosphate molecules which interfere with acetylcholinesterase. Nerve gases, produced for chemical warfare (e.g., Sarin), also belong to the organophosphates. They are much less persistent and were therefore introduced as alternatives for the chlorinated hydrocarbons. The common molecular structure of organophosphates is a tri-ester of phosphate, phosphonate, phosphorthionate, phosphorthiolate, phosphordithionate or phosphoramidate (Figure 2). With two of the three ester bonds, a methyl- or ethyl- group is bound to the P atom, while the third ester bond binds the rest group or "leaving group".
Figure 2:Chemical structure of organophosphates. R = methyl of ethyl group. (Source: Steven Droge)
Dependent on the identity of the latter group, three sub-groups may be distinguished:
1. Aliphatic organophosphates, including malathion (Figure 3) and a number of systemic chemicals.
2. Phenyl-organophosphates, which are more stable than the aliphatic ones but also less soluble in water, like parathion (no longer allowed in Europe; Figure 3).
3. Heterocyclic organophosphates, including chemicals with an aromatic ring containing a nitrogen atom like chlorpyrifos (Figure 3).
Figure 3: Malathion:diethyl 2-[(dimethoxyphosphorothioyl)sulfanyl]butanedioate (left), parathion: O,O-diethyl-O-4-nitrophenyl-phosphorthioate (middle), and chlorpyrifos: O,O-diethyl O-3,5,6-trichloropyridin-2-yl phosphorothioate (right). (Source: Steven Droge)
Carbamates
Where organophosphates are derived from phosphoric acid, carbamates are derived from carbamate (Figure 4). Their mode of action is similar to that of the organophosphates. The use of older representatives of this group, like aldicarb, carbaryl, carbofuran and propoxur, is no longer allowed in Europe, but diethofencarb (Figure 4), oxamyl and methomyl are still in use.
Figure 4:Basic structure of carbamates (left) and diethofencarb: isopropyl 3,4-diethoxycarbanilate (right). (Source: Steven Droge)
Pyrethroids
A number of modern pesticides are derived from natural products. Pyrethroids are based on pyrethrum, a natural insecticide from flowers of the Persian ox-eyed daisy, Chrysanthemum roseum. Typical for the molecular structure of pyrethroids is the cyclopropane-carboxyl group (the triangular structure), which is connected with an aromatic group through an ester bond (Figure 5). Pyrethrum is rapidly degraded under the influence of sunlight. Synthetic pyrethroids, which are much more stable and therefore used on a large scale against many different insects, include cypermethrin (Figure 5), deltamethrin, lambda-cyhalothrin, fluvalinate and esfenvalerate.
Figure 5: Cypermethrin:[cyano-(3-phenoxyphenyl)methyl]3-(2,2-dichloroethenyl)-2,2-dimethylcyclopropane-1-carboxylate. (Source: Steven Droge)
Neonicotinoids
Based on the natural compound nicotine, which acts as a natural insecticide against plant herbivores, but which was banned as an insecticide due to its high human toxicity, in the 1980s a new group of more specific insecticides has been developed, the neonicotinoids (Figure 6). Several neonicotinoids (e.g., imidacloprid, thiamethoxam) are systemic. This means that they are taken up by the plant and exert their effect from inside the plant, either on the pest organism (systemic fungicides or insecticides) or on the plant itself (systemic herbicides). The systemic neonicotinoids are widely applied as seed dressing in major crops like maize and sunflower. Other compounds are mainly used in spray applications, e.g. in fruit growing (thiacloprid, acetamiprid, etc.). Although neonicotinoids are more selective and therefore preferred over the older classes of insecticides like organophosphates, carbamates and pyrethroids, in recent years they have become under debate because of their side effects on honey bees and other pollinators.
Figure 6: Nicotine:(S)-3-[1-methylpyrrolidin-2-yl]pyridine (left) and the neonicotinoid insecticides imidacloprid: N-{1-[(6-chloro-3-pyridyl)methyl]-4,5-dihydroimidazol-2-yl}nitramide (middle) and thiacloprid: {(2Z)-3-[(6-chloropyridin-3-yl)methyl]-1,3-thiazolidin-2-ylidene}cyanamide (right). (Source: Steven Droge)
Isothiocyanates
Isothiocyanates were used on a large scale as soil disinfectant against nematodes, fungi and weeds. The large number of chemicals with different chemical origin belonging to the isothiocyanates have in common that they form isothiocyanate in soil. A representative of this group is metam-sodium (Figure 7).
Figure 7.Metam-sodium:sodium methylaminomethanedithioate forming methyl isothiocyanate. (Source: Steven Droge)
Organotin compounds
Fentin hydroxide (Figure 8) was used as a fungicide against Phytophthora (causing potato -disease). Tributyltin compounds (TBT) were used as anti-fouling agent (algicide) on ships. TBTC (tributyltin chloride) is extremely toxic to shell-fish, such as oysters, and for this reason banned in many countries. Fenbutatin-oxide was used as an acaricide against spider mites on fruit trees (tributyltin chloride).
Figure 8: Fentin hydroxide:triphenyltin hydroxide. (Source: Steven Droge)
Ryanoids
Also indicated as diamide insecticides, this group includes chemically distinct synthetic compounds such as chlorantraniliprole (Figure 9), flubendiamide, and cyantraniliprole, that act on the ryanodine receptor and are used against chewing and sucking insects.
Figure 9:Chlorantraniliprole:5-bromo-N-[4-chloro-2-methyl-6-(methylcarbamoyl)phenyl]-2-(3-chloropyridin-2-yl)pyrazole-3-carboxamide. (Source: Steven Droge)
Phenoxy acetic acids
Phenoxy acetic acids are systemic herbicides, exerting their action after uptake by the leaf and translocation throughout the plant. Especially plants with broad, horizontally oriented leaves are sensitive for these herbicides. 2,4-D (Figure 10) is the best known representative of this group.
Figure 10.2,4-D:the anionic form of 2,4-dichloro phenoxy acetic acid (pKa 2.73). (Source: Steven Droge)
Triazines
Triazines are heterocyclic nitrogen compounds, whose structure is characterized by an aromatic ring in which three carbon atoms have been replaced by nitrogen atoms. Triazines are usually applied to the soil before seed germination. The use of several compounds (atrazine, simazine) has been banned in Europe, while others like metribuzin and terbuthylazine (Figure 11) are still in use.
Figure 11.Terbuthylazine:N-tert-butyl-6-chloro-N'-ethyl-1,3,5-triazine-2,4-diamine (left), common replacement of the EU-banned herbicide atrazine (right). (Source: Steven Droge)
Bipiridyls
This group contains the herbicides diquat and paraquat (Figure 12) which mainly act as contact herbicides. This means they damage the plant without being taken up. In soil, they are rapidly inactivated by strong binding to soil particles. The use of paraquat is no longer allowed in Europe, but diquat is still in use.
Figure 12.Paraquat:1,1′-Dimethyl-4,4′-bipyridinium dichloride (left), and diquat: 1,1'-Ethylene-2,2'-bipyridylium dibromide (right). (Source: Steven Droge)
Glyphosate and Glufosinate
As an alternative to the above mentioned herbicides, glyphosate and later glufosinate were developed. These are systemic broad-spectrum herbicides with a relatively simple chemical structure (Figure 13). Their low toxicity to other organisms triggered pesticide producers to introduce genetically modified crops (e.g. soybean, maize, oilseed rape, and cotton) that contain incorporated genes for resistance against these broad-spectrum herbicides. This type of resistance allows the farmer to use the herbicide without damaging the crop. For this reason, environmentalist fear an unrestricted use of these herbicides, which indeed is the case especially for glyphosate (better known under the formulation name Roundup®).
Figure 13. Glyphosate:N-(phosphonomethyl)glycine in the two species most relevant for natural pH range (left), and glufosinate: (RS)-2-Amino-4-(hydroxy(methyl)phosphonoyl)butanoic acid in the most relevant species for natural pH range (right). (Source: Steven Droge)
Triazoles
Several modern fungicides are sharing a triazole group (Figure 14). These fungicides have gained importance because of problems with the resistance of fungi against other classes of fungicides. Members of this group for instance are epoxiconazole, propiconazole and tebuconazole.
Figure 14: Triazole:1H-1,2,3-Triazole (left), and epoxiconazole: (2RS,3SR)-1-[3-(2-chlorophenyl)-2,3-epoxy-2-(4-fluorophenyl)propyl]-1H-1,2,4-triazole (right). (Source: Steven Droge)
Biological pesticides
Biological pesticides are produced in living organisms as secondary metabolites to protect themselves against predators, herbivores, parasites or competition. They can be highly effective and act at low concentrations (high toxicity), but in contrast to some synthetic pesticides they are usually sufficiently biodegradable. Compounds like pyrethrum or strobilurin are produced within the plant or within the fungus and are thus protected against photolysis or other environmental degradation. Furthermore, the living organism can produce additional quantities of the secondary metabolite on demand. When used as a pesticide applied as a spray, however, the molecule needs to be modified to enhance its stability (for example against photolysis) to remain sufficiently active over a sufficient period of time. Accordingly, synthetic derivatives of these biological molecules are often more stable, less biodegradable. Examples are the Bt insecticide, which contains an endotoxin highly toxic to insects produced by the bacterium Bacillus thuringiensis, and avermectins, complex molecules synthesized by the bacterium Streptomyces avermitilis. Avermectins act as insecticides, acaricides and have anthelminthic properties. In nature, eight different forms of avermectin have been found. Ivermectin is a slightly modified structure that is synthesized and marketed commercially. Other compounds belonging to this group are milbemectin and emamectin.
Genetically modified plants containing a gene coding for the toxin produced by the bacterium Bacillus thuringiensis (or Bt) are another example of genetic modification being applied in agriculture produce insect-resistant crops.
European legislation describes a biocide as ‘chemical substance or microorganism intended to destroy, deter, render harmless, or exert a controlling effect on any harmful organism by chemical or biological means’. The US Environmental Protection Agency (EPA), an independent agency of the U.S. federal government to protect the environment, defines biocides as ‘a diverse group of poisonous substances including preservatives, insecticides, disinfectants and pesticides used for the control of organisms that are harmful to human or animal health or that cause damage to natural or manufactured products’. The definition by the EPA includes pesticides (Chapter 2.3.1). In the scientific and non-scientific literature, the distinction between biocides, pesticides and plant protection products is often vague.
Biocides are used all around us:
The toothpaste that you used this morning contains biocides to preserve the toothpaste
The water that you used to flush you mouth is prepared with biocides for disinfection
The clothes that you are wearing are impregnated with biocides to prevent smells
The food that you ate for breakfast might have contained biocides to preserve the food
The construction materials around you have surface coatings that contain biocides to prevent biological degradation of the material
A biocide contains an ‘active substance’, which is the chemical that is toxic to its target organism, and often contain ‘non-active co-substances’, which could help in reaching desired product parameters, such as a viscosity, pH, colour, odour or increase its handling or effectiveness. The combination of active substances and non-active substances together makes up the ‘biocidal product’. An example of a well-known biocidal product is TriChlor, which contains active substance chlorine that is used to disinfect swimming pools. Because it is impractical to store chlorine gas for the treatment of swimming pools, TriChlor tablets are added to the pool water. TriChlor is trichloroisocyanuric acid (Figure 1). When dissolved in water, the Cl atoms are replaced by H atoms, forming chlorine (Cl-) and cyanuric acid (Figure 2). The free chlorine is able to disinfect the swimming pool.
A biocidal product can also contain multiple biologically active substances to enhance its effectivity, such as AQUCAR™ 742 produced by DuPont. It contains glutaraldehyde (Figure 3) and quaternary ammonium compounds (Figure 4) that have a synergistic toxic effect on microorganisms that are present in oilfields and could form biofilms in the pipelines.
The biocidal products are classified into 22 different product-types by the European Chemicals Agency (ECHA) (Table 1). It is possible that an active substance can be classified in more than one product types.
Table 1. The classification of biocides in 22 product types (www.echa.europe.eu)
Main group 1: Disinfectants and general biocidal products
Product type 1 – Human hygiene biocidal products
Product type 2 – Private area and public health area disinfectants and other biocidal products
Product-type 18 – Insecticides, acaricides and products to control other arthropods
Product-type 19 – Repellents and attractants
Product-type 20 – Control of other vertebrates
Main group 4: Other biocidal products
Product-type 21 – Antifouling products
Product-type 22 – Embalming and taxidermist fluids
Legislation
In Europe, biocides are authorised for production and use by the Biocidal Products Regulation (BPR, Regulation (EU) 528/2012) of the ECHA. The BPR ‘aims to improve the functioning of the biocidal product market in the EU, while ensuring a high level of protection for humans and the environment.’ (https://echa.europe.eu/legislation). This is an alternative regulatory framework than that for the plant protection products, managed by the European Food Safety Authority (EFSA). All biocidal products go through an extensive authorisation process before they are allowed on the market. The assessment of a new active starts with the evaluation of a product by the authorities of an ECHA member state, after which the ECHA Biocidal Products Committee forms an opinion. The European Commission then makes a decision to approve or reject the new active substance based on the opinion of ECHA. This approval is granted for a maximum of 10 years and needs to be renewed after it reaches the end of the registration period. The BPR has strict criteria for new active substances, and meeting the following ‘exclusion criteria’ will result in the new active substance not being approved:
Carcinogens, mutagens and reprotoxic substances categories 1A or 1B according to CLP regulation
Endocrine disruptors
Persistent, bioaccumulative and toxic (PBT) substances
Very persistent and very bioaccumulative (vPvB) substances
In very special cases, new active substances will be allowed on the market when meeting this exclusion criteria, if they are important for public health and public interest and there are no alternatives available. To lower the pressure on public health and the environment, there is also a candidate list for active substances to be substituted for less harmful active substances when the old active substances meet the following criteria:
It meets one of the exclusion criteria
It is classified as a respiratory sensitizer
Its toxicological reference values are significantly lower than those of the majority of approved active substances for the same product-type and use
It meets two of the criteria to be considered as PBT
It causes concern for human or animal health and for the environment even with very restrictive risk management measures
It contains a significant proportion of non-active isomers or impurities
The impact of environmental release
The release of biocides in the environment can have huge consequences, since these products are designed to cause damage to living organisms. A classic example is the release of tributyltin from shipyards, harbours and on sailing routes from the antifouling paint on the hulls of ships (De Mora, 1996). Tributyltin was used in the antifouling paint from the 1950s on to prevent microorganisms from settling on the hulls of ships, which would increase the fuel costs and repair costs. However, the release of tributyltin from the paint resulted in a toxic effect on organisms at the bottom of the food chain, such as algae and invertebrates. Tributyltin then biomagnified in the food web, this way affecting larger predators, such as dolphins and sea otters. Eventually, tributyltin entered the diet of humans. The first legislation on the use of tributyltin for ships dates back to the 1980s, but it was not until the Rotterdam Convention of 2008 that the complete use of tributyltin as an active biocide in antifouling paints was banned. Biocides can also have an effect on the capability of the environment to deal with pollution. Microorganisms are responsible for cleaning polluted areas by using the pollutant as food-source. McLaughlin et al. (2016) studied the effect of the release of biocide glutaraldehyde in spilled water from hydraulic fracturing on the microbial activity and found that the microbial activity was hampered by the biocide glutaraldehyde. Hence, because of the biocide, the environment was not or slower capable to return to its original state.
References
De Mora, S.J. (1996). Tributyltin: case study of an environmental contaminant, Vol. 8, Cambridge Univ. Press
McLaughlin, M.C., Borch, T., Blotevogel, J. (2016). Spills of hydraulic fracturing chemicals on agricultural topsoil: Biodegradation, sorption and co-contaminant interactions, Environmental Science & Technology 50, 6071-6078
2.3.3. Pharmaceuticals and Veterinary Pharmaceuticals
Author: Thomas ter Laak
Reviewers: John Parsons, Steven Droge, Stefan Kools
Leaning objectives:
You should be able to:
understand what pharmaceuticals are and how pharmaceuticals can enter the environment
understand how emissions and environmental concentrations of pharmaceuticals can be estimated / modeled
Keywords: emission, waste water treatment, disease treatment, mass balance modelling, human pharmaceuticals, veterinary pharmaceuticals
Introduction
Pharmaceuticals are used by humans (human pharmaceuticals) and administered to animals (veterinary pharmaceuticals).
The active ingredients used in human and veterinary medicine partially overlap, however, the major fraction of pharmaceutically active substances in use are restricted to human consumption. Next to that, some active ingredients are used in other applications as well, such as biocides or in plant protection products. In veterinary practice most of the applied pharmaceuticals are antibiotics and anti-parasitic agents, while in human medicine, pharmaceuticals to treat e.g. diabetes, pain, cardiovascular diseases, autoimmune disorders and neurological disorders make up a much larger portion of the pharmaceuticals in use. Worldwide pharmaceutical consumption has increased over the last century (several numbers are summarized here). It is expected that the consumption will further increase due to a wider access to pharmaceuticals in developing countries. Additionally, demographic trends such as aging populations often seen in developed countries can also lead to increased consumption of pharmaceuticals, since older generations generally consume more pharmaceuticals than younger ones (van der Aa et al. 2010). The widespread and increasing use and their biological activity makes them relevant for environmental research. Pharmaceuticals are specifically designed and used for their biological effect in humans or treated animals. For that reason, we know a lot about their potential environmental effects as well as on their application and emission. Below an overview is given on the emission, occurrence and fate(modeling) of pharmaceuticals in the environment.
Pharmaceuticals in the environment
Pharmaceuticals can enter the environment through various routes. Figure 1 gives an overview of the major emission routes of pharmaceuticals to the environment.
Figure 1.Pharmaceutical emissions routes to the (aqueous) environment. STP = sewage treatment plant (adapted from Schmitt et al., 2017).
Pharmaceutical are produced, transported to users (humans and/or animals), used by humans or animals. After use, the active ingredients are partially metabolized and both parent compounds and metabolites can be excreted by the users via urine and feces. For humans, the major routes are transport to wastewater treatment plants, septic tanks or directly emission to soil or surface water. For animals, and especially livestock, manure contains a major fraction of the pharmaceuticals that are excreted. These pharmaceuticals end up in the environment when animals are grazing outside or when centrally collected manure is applied as fertilizer on arable land. The treatment and further application of communal wastewater and manure varies between countries and regions. Subsequently, emissions can also vary leading to different compositions and concentrations of pharmaceuticals and metabolites in the environment. In Figure 2 concentration ranges of pharmaceuticals and some of their transformation products in the Meuse river and some tributaries are shown.
Figure 2.Pharmaceutical concentrations of pharmaceuticals in the River Meuse and some of its tributaries (adapted from Ter Laak et al. 2014). Parent pharmaceuticals are plotted in bleu, transformation products in red. Drawn by Wilma IJzerman.
Properties of pharmaceuticals and their behavior and fate in the environment
Pharmaceuticals in use are developed for a wide array of diseases and therapeutic treatments. The chemical structures of these substances are therefore also very diverse, considering their size, structural presence of specific atoms, and physicochemical properties such as their hydrophobicity, aqueous solubility and ionization under environmentally relevant pH values, as shown for some examples in Figure 3.
Figure 3.Examples of pharmaceuticals, illustrating the variable chemical structures. (Source: Steven Droge)
As a consequence of their structural diversity, the environmental distribution and fate of pharmaceuticals is also very variable. Nevertheless, pharmaceuticals have generally certain properties in common:
Pharmaceuticals are designed to have a specific biological activity that determines their pharmacological application.
Most pharmaceuticals are rather robust against metabolism of their users (humans or animals) in order to reach stable therapeutic levels inside the user.
Pharmaceuticals are often relatively soluble in water, since their therapeutic application requires absorption and distribution to reach specific targets sites in living organisms. Aqueous solutions such as blood are often the internal transport medium. Less soluble pharmaceuticals are metabolized to allow renal excretion, leading to soluble metabolites.
These three generic properties also make them of environmental relevance since:
Pharmaceuticals are likely biologically active in non-target organisms, thereby disturbing their behavior, metabolism or other functions.
Pharmaceuticals are rather persistent in water treatment or the environment
Pharmaceuticals and/or metabolites have rather high aqueous solubility which makes them mobile and as a consequence they may end up in (ground)water.
Continuous use leads to continuous emission and subsequently continuous presence in environmental waters, this is called ‘pseudo persistence’.
Occurrence and modelling of human and veterinary pharmaceuticals in the environment
Pharmaceuticals in the environment have been studied since the 1990s. Most studies have been performed in surface waters, but wastewater (effluents), groundwater, drinking water, manure, soil and sediments were also studied. Pharmaceuticals have been observed in all these matrices in concentrations generally varying from µg/L to sub ng/L levels (Aus der Beek et al., 2016, Monteiro and Boxall, 2010). Various studies have related environmental loads and related concentrations to human consumption data. Basically such mass balance or studies that relate consumption in catchments of streams, lakes or rivers to environmental concentrations work according the following principle:
Modelling pharmaceuticals in the environment
The consumption of human pharmaceuticals, is relatively well documented and data are (publicly) available. Hence, based on consumption using several assumptions, environmental concentrations of pharmaceuticals can be related to consumption. This prediction works best for the most persistent pharmaceuticals, as these pharmaceuticals are hardly affected by transformation processes that can be variable as a result of environmental conditions. When loss factors become larger, they generally also become more variable, through seasonal variations in use as well as variation in loss during wastewater treatment and loss processes in the receiving rivers. This makes the loads and concentrations of more degradable pharmaceuticals more difficult to predict (Ter Laak et al., 2010).
Loads in a particular riverine system (such as a tributary of the river Meuse in the example below) can be predicted with a very simplified model. Here the pharmaceutical consumption over a selected period is multiplied by the fraction of the selected pharmaceuticals that is excreted unchanged by the human body (ranging from 0 to 1) and the fraction that is able to pass the wastewater treatment plant (WWTP) (ranging from 0 to 1):
When this is related to actual measured concentrations and loads calculated from these numbers, the correlation between predicted and measured loads can be plotted. Various studies have shown that environmental loads can be predicted within a factor of 3 for most commonly observed pharmaceuticals (see e.g., Ter Laak et al., 2010, 2014).
Figure 4.Measured versus predicted loads in a tributary of the Meuse river (adapted from Ter Laak et al., 2014)
For veterinary pharmaceuticals this so called ‘immision-emission balancing’ is more difficult for a number of reasons (see e.g., Boxall et al., 2003):
First, veterinary pharmaceuticals are applied in different quantities and using different application routes in different (live-stock) animals.
Second, animal excrements can be burned, stored for later use as fertilizes or directly emitted when animals are kept outside, leading to different emissions per route.
Finally, the fate of pharmaceuticals associated with animal excrements to soil, groundwater and surface water is variable and poorly understood.
In a way the emissions and fate of veterinary pharmaceuticals is similar to emissions of pesticides used in agriculture. However, the understanding on loads entering the system and the fate related to the various emission routes and emissions in combination with a complex matrix (urine, feces manure) is more limited (Guo et al., 2016). As a consequence, environmental fate studies of veterinary pharmaceuticals often describe specific cases, or cover laboratory studies to unravel specific aspects of the environmental fate of these pharmaceuticals (Kaczala and Blum, 2016, Kümmerer, 2009).
Concluding remarks
Pharmaceuticals are commonly found in the environmental compartments such as surface water, soil, sediment and groundwater (Williams et al., 2016). Pharmaceuticals consist of a single or multiple active ingredients that have a specific biological activity. The therapeutic application and pharmacological mechanisms provide valuable information to evaluate the environmental hazard of these chemicals. Their physicochemical properties are of more relevance for the assessment of the environmental fate and exposure. The occurrence in the environment and the biological activity of this group of contaminants makes them relevant in environmental science.
References
Aus der Beek, T., Weber, F., Bergmann, A., Hickmann, S., Ebert, I., Hein, A., Küster, A. (2016). Pharmaceuticals in the environment-Global occurrences and perspectives. Environmental Toxicology and Chemistry 35, 823-835.
Boxall, A.B.A., Kolpin, D.W., Halling-Soerensen, B., Tolls, J. (2003). Are veterinary medicines causing environmental risks? Environmental Science and Technology 37, 286A-293A.
Guo, X.Y., Hao, L.J., Qiu, P.Z., Chen, R., Xu, J., Kong, X.J., Shan, Z.J., Wang, N. (2016). Pollution characteristics of 23 veterinary antibiotics in livestock manure and manure-amended soils in Jiangsu province, China. Journal of Environmental Science and Health Part B: Pesticides, Food Contaminants, and Agricultural Wastes 51, 383-392.
Kaczala, F., Blum, S.E. (2016). The occurrence of veterinary pharmaceuticals in the environment: A review. Current Analytical Chemistry 12, 169-182.
Kümmerer, K. (2009). The presence of pharmaceuticals in the environment due to human use - present knowledge and future challenges. Journal of Environmental Management 90, 2354-2366.
Monteiro, S.C., Boxall, A.B.A. (2010). Occurrence and fate of human pharmaceuticals in the environment. Reviews of Environmental Contamination and Toxicology 202, 53-154.
Schmitt, H., Duis, K., ter Laak, T.L. (2017). Development and dissemination of antibiotic resistance in the environment under environmentally relevant concentrations of antibiotics and its risk assessment - a literature study. (UBA-FB) 002408/ENG; Umweltbundesamt: Dessau-Roßlau, January 2017; p 159.
Ter Laak, T.L., Kooij, P.J.F., Tolkamp, H., Hofman, J. (2014). Different compositions of pharmaceuticals in Dutch and Belgian rivers explained by consumption patterns and treatment efficiency. Environmental Science and Pollution Research 21, 12843-12855.
Ter Laak, T.L., Van der Aa, M., Houtman, C.J., Stoks, P.G., Van Wezel, A.P. (2010). Relating environmental concentrations of pharmaceuticals to consumption: A mass balance approach for the river Rhine. Environment International 36, 403-409.
Van der Aa, N.G.F.M., Kommer, G.J., van Montfoort, J.E., Versteegh, J.F.M. (2011). Demographic projections of future pharmaceutical consumption in the Netherlands. Water Science and Technology 63, 825-832.
Williams, M., Backhaus, T., Bowe, C., Choi, K., Connors, K., Hickmann, S., Hunter, W., Kookana, R., Marfil-Vega, R., Verslycke, T. (2016). Pharmaceuticals in the environment: An introduction to the ET&C special issue. Environmental Toxicology and Chemistry 35, 763-766.
2.3.4. Drugs of abuse
Author: Pim de Voogt
Reviewer: John Parsons, Félix Hernández
Leaning objectives:
You should be able to:
distinguish between licit and illicit drugs
know what sources cause illicit drugs to show up in the environment
Since about little more than a decade, drugs of abuse (DOA) and their degradation products have been recognized as emerging environmental contaminants. They are among the growing number of chemicals that can be observed in the aquatic environment.
Figure 1.Chemical structures of most popular drugs of abuse, in their predominant speciation under physiological conditions (pH7.4). (Source: Steven Droge)
The residues of a major part of the chemicals used in households and daily life end up in our sewer systems. Among the many chemicals are cleaning agents and detergents, cosmetics, food additives and contaminants, pesticides, pharmaceuticals, and surely also illicit drugs. Once in the sewer, they are transported to wastewater treatment plants (WWTPs), where they may be removed by degradation or adsorption to sludge, or end up in the effluent of the plant when removal is incomplete.
The consumption of both pharmaceuticals and DOA has increased substantially over the last couple of decades as a result of several factors, including ageing of the population, medicalization of society and societal changes in life-style. As a result the loads in wastewater of drugs and their transformation products formed in the body after consumption have steadily increased. More recently, it has been observed that chemical waste from production sites of illicit drugs is being occasionally discharged into sewer systems, thereby dramatically increasing the loads of illicit drug synthesis chemicals and end products transported to WWTPs. As WWTPs are not designed to remove drugs, a substantial fraction of the loads may end up in receiving waters and thus pose a threat to both human and ecosystem health.
Drugs of Abuse (DOA)
Europe’s most commonly used illicit drugs are THC (cannabis), cocaine, MDMA (ecstasy) and amphetamines. The structure of these drugs is given in Figure 1. Other important DOA include the opioids such as heroine and fentanyl, GHB, Khat and LSD.
Drugs of abuse are controlled by legislation, in The Netherlands by the Opium Act. The Opium Act encompasses two lists of substances. List one chemicals are called hard drugs while List II chemicals are known as soft drugs. Some narcotics are also being used for medicinal purposes, e.g., ketamine, diazepines, and one of the isomers of amphetamine. New psychoactive substances (NPS), also known as designer drugs or legal highs (because they are not yet controlled as they are not listed on the Opium Act lists), are synthesized every year and become available on the market in high numbers (see Figure 2).
Figure 2.Number and categories of new psychoactive substances notified to the EU Early Warning System for the first time, 2005–2017. Redrawn from EMCDDA, European Drug Report 2018, Lisbon, by Wilma Ijzerman.
Wastewater-based epidemiology
Central sewage systems collect and pool wastewater from household cleaning and personal care activities as well as excretion products resulting from human consumption and thus contain chemical information on the type and amount of substances used by the population connected to the sewer. Drugs that are consumed are metabolized in the body and subsequently excreted. Excretion products can include the intact compounds as well as the transformation products, that can be used as biomarkers. An example of the latter is benzoylecgonine, which is the major transformation product of consumed cocaine. The collective wastewater from the sewer system carrying the load of chemicals is directed to the WWTP, and this wastewater influent can be sampled at the point where it enters the WWTP. By appropriate sampling of the influent during discrete time-intervals, e.g. 24 h, a so-called composite sample can be obtained and the concentrations of the chemicals can be determined. The volume of influent entering the WWTP is recorded continuously. Multiplying the observed 24 h average concentration of a compound with the total 24 h volume yields the daily load of the chemical entering the WWTP. This load can be normalised to the number of people living in the sewer catchment, resulting in a load per inhabitant. The loads of drugs in wastewater influents are usually expressed as mg.day-1.1000 inh-1. Normalised drug load data allow comparison between sewer catchments, such as shown in Figure 3. Obtaining chemical information about the population through wastewater analysis is known as Wastewater-based epidemiology, WBE (Watch thevideo). While WBE was developed originally to obtain data on consumption of DOA, the methodology has been shown to have a much wider potential: in calculating the consumption of e.g., alcohol, nicotine, NPS, pharmaceuticals and doping, as well as for assessing community health indicators, such as incidence of diseases or stress biomarkers.
Figure 3.Consumption of cocaine in 19 European cities in 2011 calculated from chemical analysis of influent loads of benzoylecgonine, a urinary biomarker of human cocaine consumption. Redrawn from Thomas et al. (2011) by Wilma Ijzerman.
DOA and the environment
Barring direct discharges into surface waters or terrestrial environments, the major sources of DOA to the environment are WWTP effluents. Conventional treatment in municipal WWTPs has not been specifically designed to remove pharmaceuticals or DOA. Removal rates of DOA vary widely and depend on compound properties such as persistence and polarity as well as WWTP operational conditions and process configurations. Some DOA cross WWTPs almost unhindered, thus ending op in the receiving waters. Examples of the latter are MDMA and some diazepines (see Figure 4). Despite that several studies report the presence of DOA or their transformation products in surface waters, until now there is very little information about their aquatic ecotoxicity available in the scientific literature.
Figure 4.Data demonstrating that WWTP emit DOA to receiving waters. A) Removal efficiencies of DOA recorded in five Dutch WWTPs; B) Estimated discharges (g/day) of DOA from WWTPs based on monitoring data and WWTP effluent flow rates in 2009 (Sources: Bijlsma et al, 2012; Van der Aa et al, 2013). Drawn by Wilma IJzerman.
Recently, chemical waste from synthetic DOA manufacturing including their precursors and synthesis byproducts have been observed to be discharged directly into sewers. In addition, containers with chemical waste from DOA production sites have been dumped on soil or surface waters. Apart from solvents and acids or bases this waste often contains remainders of the synthesis products, which can then be dissipated in the aquatic environment or seep through the soil into groundwater.
Considering that DOA are highly active in the human body, it can be expected that some of them, in particular the more persistent ones, may exert some effects on aquatic biota when their levels increase in the aquatic environment.
References
Van der Aa, M., Bijlsma, L., Emke, E., et al. (2013). Risk assessment for drugs of abuse in the Dutch watercycle. Water Research 47(5), 1848-1857.
Bijlsma, L., Emke, E., Hernández, F., de Voogt, P. (2012). Investigation of drugs of abuse and relevant metabolites in Dutch sewage water by liquid chromatography coupled to high resolution mass spectrometry. Chemosphere 89(11), 1399-1406.
Thomas, K. V., Bijlsma, L., Castiglioni, S., et al. (2012). Comparing illicit drug use in 19 European cities through sewage analysis. Science of the Total Environment 432, 432-439.
2.3.5. Hydrocarbons
Author: Pim N.H. Wassenaar
Reviewer: Emiel Rorije, Eric M.J. Verbruggen, Jonathan Martin
Learning objectives:
You should be able to
explain the diversity/variation in hydrocarbon structures.
explain the specific and non-specific toxicological effects of several hydrocarbons.
Hydrocarbons are a class of chemicals that only consist of carbon and hydrogen atoms. But despite their simplicity in building blocks, this group of chemicals consists of a wide variety of structures, as there are differences in chain length, branching, bonding types and ring structures. The main sources of hydrocarbons are crude oil and coal, which are formed over millions of years by natural decomposition of the remains of plants, animals or wood, and are used to derive products we are using on a daily basis, including fuels and plastics Other natural sources include natural burning (forest fires) and volcanic sources
Hydrocarbon classification
The major classes of hydrocarbons are paraffins (i.e. alkanes), naphthenics (i.e. cycloalkanes) and aromatics (Figure 1), and within these classes, several subclasses can be identified. Paraffins are hydrocarbons that do not contain any ring structures. Paraffins can be subdivided in normal (n-) paraffins, which do not contain any branching (straight chain), and iso-paraffins (i-), which do contain a branched carbon-chain. When alkanes include at least one carbon-carbon double bond, they are considered olefins (or alkenes).
Naphthenic and aromatic hydrocarbons both contain ring-structures but differ in the presence of aromatic or non-aromatic rings. The naphthenics and aromatics can be further specified based on their ring count; often mono-, di- and poly-ring structures are distinguished from each other. Of all these classes, the polycyclic aromatic hydrocarbons (PAHs) are the best-studied category in terms of all kinds of environmental aspects.
Figure 1.Chemical structures of common hydrocarbon classes. (by author)
Besides the classes considered in Figure 1., combinations of these classes also exist. Naphthenic or aromatic structures with an alkane side chain are mostly still considered as naphthenic or aromatic hydrocarbons, respectively. However, when a non-aromatic-ring is fused with an aromatic-ring, the hydrocarbon is classified as a naphthenic-aromatic structure. Depending on the ring-count several subclasses can be identified, including naphthenic-mono-aromatics and naphthenic-poly-aromatics.
Concerns for human health and the environment
Because of their lack of polar functional groups, hydrocarbons are generally hydrophobic and, as a consequence, many are able to cause acute toxic effects in aquatic animals by a non-specific mode of action known as narcosis (or baseline toxicity). Narcosis is a reversible state of inhibited activity of membrane structures within the cells of organisms. Narcosis type toxicity is considered the minimum toxicity that any substance will be able to have, just by reaching concentration levels in the phospholipid bilayer of the cell membranes that disturb membrane transportation process. Hence the name “baseline” or minimum toxicity. When these events take place above a certain threshold, systemic toxicity can be observed in the organism, such as lethality. This threshold concentration is also known as the critical body residue (CBR) (Bradbury et al., 1989; Parkerton et al., 2000; Veith & Broderius, 1990).
Nevertheless, hydrocarbons can also have a more specific mechanisms of action, resulting in greater toxicity than baseline toxicity. For example, the toxicity of several PAHs increases in combination with ultraviolet radiation due to photo-induced toxicity. Photo-induced toxicity may be caused by photoactivation, in which a PAH is degraded into an oxidized product with a higher toxicity, or rather by photosensitization, in which reactive oxygen species (ROS) are formed due to an excited state of the PAHs (Figure 2) (Roberts et al., 2017). PAHs are especially vulnerable to photodegradation as their absorption spectrum falls within the range of wavelengths reaching the earth’s surface (> 290 nm), which is not the case for most monoaromatic and aliphatic hydrocarbons (EMBSI, 2015). The photo-induced effects are of particular concern for aquatic species with transparent bodies, like zooplankton and early life stages, as more UV-light can penetrate into their organs and tissues (Roberts et al., 2017).
Figure 2.Mechanism of photo-induced toxicity of the polycyclic aromatic hydrocarbon anthracene via photosensitization or photomodification reactions, respectively. Adapted from Roberts et al. (2017) by Steven Droge.
Several hydrocarbons are also able to cause genotoxicity and cancer upon exposure, including benzene, 1,3-butadiene and some PAHs. The carcinogenicity of PAHs is caused by biotransformation into reactive metabolites, specifically into epoxides which are the first step in oxidation of aromatic ring structures into dihydrodiol ring systems (Figure 3). In general, the biotransformation step increases the water solubility of the hydrocarbons (Phase I metabolism) and promotes subsequent conjugation and excretion (Phase II metabolism). However, several epoxide metabolites – more specifically the most stable aromatic epoxides - can reach the cell nucleus and covalently react with DNA, forming DNA adducts, and induce mutations (Figure 3). Ultimately, if not repaired such mutations can accumulate and may result in the formation of tumors (Ewa & Danuta, 2016). Specifically, PAHs with a bay-like region are of concern as biotransformation results in relatively stable reactive epoxides that are not accessible to epoxide hydrolase enzymes (Figure 3) (Jerina et al. 1980). Similar to PAHs, 1,3-butadiene and benzene are also able to cause cancer via the effects of their respective reactive metabolites (Kirman et al., 2010; US-EPA 1998).
Figure 3.The biotransformation pathways of benzo(a)pyrene and binding to the DNA of reactive intermediates. Adapted from Homburger et al. (1983) by Steven Droge.
Besides their toxicity, some hydrocarbons such as the high molecular weight PAHs can be persistent in the environment and may accumulate in biota as a result of their hydrophobicity. It is therefore expected that internal concentrations are higher for such hydrocarbons and it is interesting that there is thus a relationship between narcosis and bioaccumulation potential. Consequently, these hydrocarbons might be of even greater concern.
Characterization of mixtures of hydrocarbons
As most research focused on specific hydrocarbons, including several PAHs, it is important to note that the biodegradation, bioaccumulation and toxicity potential of many hydrocarbons is still not fully known, such as for alkylated PAHs and naphthenics. As there is such a wide variety in hydrocarbon structures, it is impossible to assess the (potential) hazards of all hydrocarbons separately. Therefore, grouping approaches have been developed to speed up the risk assessment. Within a grouping approach, hydrocarbons can be clustered based on structural similarities. The underlying assumption is that all chemicals in a group are expected to have fairly similar physicochemical properties, and subsequently also fairly similar environmental fate and effect properties. As a result, such a group could potentially be assessed as if it is one single hydrocarbon.
The applicability of a hydrocarbon specific grouping approach, known as the Hydrocarbon Block Method (King et al., 1996), to assess the biodegradation and bioaccumulation potential of hydrocarbons is currently being investigated. Within this approach, all hydrocarbons are grouped based on their functional class (e.g. paraffin, naphthenic, aromatic) and the number of carbon atoms. The number of carbon atoms is thought to highly correlate with the boiling point of the hydrocarbons. An example matrix of the Hydrocarbon Block Method is presented in Figure 4. The composition of an oil substance could be expressed in such a matrix following GC-GC/MS analysis. Subsequently, the PBT-properties of the individual blocks could potentially be assessed by analyzing and extrapolating the PBT-properties of representative hydrocarbons for varying hydrocarbon blocks (see Figure 4).
Figure 4.Theoretical example matrix of the hydrocarbon block method based on functional classes (columns) and carbon number (rows). Percentages represents the relative presence of specific hydrocarbon block within an oil substance. The PBT-properties of a block can potentially be assessed by analyzing and extrapolating PBT-properties of representative hydrocarbon structures.
References
Bradbury, S.P., Carlson, R.W., Henry, T R. (1989). Polar narcosis in aquatic organisms. In Aquatic Toxicology and Hazard Assessment: 12th Volume. ASTM International.
EMBSI (2015). Assessment of Photochemical Processes in Environmental Risk Assessment of PAHs
Ewa, B., Danuta, M.Š. (2017). Polycyclic aromatic hydrocarbons and PAH-related DNA adducts. Journal of applied genetics 58, 321-330.
Homburger, F., Hayes, J.A., Pelikanm E.W. (1983). A Guide to General Toxicology. Karger/Base, New York, NY.
Jerina, D.M., Sayer, J.M., Thakker, D.R., Yagi, H., Levin, W., Wood, A.W., Conney, A.H. (1980). Carcinogenicity of polycyclic aromatic hydrocarbons: the bay-region theory. In Carcinogenesis: Fundamental Mechanisms and Environmental Effects (pp. 1-12). Springer, Dordrecht.
King, D.J., Lyne, R.L., Girling, A., Peterson, D.R., Stephenson, R., Short, D. (1996). Environmental risk assessment of petroleum substances: the hydrocarbon block method. CONCAWE report no. 96/52.
Kirman, C.R., Albertini, R.A., & Gargas, M.L. (2010). 1, 3-Butadiene: III. Assessing carcinogenic modes of action. Critical reviews in toxicology 40(sup1), 74-92.
Parkerton, T.F., Stone, M.A., Letinski, D. J. (2000). Assessing the aquatic toxicity of complex hydrocarbon mixtures using solid phase microextraction. Toxicology letters 112, 273-282.
Roberts, A.P., Alloy, M.M., Oris, J.T. (2017). Review of the photo-induced toxicity of environmental contaminants. Comparative Biochemistry and Physiology Part C: Toxicology & Pharmacology 191, 160-167.
US-EPA (1998). Carcinogenic Effects of Benzene: An Update. EPA/600/P-97/001F.
Veith, G.D., Broderius, S.J. (1990). Rules for distinguishing toxicants that cause type I and type II narcosis syndromes. Environmental Health Perspectives 87, 207.
2.3.6. CFCs
(draft)
Authors: Steven Droge
Reviewer: John Parsons
Leaning objectives:
You should be able to:
realize what the ozone layer depletion was all about
understand why certain replacement chemicals are still problematic
CFCs (chlorofluorocarbons) were very common air pollutants in the 20th century because they were the basic components of refrigerants and air conditioning, propellants (in spray can applications), and solvents, since the 1930s. They are still very common air pollutants, because they are very persistent chemicals, and emissions do still continue. In the first years as refrigerants, they replaced the much more toxic components ammonia (NH3), chloromethane (CH3Cl), and sulfur dioxide (SO2). Particularly the CFCs leaking from old refrigerating systems in landfills and waste disposal sites caused high emissions into the environment. Typically, these volatile CFC chemicals are based on the smallest carbon molecules methane (CH4), ethane (C2H6), or propane (C3H8). All hydrogen atoms in these CFC molecules are replaced by a mixture of chlorine and fluorine atoms.
Figure 1. Different common refrigerants and their boiling points. Freon 134 is chlorine free. . (Source: Steven Droge)
CFCs are less volatile than their hydrocarbon analogue, because the halogen atoms polarize the molecules, which causes stronger intermolecular attractions. Depending on the substitution with Cl or F, the boiling point can be tuned to the desired point for refrigerating cooling processes. The CFCs are also much less flammable than hydrocarbon analogues, making them much safer in all kinds of applications.
Naming of CFCs
CFCs were often known by the popular brand name Freon. Freon-12 (or R-12) for example stands for dichlorodifluoromethane (CCl2F2, boiling point -29.8 °C, while methane has -161 °C), as shown in Figure 1. The naming reflects the amount of fluor atoms as the most right number. The next value to the left is the number of hydrogen atoms plus 1, and the next value to the left is the number of carbon atoms less one (zeroes are not stated), and the remaining atoms are chlorine. Accordingly, Freon-113 could apply to 1,1,2-trichloro-1,2,2-trifluoroethane (C2Cl3F3, boiling point 47.7 °C, while ethane has -161 °C). The structure of any Freon-X number can also be derived from adding +90 to the value of X, so Freon-113 would give a value of 203. The first numerical is the number of C (2), the second numerical H (0), the third numerical F (3), and the remaining substitutions are by chlorine (C2X6 gives 3 chlorines).
The reason CFC depletes the ozone layer
The key issue with CFC emissions is the reaction under influence of light (“photodegradation”) that ultimately reduces ozone concentrations (“ozone depletion”) in the upper atmosphere (“stratosphere”). Ozone absorbs the high energy radiation of the solar UV-B spectrum (280–315nm), and the ozone layer therefore prevents this to reach the Earth's surface. The even more energetic solar UV-C spectrum (100-280nm) is actually causing the formation of ozone (O3) when reacting with oxygen (O2), as shown in Figure 2. Under the influence of intense light-energy in the upper atmosphere, CFC molecules can disintegrate into two highly reactive radicals (molecules with a free electron . ), for Freon-11:
It is the radical Cl. that catalyzes the conversion of ozone back into O2. The environmentally relevant role of the fluorine atoms in CFCs is that they make these chemicals very persistent after emission, because the C-F bond is one of the strongest covalent bonds known. With half-lives up to >100 years, high CFC levels can reach the upper atmosphere. James Lovelock was the first to detect the widespread presence of CFCs in air in the 1960s, while the damage caused by CFCs was discovered only in 1974. Another undesirable effect of CFC in the stratosphere is that they are a much more potent greenhouse gases than CO2.
Figure 2.The influence of UV on formation of chlorine radicals from CFCs, oxygen radicals from O2 and oxygen radicals from disintegration of O3. Ozone is not formed in the absence of UV (night time), but can still be reacting away by chlorine radicals. (Source: Steven Droge)
CFC replacements.
In 1978 the United States banned the use of CFCs such as Freon in aerosol cans. After several years of observations of the ozone layer depletions globally (Figure 3), particularly above Antarctica, the Montreal Protocol was signed in 1987 to drastically reduce CFC emissions worldwide. CFCs were banned by the late 1990s in most EU countries, and e.g. in South Korea by 2010. Due to the persistency of CFCs it may take until 2050-2070 before the ozone layer will return to 1980 levels (which were bad already).
The key damaging feature of CFCs in terms of ozone depletion is their persistency, so that emissions reach and build up in the stratosphere (starting from 20km above the equator, but only at 7km above the poles). CFC replacement molecules were initially found simply by adding more hydrogens in the CFC structures and somewhat less Cl (HCFCs), but fractions still contributed to Cl. radicals. Later alternatives lack the chlorine atoms and have even shorter lifetimes in the lower atmosphere, and simply cannot form the Cl radicals. These “hydrofluorocarbons” (HFCs) are currently common in automobile air conditioners, such as Freon-134 (do the math to see that there is no Cl, boiling point -26.1 °C).
Figure 3.The 2006 record size hole in the ozone layer above Antarctica (Source https://en.wikipedia.org/wiki/Ozone_depletion)
Still, HCFC as well as HFCs are still very potent greenhouse gasses, so the worldwide use of such chemicals remains problematic and gives rise to new legislations, regulations, and searches for alternatives. R-410A (which contains only fluorine) is becoming more widely used but is 1700 times more potent than CO2 as greenhouse gas, equal to Freon-22. Simple hydrocarbon mixtures such as propane/isobutane are already used extensively in mobile air conditioning systems, as they have the right thermodynamic properties for some uses and are relatively safe. Unfortunately, we did not have the technological skills, nor the awareness to apply this back in the 1930s.
2.3.7. Cosmetics/personal care products
(draft)
Author: Mélanie Douziech
Reviewers: John Parsons
Learning objectives:
You should be able to:
Define what personal care products and cosmetics are
Explain how chemicals from personal care products end up in the environment
Cite and describe some of the most common chemicals found in personal care products
Keywords: wastewater, chemical function, surfactants, microbeads
Introduction
Personal care products (PCPs) cover a large range of products fulfilling hygiene, health, or beauty purposes (e.g. shampoo, toothpaste, nail polish). They are categorized into oral care, skin care, sun care, hair care, decorative cosmetics, body care and perfumes. Overall, most PCPs are classified as cosmetics and regulated accordingly. In the European Union (EU) the Cosmetic Regulation governs the production, safety of ingredients and the labelling and marketing of cosmetic products. The United States of America (USA), on the other hand, have a narrower definition of cosmetics so that products not fulfilling the definition are regulated as pharmaceuticals (e.g. sunscreen) (Food and Drug Administration, 2016).
PCPs come in a range of formats (e.g. liquids, bars, aerosols, powders) and typically contain a wide range of chemicals, each fulfilling a specific function within the product. For example, a shampoo can include cleansing agents (surfactants), chemicals to ensure product stability (e.g. preservatives, pH adjusters, viscosity controlling agents), diluent (e.g. water), perfuming chemicals (fragrances), and chemicals to influence the product’s appearance (e.g. colourants, pearlescers, opacifiers). The chemicals present in PCPs ultimately enter the environment either through air during direct use, such as the propellants in aerosols, or through wastewater via down the drain disposal following product use (e.g. shower products, toothpaste). The release of PCP chemicals into the environment needs to be monitored and the safety of these chemicals understood in order to avoid potential problems. In developed countries, the use of wastewater treatment plants (WWTPs) is key to effectively removing the PCP chemicals and other pollutants from wastewater prior to their release to rivers and other watercourses. The removal mechanisms occurring in WWTPs include biodegradation, sorption onto solids, and volatilization to the air. The extent of removal is influenced by the physicochemical properties of the chemicals and the operational conditions of the WWTPs. In regions where wastewater treatment is lacking, the chemicals in PCPs enter the environment directly.
The wide scale daily use of PCPs and the associated large volumes of chemicals released explain why they are scrutinized by environmental protection agencies and regulatory bodies. The following sections will briefly review some of the classes of chemicals used in PCPs by describing their behavior in the environment and their potential effect on ecosystems.
Cleansing agents - surfactants
Surfactants are an important and widely used class of chemicals. They are the key components of many household cleaning agents as well as PCPs, such as shampoos, soaps, bodywash and toothpaste, because of their ability to remove dirt. These dirt-removing properties also make surfactants inherently toxic to aquatic organisms. The biodegradability of surfactants is a key legal requirement for their use in PCPs to minimize the likelihood of unsafe levels in the environment. Different types of surfactants exist and are often classified based on their surface charge. Anionic surfactants, which carry a negative surface charge, interact and help remove positively charged dirt particles from surfaces such as hair and skin. Sodium lauryl sulfate is a typical example of an anionic surfactant used in PCPs. Cationic surfactants, such as cetrimonium chloride, are positively charged and may be used as hair conditioning agents to make hair shinier or more manageable. Non-ionic surfactants (uncharged), such as cetyl alcohol, help formulate products or increase foaming. Amphoteric surfactants, such as sodium lauriaminodipropionate, carry both positive and negative charges and are commonly used to counterbalance the potentially irritating properties of anionic surfactants.
Fragrances
Fragrances are mixtures of often more than 20 perfumery chemicals used to provide the smell of PCPs. Typically, fragrances are present at very low levels in most PCPs (below 0.01%) so that their exact compositions are not disclosed. Disclosed, however, are any allergens present in the fragrance to help dermatologists and consumers avoid certain fragrance chemicals. Despite the wish to protect trade secrets, a recent trend increasingly sees companies disclose the full fragrance compositions of their products on their websites (e.g. L’Oréal, Unilever). Well-known examples of fragrances include hexyl cinnamal, linalool, and limonene. Potential concerns about the ecotoxicological impact of fragrances have arisen on the one hand because of a lack of disclosure of fragrance formulations and on the other hand because of the detection of certain persistent fragrances in the environment (e.g. nitromusks).
Preservatives
Preservatives are usually added to PCPs containing water for their ability to protect the product from contamination by bacteria, yeasts, and molds during storage or repeated use. Given their targeted action against living organisms, the use of preservative in chemical products including PCPs is under constant scrutiny. For example, in 2016 and 2017, the European Commission tightened the regulation around the use of methylisothiazolinone in cosmetics products due to human safety concerns. Other preservatives that have been restricted in use, because of both human safety and environmental safety concerns (e.g. endocrine disruption effects), include certain types of parabens and triclosan.
UV filters
UV filters are used in sunscreen products as well as in other PCPs such as foundation, lipstick, or moisturizing cream to protect users from UV radiation. UV filters can be organic or inorganic. Inorganic UV filters, like titanium oxide and zinc oxide, form a physical boundary protecting the skin from UV radiation. Organic UV filters, on the other hand, protect the skin by undergoing a chemical reaction with the incoming UV radiation. Organic UV filters commonly found in PCPs include butyl methoxydibenzoylmethane, ethylhexyl methoxycinnamate, and octocrylene. Organic UV filters are poorly biodegradable and have the potential to accumulate in organisms. Further, a number of organic UV filters have been shown to be toxic to coral organisms in laboratory tests. They are suspected to cause coral bleaching by, for example, promoting viral infections but research is still on-going to understand their potential ecotoxicological effects at realistic environmental concentrations.
Volatile chemicals
Certain chemicals used in PCPs are highly volatile and may end up in the air following product use. Examples include propellants, such as propane butane mixes or compressed air/nitrogen, used in aerosols to apply ingredients in hairsprays or deodorants and antiperspirants. Fragrances also volatilize when the product is applied to skin or hair to provide smell. Volatile silicones, chemicals used to assist the deposition of ingredients in liquids and creams, are another example of chemicals emitted to air upon PCP use.
The special case of plastic microbeads
Plastic microbeads, with a diameter smaller than 5mm, have been used in PCPs such as face scrubs or shower gels for their scrubbing and cleansing properties. The growing concern about plastic pollution in water has drawn attention to the use of microbeads in PCPs. As a result, a number of initiatives were launched both to highlight the use of plastic microbeads and to encourage replacement with natural alternatives. An example thereof is the “Beat the microbead” coalition (https://www.beatthemicrobead.org/) sponsored by the United Nations Environment Program, launched to help consumers identify and avoid PCPs containing microbeads. Such initiatives together with voluntary commitments by industry have led to a large decrease in the use of microbeads in wash-off cosmetic products: In the EU, for example, the use of microbeads in wash-off products was reduced by 97% from 2012 to 2017. Legislation to restrict the use of microbeads has also recently been put in place. In the USA microbeads in PCPs were banned in July 2017 and a number of EU countries (e.g. United Kingdom, Italy) have also banned their use in wash-off products.
Further reading
For more information on PCP chemicals and their function in products, please see (European commission 2009; Grocery Manufacturers Association 2017).
For more information on the different types of surfactants, please see Tolls et al. (2009) and Section 2.3.8.
Manova et al. (2013) list the different types of UV filters.
The report of Scudo et al. (2017) gives more information on the use of microplastics in Europe.
Grocery Manufacturers Association (2017). Smartlabel. [cited 2017 11]; Available from: http://www.smartlabel.org/.
Manova, E., von Goetz, N., Hauri, U., Bogdal, C., Hungerbuhler, K. (2013). Organic UV filters in Personal Care Products in Switzerland: A Survey of Occurrence and Concentrations. International Journal of Hygiene and Environmental Health 216, 508-514.
Scudo, A., Liebmann, B., Corden, C., Tyrer, D., Kreissig, J., Warwick, O. (2017). Intentionally Added Microplastics in Products. in: Limited A.F.W.E.a.I.U., ed. United Kingdom
Tolls, J., Berger, H., Klenk, A., Meyberg, M., Müller, R., Rettinger, K., Steber, J. (2009). Environmental safety aspects of Personal Care Products - a European perspective. Environmental Toxicology and Chemistry 28, 2485-2489.
2.3.8. Detergents and surfactants
Author: Steven Droge
Reviewer: Thomas P. Knepper
Leaning objectives:
You should be able to:
explain why surfactants remove dirt
discuss historical progress on surfactant biodegradability
describe the different types of common surfactants.
describe examples of how surfactants enter the environment.
Surface active agents (“surf-act-ants”) are a wide variety of chemicals produced in bulk volumes (>10.000 tonnes annually) as a key ingredient in cleaning products: detergents. Typical for surfactants is that they have a hydrophobic tail and a hydrophilic head group (Figure 1).
Figure 1.Different forms (micelle and surfactant monomer) and types of surfactants. (Source: Steven Droge)
At relatively high concentrations in water (typically >10-100 mg/L), surfactants spontaneously form aggregated structures called micelles (Figure 1), often in spheres with the hydrophobic tails inward and the hydrophilic head groups towards the surrounding water molecules. These micelle super-structures allow surfactants to dissolve grease and dirt from e.g. textile or dishes into water, which can then be flushed away. Besides this common use of surfactants, their amphiphilic (i.e., both hydrophilic and lipophilic) properties allow for a versatile use in our modern world:
• During the large 2010 oil spill in the Mexican Gulf, enormous volumes (>6700 tonnes) of several types of surfactant formulations (e.g. "Corexit") were used to disperse the constant stream of oil leaking from the damaged deep water well into small dissolved droplets, in order to facilitate microbial degradation and prevent the formation of floating oil slabs that could ruin coastal habitats.
• The ability of a layer of surfactants to maintain hydrophobic particles in solution is a key process in many products, such as paints and lacquers.
• The ability to emulsify dirt particles is a key feature in process fluids during deep drilling in soil or sediment.
• Fabric softners, and hair conditioners, have cationic surfactants as key ingredients that stick with the positively charged head groups onto the negatively charged fibers of your towel or hair. After the final flushing, these cationic surfactants still stick on the fibers and because of the hydrophobic head groups sticking out make these materials feel soft and smooth. Often only during the next washing event (with anionic or nonionic surfactants) the cationic surfactants are flushed off the fibers.
• Many cationic surfactants have biocidal properties at relatively low concentrations and are therefore used in a few percent in many cosmetic products as preservatives, e.g. in cosmetics, or used to kill microbes in food processing, antibacterial hand wipes or during swimming pool cleaning. Examples are chloride salts of benzalkonium, benzethonium, cetylpyridinium.
• Surfactants lower the surface tension of water, and therefore are used (as “adjuvants”) in pesticide products to facilitate the droplet formation during spraying and to improve contact of the droplets with the target leaves in case of herbicides. Examples are fluorinated surfactants, silicone based surfactants (Czajka et al. 2015), and polyethoxylated tallow amine (POEA) used for example in the glyphosate formulation Roundup.
The hydrophobic tail of surfactants is mostly composed of a chain of carbon atoms, although also fluorinated carbon (-CF2-) chains or siloxane (Si(CH3)3-O-[..]-Si(CH3)3) chains are also possible.
The first bulk volume produced surfactants for washing machines were branched anionic alkylbenzenesulfonates (ABS) and alkylphenolethoxylates (APEO), with the hydrocarbon source obtained from petroleum. Because of the variable petroleum source, these chemicals are often complex mixtures. However, hydrophobic branched alkylchains are poorly biodegraded, and the constant disposal of these surfactants into the waste water caused very high environmental concentrations, often leading to foaming rivers (Figure 2).
Figure 2.Foaming in sewage treatment and on rivers in the 1950s caused by non-biodegradable tetrapropylene sulfonate (TPS) (Source: Kümmerer 2007, who obtained permission from the Archives of the Henkel Company, Düsseldorf).
Surfactant producers ‘voluntarily’ switched to carbon sources such as palm oil or controlled polymerization of petrol-based ethylene, that could be used to generate surfactants with linear alkyl chains: linear alkylbenzenesulfonate (LAS) and alcohol ethoxylates (AEO). Some surfactants have the hydrophilic headgroup attached to two carbon chains, such as the anionic docusate (heavily used in the BP oil spill) and the cationic dialkyldimethylammonium chemicals. Common detergent surfactants are nowadays designed to pass ready biodegradability tests (>60% mineralisation to CO2 within a 10 d window following a lag phase, in a 28 d test). Early examples of fabric softners are double chain (dialkyl)dimethylammonium surfactants, but the environmental persistency of these compounds (DODMAC and DHTDMAC, see e.g. EU and RIVM-reports) has led to a large replacement by diesterquats (DEEDMAC), which degrade more rapidly through the weak ester linkages of the fatty acid chains (Giolando et al. 1995). A switch to sustainable production of the carbon sources is ongoing. Whereas petroleum based ethylene oil was mostly used, it is being replaced increasingly by the linear fatty acid carbon chains from either palm-oil (mostly C16/C18), coconu oil (mostly C12/C14), but also such raw materials needs to be as sustainably derived as possible.
The hydrophilic headgroups can vary extensively. Nonionic surfactants can have a simple polar functional group (amide), glucose based (polyglycoside), or contain a variable lengths of repetitive ethoxlyate and/or propoxylate units. Because the ethoxylation process is difficult to control, such surfactants are often complex mixtures. Anionic surfactants are often based on sulfate (SO4-) or sulfonate (SO4-), but also phosphonate and carboxylates are common. A key difference between anionic surfactants is that sulfate and sulfonates are fully anionic (pKa ~<0) over the entire environmental pH range (pH4-9), while carboxylates are weaker acids that are still partially neutral species (pKa ~5). Most cationic surfactants are based on permanently charged quaternary ammonium headgroups (R-(N+)(CH3)3), although several ionizable amine groups are applied in cationic surfactants too (e.g., diethanolamines).
The key ingredient property of most surfactants is the critical micelle concentration (CMC), which defines the dissolved concentration above which micellar aggregates start to form that can remove grease or fully emulsify particles. The CMC decreases proportionally with the hydrophobic tail length, and this means that with longer tails, you need less surfactant to start to form micelles. However, with increasing hydrophobic tails anionic surfactants more readily precipitate with dissolved inorganic cations such as calcium. Also, surfactant toxicity increases proportionally with hydrophobic tail lengths. If the alkyl chain is too long, the surfactant may bind strongly to all kinds of surfaces and not be available for micelle formation. The optimum hydrophobic chain length is thus often a balance between the desired properties of the surfactant and several critical processes that influence the efficiency and risk of surfactants.
References
Kümmerer, K. (2007). Sustainable from the very beginning: rational design of molecules by life cycle engineering as an important approach for green pharmacy and green chemistry. Green Chemistry 9, 899–907 . DOI: 10.1039/b618298b
Giolando, S.T., Rapaport, R.A. , Larson, R.J., Federle, T.W., Stalmans, M., Masscheleyn P. (1995). Environmental fate and effects of DEEDMAC: A new rapidly biodegradable cationic surfactant for use in fabric softeners. Chemosphere 30, 1067-1083. DOI: 10.1016/0045-6535(95)00005-S
Czajka, A., Hazell, G., Eastoe, J. (2015). Surfactants at the Design Limit. Langmuir 31, 8205−8217. DOI: 10.1021/acs.langmuir.5b00336
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