13.1 Introduction

Identifying and listing substances or materials as contaminants of emerging concern (CECs) is not a simple task, and for the marine environment specifically is a challenge for environmental regulators, managers and researchers worldwide (Box 13.1) (Tornero and Hanke 2017). Some of these agencies have widely different definitions of what a CEC actually is (Halden 2015).

The meaning of the term contaminant is relatively well understood and is discussed in Chapter 1 of this book. Although the text used by various authors and agencies to define contamination varies, it usually includes or implies the involvement of human-related activities and results in the production of an unnatural concentration of material in a specific environment leading to an associated adverse consequence or impairment to the natural condition for one or more attributes of that environment. However, the terms emerging and concern are more subjective, and are subject to time scales and prevailing circumstances.

13.1.1 What is Meant by “Emerging”?

A meta-analysis of 143,000 publications about 12 prominent CECs ranging from the pesticide DDT to nanoparticles and microplastics (Halden 2015) showed a common time course of emergence and subsidence of concern spanning about 29 years. That study noted that a number of factors can trigger and accelerate the emergence of new CECs, for example, new methods of detection and lowered detection limits, paradigm shifts in scientific understanding, breakthroughs in the design and manufacture of materials and changes in marketing and consumer behaviour leading to increased chemical consumption. Each of these factors can bring long-ignored environmental contaminants into the public eye and drive an increasing level of concern. This increase in the level of concern about a substance or material often triggers further research and publishing activity and the development of new regulations.

13.2 Contaminants of Emerging Concern in the Marine Environment

The list of potential candidate substances to be CECs in the marine environment is very large. In excess of 100 million chemical substances are currently registered in the Chemical Abstracts Service (CAS) and about 4000 new ones are registered every day. The number of registered and pre-registered substances in REACH (the European Union legislation for the Registration, Evaluation, Authorisation and Restriction of Chemicals) lists 30,000–50,000 industrial chemicals present in daily-use products, all of which are potentially ultimately released into the environment (Dulio et al. 2018). However, not all of these chemicals are of concern once released to the environment, and many are unlikely to become CECs. Numerous international environmental agencies and regulators have compiled individual lists of chemicals and substances they regard as being of concern but there is no common list accepted by all the relevant organizations.

The European Commission Joint Research Centre has compiled a “comprehensive list of chemical substances considered relevant” under European Union legislation and by international organizations (Tornero and Hanke 2017). Although not all of the listed contaminants are of concern for the marine environment, this list is invaluable in presenting in one table the total of approximately 2700 of concern substances (or groups of substances) identified under relevant global conventions (e.g. the Stockholm Convention on POPs), European legislation (e.g. REACH), government agencies (e.g. the United States Envrionmental Protection Agency (US EPA) Priority Pollutants legislation) and international research groups (e.g. the NORMAN Network), together with the status of each contaminant on its source list.

Several large-scale monitoring programs in the marine environment have focused on detecting and monitoring emerging contaminants of concern. The most well known (and possibly the largest, longest lasting and best resourced) of these programmes is the Mussel Watch Program conducted by the National Oceanic and Atmospheric Administration (NOAA) in North America since 1986 (NOAA 2008). (See also Chapter 2, Box 2.1).

In the marine environment, a good example of an emerging contaminant of worldwide concern at the time (some 50 years before the present) is provided by the emergence of concern over the use of tributyltin (TBT) as an active ingredient in anti-fouling coatings applied to the hulls of ships. The published scientific material on the TBT issue is very extensive, but an overview published by the European Environment Agency (EEA) is succinct and comprehensive (Santillo et al. 2002) (see also Chapter 7). In summary, the use of antifoulant protection on the submerged portion of ship hulls is essential to minimize the growth of marine life (fouling) that causes hull damage to timber vessels and reduces speed and increases fuel consumption in all affected vessels regardless of the material from which they are constructed. Initially, wooden ships were protected with metallic copper sheathing. In later times, copper-containing paints were used on vessels of all types, and in the late 1960s, organotin compounds (in particular, TBT) were found to be a very effective ingredient in anti-fouling paints and these compounds rapidly became the active ingredient of choice in hull paints and use was widespread by the early 1970s.

However, the widespread use of TBT-based antifoulant paints by commercial shipping, including fishing fleets, and by leisure craft became associated with a marked decline in many commercially important marine mollusc fisheries (for example, mussels and oysters), characterized by declining populations of many resource species especially where there was a high density of boat traffic. Research demonstrated the toxic consequence of the exposure of marine molluscs to low (ng/litre) concentrations of water-borne TBT was primarily imposex (the development of male sexual structures in females—leading to reproductive failure), but also shell deformities, failure of larval settlement and bioaccumulation of TBT. (See also Box 7.2).

Subsequently, TBT was found to be environmentally persistent, particularly in sediments (a half-life of 4 years) but much less so in waters (half-life of 6 days), and increasing concentrations were found in the tissues of a wide range of marine life including fish and marine mammals. The sources of TBT to the marine environment were not only its release from vessel coatings, but also from poorly or non-regulated disposal of TBT-containing paint residues stripped from vessels when hulls were repaired and when regularly scheduled repainting was carried out (Figure 13.2).

Figure 13.2
figure 2

Photo: M. Mortimer

Stripping TBT-based antifoulant paint from a ship hull during drydocking for maintenance and repaint in the Port of Brisbane, Australia during the 1990s. After work completion the drydock was re-flooded, and paint debris accumulated on the dock floor was flushed into the river.

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The progressive introduction from 1982 by countries and international organizations (see timeline in Santillo et al. 2002) of restrictions on the use of TBT-based antifoulants, culminating in their effective phase-out by the International Maritime Organization (IMO), a specialized agency of the United Nations responsible for regulating shipping in its adoption of the International Convention on the Control of Harmful Antifouling Systems on Ships (IMO 2001). This convention imposed a global prohibition of the application of organotin compounds which act as biocides in anti-fouling systems on ships by 1 January 2003 and a complete prohibition of the presence of organotin compounds which act as biocides in anti-fouling systems on ships by 1 January 2008 and has succeeded in successfully managing the TBT contamination problem.

13.3 The Relationship Between CECs and Endocrine Disrupting Chemicals

Since the late 1980s, there has been growing evidence of the feminization of male fish in waters receiving sewage treatment plant discharges (e.g. Jobling et al. 1996, 1998) and this triggered concern in the general community and the attention of regulatory agencies and researchers concerning the presence of estrogenic chemicals in outfalls and receiving waters.

Common usage of the term endocrine disruption in the context of chemical pollution originated in 1991 as a consensus statement at a conference workshop series publication in Wisconsin, USA (Colborn and Clement 1992). The convenor of that conference, Theo Colborn along with others, subsequently published the book Our Stolen Future (Colborn et al. 1997), a landmark publication in raising public attention to the issues relating to endocrine disruption in wildlife and potentially humans.

The growing attention, given to endocrine disrupting chemicals (EDCs) phenomenon (Box 13.3), raised the concern levels about contaminants in aquatic environments, and in particular chemicals that are EDCs, and as a consequence numerous chemicals and substances became CECs. However, it is important to note that the EDC phenomenon is an expression of toxic effect, and although many CECs are associated with the EDC phenomenon, many are regarded as CECs for other reasons.

13.4 Pharmaceuticals and Personal Care Products (PPCPs) as CECs

The group of chemicals and substances collectively known as PPCPs includes both pharmaceuticals and personal care products used for personal health/well-being or for cosmetic purposes (see Chapter 12). The common usage of the term PPCPs also includes non-medicinal/non-cosmetic household products or their ingredients such as disinfectants (e.g. triclosan) and antiseptics, soaps, detergents and other cleaning products, synthetic musks and fragrances cosmetics, lotions, preservatives and sunscreen agents (e.g. oxybenzone). A recent overview of the global extent of discharges of PPCPs was provided in Dey et al. (2019).

Pharmaceuticals are defined as prescription, over-the-counter and veterinary therapeutic drugs used to prevent or treat human and animal diseases, while personal care products are used mainly to improve the quality of daily life (Boxall et al. 2012).

Pharmaceuticals can be classified by their therapeutic uses. The common uses being: anti-diabetics (e.g. alpha-glucosidase inhibitor), ß-blockers (e.g. atenolol, metoprolol), antibiotics (e.g. trimethoprim), lipid regulators (e.g. gemfibrozil), anti-epileptic (e.g. acetazolamide), tranquilizers (e.g. diazepam), anti-microbials (e.g. penicillins), anti-ulcer and anti-histamine drugs (e.g. cimetidine, famotidine), anti-anxiety or hypnotic agents (e.g. diazepam), anti-inflammatories and analgesics (e.g. ibuprofen, paracetamol, diclofenac), anti-depressants (e.g. benzodiazine-pines), anti-cancer drugs (e.g. cyclophosphamide, ifosfamide), anti-pyretics and stimulants (e.g. dexamphetamine, methylphenidate, modafinil), and estrogens and hormonal compounds (e.g. estriol, estradiol, estrone).

Currently more than 5000 manufactured pharmaceutical medicines are consumed by humans and/or domesticated animals, with an estimated total annual worldwide consumption in the range of 90,000–180,000 tonnes with the largest national consumptions being Russia, China, South Africa, India and Brazil (Van Boeckel et al. 2015; Tijani et al. 2016). A comprehensive overview of the current understanding of the extent and potential impact of contamination of the marine environment by pharmaceuticals is provided in the recent review by Ojemaye and Petrik (2019).

A large portion of medications that are ingested orally or by infusion are excreted through urine and/or faeces due to their incomplete absorption (metabolism) in humans and animals, these ultimately end up in wastewater treatment plants. Subsequently, municipal sewage treatment plants are major points of release of pharmaceuticals into the marine environment because wastewater treatment plants are not designed to decompose the vast majority of pharmaceutical compounds, which are by intent stable and robust, polar and non-volatile in nature. The most frequent and widespread pharmaceuticals in sewage and the discharge from marine outfalls are antibiotics and nonsteroidal anti-inflammatory drugs (Ojemaye and Petrik 2019). Other pathways for pharmaceuticals to be delivered into the marine system are via landfill sites, septic tanks, urban wastewater, showering and bathing, industrial effluent and agricultural runoff.

Measured concentrations of pharmaceuticals from worldwide coastal environment locations in seawater, sediments and organisms (Ojemaye and Petrik, 2019) range from 0.21 to 5000 ng/L (seawater), 0.0402 ng/g dry weight to 208 ng/g wet weight (biota) and 0.2 µg/kg dry weight to 466 µg/kg wet weight (sediments). However, despite evidence of their increasing presence, little attention has been directed towards understanding the release of pharmaceuticals into coastal-marine environments and their potential negative impact on marine ecosystems. This qualifies many pharmaceuticals as CECs in the marine environment.

Since the active ingredients in pharmaceuticals are chosen on the basis that their physicochemical and biological properties can produce specific biological effects in humans and animals, they have a high potential to trigger negative impacts on non-target organisms. In addition, anti-infection agents could create an ecological hazard by advancing the spread of resistant genes in the environment (Costanzo et al. 2005).

Other concerns are that the metabolites of many pharmaceuticals are potentially active and unsafe in the environment. For example, paracetamol and amitriptyline are mostly metabolized into highly reactive compounds (Graham et al. 2013). Also, of concern is that pharmaceuticals are discharged into the marine environment from sewage treatment plants as complex mixtures thus exposing marine life to potential synergetic environmental effects. For example, a synergistic antioxidant response in fish was demonstrated in a laboratory study involving co-exposure to a mixture of fluoxetine (an antidepressant medication) and roxithromycin (an antibiotic), and also with a mixture of fluoxetine and propranolol (a β-blocker used to treat a range of cardiac disease symptoms) (Ding et al. 2016).

Similarly, some ingredients of non-medicinal/non-cosmetic household products (e.g. triclosan—a widely used bactericide in healthcare products such as skin care ointments and lotions, mouthwashes and toothpastes, shower gels and shampoos) are not efficiently broken down in typical municipal sewage treatment plants, so that the end-of-treatment discharges from these facilities are major point sources of release into the marine environment (Cui et al. 2019). Triclosan is persistent and bioaccumulative in the aquatic environment and triggers a number of toxic responses (Maulvault et al. 2019).

The array of PPCPs in sewage discharges is extensive, but the potential for adverse effects is largely unknown for most of the active ingredients present (Ojemaye and Petrik 2019). The NORMAN Network (Box 13.2) currently lists almost 300 PPCPs substances as CECs.

13.5 Nanomaterials

The manufacture and use of nanoparticles and nanostructured materials (also known as nanomaterials) is an expanding field of modern technology. As a consequence, the perceived risks associated with potentially toxic properties of these novel materials have resulted in their attracting attention as a new class of CECs.

By their nature, nanoparticles are units of particulate materials with a maximum dimension sized in nanometres (10–9 m). Although there is no single internationally accepted definition for nanomaterials (Jeevanandam et al. 2018), they are commonly defined as materials in which a single unit is sized in the range of 1–100 nm in at least one dimension. The term aerosols is often applied to nanoparticles when they are airborne, for example, in wind-borne dust or otherwise suspended in the atmosphere. The US EPA routinely uses the term ultrafine particles when discussing natural nanomaterials and aerosols. A summary of types and classifications of nanomaterials, and common technical descriptors is at Box 13.4.

Interestingly, the use and manufacture of nanomaterials are not an entirely modern phenomenon. The Ancient Egyptians used nanoparticulate lead sulfide as a hair dye some 4000 years ago (Walter et al., 2006) and more recently (400‒100 BC) red enamels used by Ancient Celtic cultures were based on nanoparticulate copper oxides (Brun et al. 1991) and stained glass in medieval churches incorporated gold and silver nanoparticles (Schaming and Remita 2015).

The origin and source of nanoparticles and nanomaterials is diverse (Box 13.5). Naturally occurring nanoparticles (colloids) and nanomaterials are widespread in both the living and inanimate world. In addition, nanoparticles and nanomaterials may be produced as an incidental by-product of an industrial process, or they may be manufactured explicitly by an engineered process to exploit specific features that stem from their small size.

The application of nanoparticulate and nanostructured materials has increased over the past decade because they provide enhanced or unique physicochemical properties (e.g. melting point, wettability, electrical or thermal conductivity, catalytic activity, light absorbance or scattering) that are different from those of their bulk counterparts. Manufactured nanomaterials can significantly improve the characteristics of bulk materials, in terms of strength, conductivity, durability and lightness, and they can provide useful properties (e.g. self-healing, self-cleaning, anti-freezing and antibacterial) and can function as reinforcing materials for construction. By 2014, some 1814 nanotechnology-based consumer products were commercially available in over 20 countries (Vance et al. 2015). Examples of the incorporation of nanoparticles in consumer products include titanium oxide nanoparticles as a white pigment in paints, cosmetic creams and sunscreens, and silver nanoparticles used in numerous personal care products such as air sanitizers, wet wipes, shampoos and toothpastes, as well as in clothing and laundry fabric softeners (PEN 2019). Nanoscale zero-valent particulate iron (nZVI) is a widely used remediant for treating toxic wastes due to its large specific surface area and high reactivity (Stefaniuk et al. 2016).

Unfortunately, the highly sought physicochemical properties of nanomaterials that have led to their increasing applications can also have an associated environmental downside. For example, nanoparticulate zinc oxide (ZnO) used in sunscreens is toxic to marine algae largely because of its dissolution as Zn2+ (Franklin et al. 2007), and nZVI use in contaminant remediation presents a range of potentially harmful environmental consequences that are not well understood (Stefaniuk et al. 2016). The enhanced toxic potential of nanosized materials may arise from their capacity to penetrate and disturb the cells and cellular systems of living tissues.

The challenge for regulators is to determine whether nanomaterials should be regulated in the same way as micron-sized particles. Among metal nanomaterials, cerium dioxide (used as a diesel fuel additive) and nanosilver are more toxic than their micron-sized forms, whereas because of their solubility there is no difference in toxicity for zinc oxide nano- and micron-sized particles in freshwaters (Batley et al. 2013). The enhanced surface area of nanosized materials can result in different cellular uptake rates, oxidative mechanisms and processes including translocation relative to that of exposure to the same material when it is not nanosized (Oberdörster et al. 2005). In the environment, aggregation is a common feature of nanomaterials, and often coatings are used (e.g. citrate or polyvinylpyrrolidone (PVP)) to minimize this. Aggregation is greatest in marine waters due to their high ionic strength, leading to sizes > 100 nm and in many cases resulting in sedimentation (Klaine et al. 2008). The presence of organic particles such as those formed from extracellular polymeric substances can briefly stabilize nanomaterials (<48-h) (Gondikas et al. 2020). Seawater enhances the dissolution of silver from coated Ag nanomaterials, largely through chloride complexation, which reduces silver toxicity (Angel et al. 2013).

Some nanoparticles and nanomaterials are released directly into the environment from the use of consumer products (e.g. silicon nanoparticles in car tyres are released by abrasion in normal vehicle use), or indirectly (e.g. nanoparticles in pharmaceuticals and cosmetics can end up in sewage, and then be discharged to the marine environment).

13.6 PFAS (Per- and Polyfluoroalkyl Substances)

The term PFAS (per- and polyfluoroalkyl substances) applies to the set of more than 4700 synthetic substances manufactured and used in a variety of industries since the 1940s (OECD 2019), and some have been classified as POPs (Chapter 8). All PFAS constitute an array of highly persistent environmental CECs that has triggered a global response by research and regulatory organizations over the past two decades.

PFAS comprise a set of compounds each of which has a molecular structure comprising an aliphatic moiety (i.e. a group of covalently bonded carbon atoms in a straight or branched chain, and in some cases including non-aromatic rings) that is highly fluorinated and linked to a functional group moiety. This PFAS molecular structure can be conceptualized as an alkyl tail of carbon atoms with fluorine atoms attached to a a functional group head (Figure 13.3). The degree of fluorination of the aliphatic moiety in a PFAS structure can be partial or total. In polyfluoroalkyl substances, fluorine atoms replace only some of the hydrogen atoms in the aliphatic chain, whereas in perfluoroalkyl substances, fluorine atoms replace all of the hydrogen atoms in the aliphatic chain. The general formula for a perfluorinated PFAS is CnF2n+1-R where n is 3 or greater and -R is a functional group such as carboxylic acid (COOH), sulfonic acid (SO3H) or sulfonamide (SO2NH2) (Figure 13.3).

Figure 13.3
figure 3

Typical perfluorinated PFAS molecules showing the basic structure comprising a perfluorinated alkyl tail attached to a functional group head. Structures here are the linear isomers. A mixture of linear and branched isomers may be present in an environmental sample. Adapted from Mueller and Yingling (2017) by M. Mortimer

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Note that the term PFAS sometimes appears in print in the context of more than one fluorinated chemical, but the addition of the s is redundant since the acronym PFAS includes the plural (ATSDR 2017). Also, it is important to note that PFAS (per- and polyfluoroalkyl substances) are sometimes called perfluorinated chemicals and the acronym PFC is then used. However, this use of the PFC acronym can be confusing since PFC is also commonly used for a related, but distinctly different group of substances: the perfluorocarbons (Box 13.6).

The range of structurally related compounds comprising more than 4700 member group of PFAS substances is illustrated in the PFAS family tree in Figure 13.4.

Figure 13.4
figure 4

The PFAS family tree with examples. Adapted from Wang et al. (2017) by M. Mortimer. PFCAs = Perfluoroalkyl carboxylic acids; PFSAs = Perfluoroalkane sulfonic acids; PFPAs = Perfluoroalkyl phosphonic acids; PFPiAs = Perfluoroalkyl phosphinic acids; PFECAs and PFESAs = Perfluoroether carboxylic and sulfonic acids; PASF = Perfluoroalkane sulfonyl fluoride

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13.7 Summary

There is a large body of research papers and reports concerning the topic of Contaminants of Emerging Concern (CECs) but the term itself is not definitive since both emerging and concern may be subjective, and the list of materials identified as CECs changes over time and in response to community perceptions of risks to health and the environment. The NORMAN Network is a key organization in identifying such materials and coordinating meaningful related research.

In the marine environment, since the late 1900s the priority focus has moved from concern over unintended impacts from the widespread use of organic tin-based antifoulants used on the hulls of sea-going vessels to impacts relating to a wide range of material types including EDCs, PPCPs, nanomaterials, PFAS compounds as well as environmental contamination by polymer and plastic debris.

Each of these current CECs covers a large number of chemical identities. Overall this is a wide-ranging and dynamic area of risk assessment, priority setting and ongoing scientific research.

13.8 Study Questions and Activities

  1. 1.

    In the context of the marine environment draw up a short list of up to five contaminants of emerging concern that are highlighted in recent media publications (noting that the media may not use the term contaminant of emerging concern as a descriptor) and compare this short list with contaminants which are popular topics in the programmes of recent conference presentations and journal publications. What do you suggest are reasons for similarities and differences between these two sets of CECs?

  2. 2.

    Identify two CECs in the marine environment that have been receiving frequent attention for a period longer than two or three years. Why are they still considered emerging (for example, has the baseline of residual concern changed)?

  3. 3.

    In this chapter, lead in the marine environment near Port Pirie, South Australia is used as an example. What other locations in Australia and other countries with territorial waters in the Pacific Ocean also have an emerging problem associated with lead mining and processing?

  4. 4.

    Which metallic contaminants are CECs in European marine waters?

  5. 5.

    What potential contaminants of marine waters are likely to become CECs as a consequence of the shift from fossil-fuel-based energy sources to renewables? What are some geographic locations where these may first emerge as CECs—explain why?