Global occurrence, chemical properties, and ecological impacts of e-wastes (IUPAC Technical Report)

1 Introduction

Electrical and electronic waste (e-waste) is a broad term that describes electrical and electronic equipment (EEE) that has become unwanted, non-working, or obsolete, and has essentially reached the end of its useful life. In Europe, the revised Waste Electrical and Electronic Equipment Directive (WEEE 2012/19/EU) came into force in 2018, streamlining e-waste categories from 10 to 6 (Table 1) [1]. Over the last two decades, the lifespans of the related products have become shorter and the global market of e-waste has grown exponentially. In 2016, the total tonnage of e-waste worldwide reached 44.7 × 10⁶ tonnes (t, often referred to as ‘metric tons’ in the US) and is expected to rise to 52.2 × 10⁶ t by 2021 [2]. By load (% tonnage), Asia is the biggest contributor (40.7 %), followed by Europe (27.5 %). China produced the highest e-waste quantity in Asia and overtook the United States of America (USA) to be the highest generator of e-waste in the world (7.2 × 10⁶ t) in 2017. A global map of e-waste generated (per capita) is presented in Fig. 1a. Norway produces the highest amount of e-waste per capita (∼28.5 kg); North American countries, such as USA and Canada, produced approximately 20 kg per capita, whereas countries in Latin America (LATAM) contribute up to 10 % of the e-waste generated worldwide [3] and ∼7 kg per capita [2]. Only a few LATAM countries have specific legislation on e-waste management, as the majority are regulated under general hazardous waste legislation, with e-waste management in the region generally linked to metal scrap processing by informal and private companies [3]. The Middle East and North Africa (MENA) is a relative newcomer to readily disposable electrical products, such as cellular phones and computers. Over the last decade, the consumption of electronic goods has increased markedly, especially in the Gulf nations (Saudi Arabia, Qatar, and the United Arab Emirates), Egypt, and Israel. The quantity of e-waste generated in the MENA region has been estimated to be 2.9 × 10⁶ t, with Saudi Arabia the highest contributor (0.5 × 10⁶ t) [2]. Whilst the take-up of computers, cellular phones, and household gadgets among these countries has been high, there have been little or no initiatives to handle the resulting e-waste, recover resources, and promote circularity. The African nations have the least number of direct manufacturers of EEE, but they produce around 5 % of the global e-waste (2.2 × 10⁶ t). Up to 50 % of the e-waste generation in Africa results from transboundary imports from developed countries [4]. The actual amount of e-waste in Africa is believed to be much higher, as high levels of e-waste, enters the countries through illegal routes and is therefore not captured by official audits.

Fig. 1:

(a) Global distribution of e-waste generated (kg per capita); (b) percentage increase of e-waste generated from worldwide 2014–2017.

In terms of recycling of e-waste, Europe and Africa have the highest (35 %) and lowest recycling rates (0.2 %), respectively [2]. The percentage increase of e-waste production worldwide, based on data obtained in 2014 [5] and 2017 [2], is presented Fig. 1b. A number of key e-waste-producing countries appeared to produce similar (e. g., Norway) or even reduced quantities of e-waste (e. g., USA) over the 3-year period, probably as a result of tighter regulations and the introduction of measures to reduce e-waste. Several countries not usually associated with high volumes of e-waste production showed a rapid rate of increase, most markedly in Zimbabwe, Mongolia, and Iceland (2.0-, 1.6-, and 1.5-fold increases respectively). These countries currently do not have national regulations to control e-waste [2].

Several notable international initiatives, as well as legislation are in place to limit the adverse impact of e-waste (Tables 2 and 3), but substantial quantities of historical e-waste remain a challenge in waste management. Legislative loopholes have resulted in near end-of-life e-waste being exported from developed to developing countries [6], despite the Basel Convention, which prohibits transboundary movement of e-waste. For example, e-waste is being exported under the guise of second-hand or used electrical and electronic equipment (UEEE) or recycling materials, which are not under the remit of the Basel Convention. Much of the UEEE destined for developed countries, such as the United States, Taiwan, Japan, South Korea, and Canada, is re-exported to international trade hubs, for example, Hong Kong, United Arab Emirates, Lebanon, and Macau [7], [8], [9], where the e-waste is transported to highly populated regions with very cheap labour [10], [11].

In developed countries, e-waste management has revolved around two major strategies, either: (i) recycling, recovery, and disposal within their own countries, or (ii) exportation to developing nations [9], [12]. For developing countries, the management of e-waste is further complicated by the illegal import of waste, as exemplified above, undermined by weak environmental regulations and constrained by inadequate organizational structures. Rudimentary methods, such as manual dismantling, chipping, melting, and burning, as well as uncontrolled chemical dissolution, such as leaching by strong oxidizing acids, primarily aqua regia (1:3 HNO₃–HCl concentrated solutions), are often used by the informal sector in developing countries to salvage and recover the valuable materials present in e-waste, including copper, gold, and silver. These informal recycling practices contribute to the release of toxic leachate and fumes as well as persistent organic pollutants (POPs) into the local surroundings, thereby posing risks to humans and the environment, including its biota. A number of extensive reviews have been performed in an attempt to evaluate the impacts of e-waste on human health [6], [13], [14], [15], [16], [17], [18], [19]. The potential adverse health effects of exposure to informal e-waste disposal and recycling have been reported to include elevated levels of contaminants in blood, milk, and other bodily fluids, genotoxicity, endocrine disruption, immunotoxicity, nephrotoxicity, abnormal reproductive development, intellectual impairment, and damage to different organs. In contrast, information on ecological and environmental risk assessment is less well integrated. Here, we consolidate primary literature evidence to provide an extensive review on the global generation and distribution of e-waste, the complexity (chemical and physical) of e-waste, its potential environmental impact, and the diverse approaches used to appraise the associated environmental risks.

Rapid economic growth, coupled with urbanisation and the global demand of EEE, greatly accelerates the e-waste stream; the safe disposal and recycling of e-waste in both developed and developing countries are critical in tackling a waste phenomenon that could seriously impact human health, the ecosystem, and our limited resources. The complex and multi-faceted issues of e-waste can only be managed by engaging all the stakeholders. The knowledge provided by this technical report will be valuable to academics, manufacturers, policy makers, waste professionals, and consumers in tackling this 21^st century global challenge.

2 E-waste in ‘hotspot’ countries

Economic growth, advancement in technologies, and short hardware innovation cycles have led to a significantly higher turnover of devices compared to past decades [20], [21], [22]. It is estimated that 75 to 80 % of the estimated 50 × 10⁶ t of e-waste generated worldwide is shipped to developing countries in Africa and Asia for ‘recycling’ and disposal [6], [10]. The major flow of e-waste exports globally is presented in Fig. 2. Nigeria, Ghana, India, Pakistan, and China have been identified as hotspot destinations of e-waste in a number of studies [18], [23], [24], [25], [26]. Other popular destinations include Kenya, Chile, Malaysia, and the Philippines [10], with Dubai in UAE and Singapore serving as pre-distribution centres for e-waste coming from the EU and US to South Asian countries, mainly India and Pakistan [11], [27].

Fig. 2:

The global flow of e-waste exports.

Determining the volume and global flow of individual e-waste categories and their generated constituents is challenging [28], particularly in developing countries, where there is a lack of official records and audit trails. This is further complicated by unreliable data keeping and uncontrolled importation, lack of historical sales data of WEEE, the dumping of e-waste in landfills without any assessment of quantity and quality, and the difficulties in tracking data related to informal recycling [29].

3 Chemical composition of e-waste

The e-waste stream comprises a heterogeneous mix of metals (among them valuable elements, such as gold, silver, copper, and aluminium), metalloids,^[1] rare earth elements (REEs), glass, and plastics [including flame retardants (FRs) and other additives] (Fig. 3a,b). Due to ongoing technological developments, the exact composition changes constantly, which makes e-waste distinct from other waste streams [48], [49].

Fig. 3:

(a) Example of e-waste ready for recycling (source: K. Isimekhai); (b) material constituents of some equipment that end up as e-waste (adapted from Ma et al. [50]).

The hazardous chemicals present in e-waste may be released into the environment, especially if the waste is disposed of or recycled improperly. Cayumil et al. [51] have classified these hazardous chemicals into primary and secondary contaminants. Primary contaminants are important constituents that are initially present in the manufacture of EEE due to their special intrinsic characteristics and includes metals such as copper, cadmium, lead, nickel, and zinc; metalloids, such as antimony (Sb); organic compounds like polychlorinated biphenyls (PCBs); liquid crystals; and FRs. In contrast, secondary contaminants are the by-products generated after the improper recycling of e-waste or during the recovery of valuable materials. For example, polycyclic aromatic hydrocarbons (PAHs), polychlorinated dibenzo[1,4]dioxins and dibenzofurans (PCDD/Fs) are of particular concern in regions where there is a lack of environmental regulations.

3.3 Halogenated compounds in e-waste

A number of halogenated compounds are present in e-waste, mainly in the form of FRs used in WPCB, plastic housing, keyboards, chargers, and cables [72], and liquid crystals in LCDs. Many of these compounds are POPs regulated by the Stockholm Convention, for example chlorinated compounds like PCDD/Fs, PCBs, and the brominated flame retardant (BFR) group [e. g., polybrominated biphenyls (PBBs), polybrominated diphenyl ethers (PBDEs), hexabromocyclododecanes (HBCDDs) and tetrabromobisphenol-A (TBBP-A)], amongst others. Liquid crystals found in LCDs are mainly organic compounds comprising carbon, hydrogen, oxygen, and fluorine. Liquid crystals used as mixtures in display technology have been deemed non-toxic, but their uncontrolled burning/incineration could well give rise to hydrogen fluoride and organofluoro-compounds.

As one of the main manufactured halogenated flame retardants, BFRs have been used in products since the 1960s [73]. The use of BFRs has been of concern for the past two decades due to their persistency and their bioaccumulative and toxicity characteristics [74]. PBDEs have been sold as three main commercial mixtures: pentaBDE, octaBDE, and decaBDE, each of which consists of several BDE congeners. The use of penta- and octaBDE in products was restricted in the EU in early 2000 [75] and this was followed in 2008 by the EU Registration, Evaluation, Authorization, and Restriction of Chemicals (REACH) regulation banning the use of decaBDE in electronics in EU [76]. Congeners of penta- and octaBDE mixtures were added to the list of POPs covered by the Stockholm Convention in 2009. Additionally, commercial mixtures of decaBDE were added to the list in 2017 [76]. A high fraction (61 %) of the cathode ray tube (CRT) casings contained more than 10,000 mg kg⁻¹ bromine from BFRs; decaBDE was the major FR used in TV sets and TBBP-A for computer CRTs [77], [78].

The halt in production and use of PBDEs has resulted in the increased application of alternative FRs with less well-known physicochemical properties. Examples of novel brominated flame retardants (NBFR) include: 1,2-bis(perbromophenyl)ethane, systematic name: 1,2,3,4,5-pentabromo-6-[2-(2,3,4,5,6-pentabromophenyl)ethyl]benzene)(DBDPE); 1,2-bis(2,4,6-tribromophenoxy)ethane, systematic name: 1,3,5-tribromo-2-[2-(2,4,6-tribromophenoxy)ethoxy]benzene (BTBPE); bis(2-ethyhexyl) tetrabromophthalate, systematic name: bis(2-ethylhexyl) 3,4,5,6-tetrabromobenzene-1,2-dicarboxylate (TBPH); 2-ethylhexyl-2,3,4,5-tetrabromobenzoate (TBB); hexabromobenzene, systematic name: 1,2,3,4,5,6-hexabromobenzene (HBB), 2,4,6-tribromophenol (TBP); decabromobiphenyl, systematic name: 1,2,3,4,5-pentabromo-6-(2,3,4,5,6-pentabromophenyl)benzene) (BB209); triphenyl phosphate (TPP); tri-p-cresyl phosphate, systematic name: tris(4-methylphenyl) phosphate (TCP); tris(2-chloroethyl) phosphate, systematic name: tris(2-chloroethyl) phosphate (TCEP), and tris(1-chloropropan-2-yl) phosphate (TCPP). The physicochemical properties of a list of PBDE alternatives in electronic products are presented in Table 5.

Table 5:

Molar mass, M, and decadic logarithm of n-Octanol/Water Partition Coefficient, lg K_ow, of main FRs used as PBDE alternatives in electronic products.

Abbreviation/IUPAC name	CAS number	Structure and molecular formula	M/g mol−1	lg Kow	Application in electronic products
BB209	13654-09-6		943.172	8.58	High impact polystyrene (HIPS)
Decabromobiphenyl
IUPAC name: 1,2,3,4,5-pentabromo-6-(2,3,4,5,6-pentabromophenyl)benzene
DBDPE	84852-53-9		971.226	11.1	Polystyrene (PS)
1,2-bis(perbromophenyl)ethane
IUPAC name: 1,2,3,4,5-pentabromo-6-[2-(2,3,4,5,6-pentabromophenyl)ethyl]benzene)
BTBPE	37853-59-1		687.640	8.31	HIPS, acrylonitrile–butadiene–styrene (ABS)
1,2-bis(2,4,6-tribromophenoxy)ethane
IUPAC name: 1,3,5-tribromo-2-[2-(2,4,6-tribromophenoxy)ethoxy]benzene
TBPH bis(2-ethyhexyl) tetrabromophthalate	26040-51-7		706.148	9.34	Polyurethane (PU) foam
IUPAC name: bis(2-ethylhexyl) 3,4,5,6-tetrabromobenzene-1,2-dicarboxylate
TBB	183658-27-7		549.923	7.73	PU foam
2-ethylhexyl 2,3,4,5-tetrabromobenzoate
IUPAC name: 2 ethylhexyl 2,3,4,5-tetrabromobenzoate
HBB	87-82-1		551.490	6.07	Capacitors
Hexabromobenzene
IUPAC name: 1,2,3,4,5,6-hexabromobenzene
TBP	118-79-6		330.801	3.89	As an intermediate of flame retardants (tribromophenyl allyl ether, 1,2-bis(2,4,6- tribromophenoxy)-ethane and as brominated epoxy resins
2,4,6-tribromophenol
IUPAC name: 2,4,6-tribromophenol
TPP triphenyl phosphate	115-86-6		326.288	4.59	Polymers
IUPAC name: triphenyl phosphate
TCP tri-p-cresyl phosphate	1330-78-5		368.369	5.11	Plasticizer
IUPAC name: tris(4-methylphenyl) phosphate
TCEP tris(2-chloroethyl) phosphate	115-96-8		285.482	1.47	Polymers such as polyurethanes, polyester resins and polyacrylates
IUPAC name: tris(2-chloroethyl) phosphate
TCPP tris(1-chloro-2-propyl) phosphate	13674-84-5		327.563	2.59	PU
IUPAC name: tris(2-chloropropan-2-yl) phosphate

4 Environmental levels of contaminants caused by e-waste recycling

In comparison to informal e-waste recycling, formal recycling of e-waste has received much less attention. Formal e-waste recycling activities are principally engaged in the dismantling and mechanical processing of EEE to recover valuable materials. A systematic review by Cellabos and Dong [79] reported the occupational exposure in formal recyclers is mainly via dust inhalation: the extent of exposure is dependent on the sophistication of the facilities. High levels of metal(loid)s, BFR, and novel flame retardants (NFRs) have been reported in formal e-waste recycling plants in Sweden [80], Australia [81], and China [82].

Labour intensive manual dismantling and crude recycling methods are involved in the informal recycling sector, especially in e-waste ‘hot spots’ countries (section 2). In comparison to the highly automated processes, informal recycling of e-waste can be generally grouped into three categories: (i) manual dismantling; (ii) recovery of valuable materials via open burning and/or acid stripping, and (iii) disposal of the unsalvageable waste (Fig. 4). Table 6 presents a list of informal recycling activities from different countries around the world. As a result of these crude techniques and inappropriate facilities, the informal recycling of e-waste has the potential to release a large quantity of contaminants into the recycling sites and the surrounding environment. Also, e-waste recycling sites are often located in or close to agricultural land [83] or close to water bodies [84]. Furthermore, many recycling activities have also been carried out in the backyard of family homes in India [85].

Fig. 4:

Activties associated with informal e-waste recycling – (a) manual dismantling (source: D. Chatterjee); (b) open burning of e-waste to recover valuable materials (Source: O. E. Popoola).

5 Ecological impact of e-waste on biota

A major concern of soil contamination from e-waste recycling is its impact on the biota. Analysis of soil-air exchange of pollutants has shown that soil acts as a secondary source to the atmosphere only for a few low molecular mass compounds, while the direction of the flux of most detected chemicals was from air to soil [148]. Contaminants released by e-waste activities are directly transferred to soils, then to plants and other soil organisms. For example, the mean TBBP-A level measured in biota near e-waste recycling and disposal sites in China ranges from 3.62 to 42.26 ng g⁻¹ wet mass in plant materials, 28.2 to 103.4 ng g⁻¹ dry mass in birds, 0.98 ng g⁻¹ wet mass to 1.52 ng g⁻¹ dry mass [149], indicating the potential for bioaccumulation.

The soil microbial community at e-waste sites appeared to be significantly affected by the pollutants present (e. g., PAHs, PBDEs and metals), but did not appear to correlate to the distance away from the pollution source [150]. Metal(loid)s present in an e-waste recycling site in Ziya, northern China impacted negatively on soil microbial viability, even though the levels of the chemicals were not harmful to humans [89]. Bacterial diversity was not decreased at e-waste open-burning sites, compared with a control site, possibly due to adaptation of the bacterial consortium to the environmental pollutants [151]. Other major drivers affecting the microbial composition in e-waste soil include available phosphorus, soil moisture [152], pH, texture, cation exchange capacity, total phosphorus content, and organic carbon [152], [153], [154]. Examples of the impact of metals in the environment are presented in Table 8. Key microbial processes are influenced by elevated levels of metals in e-waste recycling sites and shifts in soil microbial diversity, population, and functional activity of the microbial communities have all been observed [152], [154], [155], [156], [157]. A study that examined the combined effect of lead and BDE 209 at several e-waste recycling sites showed that the microbial richness and diversity was adversely affected [158]. Additionally, the soil biomass was significantly inhibited by these pollutants and an inverse relationship was observed for the soil respiration and metabolic quotient [159]. Metals released into the environment also affect the ecological balance in the aquatic environment or can settle into the sediment, from where they can be taken up by aquatic plants and organisms [160], [161].

Foliar uptake of atmospheric metals has been proven to be a significant pathway for metals entering into leaf and grain tissues [83]. Soil pH, redox potential, and soil composition, such as the organic matter content, play a role in the mobility and bioavailability of contaminants and, in the case of metals, chemical speciation also influences their toxicity in the plant system [162]. The root to shoot transportation of metal(loid)s is also a significant pathway. Vegetables, rice, and wild plants collected from areas surrounding residential gardens, paddy fields, and deserted land near a former e-waste incineration site in Guangdong province, China, showed levels of metals in the edible part of leafy vegetables that were significantly higher than in the edible portion of root vegetables [83], [163]. In the same studies, the levels of cadmium and lead in most vegetables exceeded the food safety limit in China. Bakare et al. [164] collected leachate from an open e-waste dumpsite in Alaba International Market in Lagos State, Nigeria, and water samples from wells between 100 and 150 m away from the dumpsite. The leachate appeared to induce chromosome aberration, decreasing mitotic index, and root growth anomalies in onions (Allium cepa).

The accumulation of contaminants up the food chain is an important pathway for the transportation of contaminants within the ecological network. Persistent toxic substances (such as PBDEs, PCDDs/PCDFs) and metals (such as lead) may end up in the ocean and may re-distribute into the environment, where bioaccumulation and biomagnification may occur [165]. Bioaccumulation of PCBs and PBDEs was observed in a number of studies. Earthworms in soil from an e-waste dismantling area in Taizhou, China, indicated that PCBs and PBDEs had higher bioaccumulation potential compared to PCDD/Fs [166]. DP was also detected in earthworms around e-waste recycling sites in China, although the biota-to-soil-accumulation factor was low (0.0007 to 1.85), with an average value of 0.23 [167].

Fu et al. [168] analysed contaminants in soil samples and apple snails in Fengjiang, southeast China, and reported a correlation between the dismantling activities and the release and transport of PCBs and PBDEs in the surrounding areas, where PCB and the PBDE (excluding BDE 209) were found to accumulate in apple snails. High levels of BDE 47 and other PBDEs were also detected in Chinese Mystery Snails (Cipangopaludina chinensis) collected in several e-waste recycling sites in China [169], [170]. DP and related compounds were detected in wild frogs (Rana limnocharis) and the freshwater food web near e-waste recycling sites in southern China [169], [171].

Fish obtained from a reservoir surrounded by several e-waste dismantling workshops in Qingyuan, South China, were found to contain high levels of BDE 47 [170]. The levels of PBDEs found in fish and shellfish in the Nayang and Lianjiang rivers (around an active e-waste site in Guiyu, China) were 15,000 times higher than levels reported from other regions and about 200 to 600 times higher than PBDE levels in bottom sediments collected from the same rivers [172]. The accumulation of the PBDE congener appeared to be species-specific, with carnivorous fish appearing to accumulate more organic pollutants [173], [174]. Nuclear abnormalities were observed in peripheral erythrocytes of African sharptooth catfish (Clarias gariepinus) in Alaba International Market in Lagos State, Nigeria [164]. The data suggested that the water and leachate samples contained clastogenic and/or aneugenic substances capable of increasing DNA damage and genome instability in the tested organisms.

Relatively high levels of ΣPBDEs (1000 to 5200 ng g⁻¹ wet mass), PBBs (110 to 340 ng g⁻¹ wet mass), and polybutylene terephthalate (PBT; 0.5 to 1.2 ng g⁻¹ wet mass) were found in birds sampled at e-waste recycling areas in China [175]. These water birds feed predominately on berries, soft fruits, and vegetables, or insects and small animals. Therefore, the detection of PBDE congeners in their tissues indicates that bioaccumulation can occur in wild bird populations [175]. Elevated DP levels and ethoxyresorufin-O-deethylase (EROD), an enzyme used as a biomonitor of persistent organic pollutants exposure, was observed in common kingfishers (Alcedo atthis) near an e-waste site [176]. Furthermore, a population abundance and species diversity study on birds in areas extended beyond the point source of e-waste pollution in South China showed a severe decline of migratory and resident bird functional assemblages [177].

Six food groups from Taizhou, China were also found to contain high levels of PCDD/Fs (at least 30 times higher than a control site). The highest level of the total WHO-TEQ was in crucian carp (10.87 pg g⁻¹ wet mass), followed by duck (3.77 pg g⁻¹ wet mass), hen eggs (2.80 pg g⁻¹ wet mass), chicken (2.43 pg g⁻¹ wet mass), rice (0.08 pg g⁻¹ wet mass), and vegetables (0.022 pg g⁻¹ wet mass) [178]. A study measuring HFRs over six years in free-range eggs in Longtang, China, showed that there was no decline in HFR levels from 2010 to 2016, despite the legislation and regulation of informal e-waste recycling activities. PBDEs are the dominant compounds in eggs, followed by DPs, PBBs, HBCDD, and DBDPE [179].

6 Contamination indices

Currently, a number of pollution indices are widely used in the literature to indicate the degree or extent of soil contaminations due to e-waste, including: Pollution Load Index (PLI), Potential Ecological Risk Index (RI), Geoaccumulation index (I_geo), Nenerow Pollution Index (PI_Nemerow), and Enrichment Factor (EF). The formulae to calculate these indices are presented in Tables 9 and 10. Examples of their usage in e-waste literature is presented in Table 11. The PLI is an integrated index to calculate contamination of a number of pollutants; it is widely used to compare pollution in different sites, but omits the availability of metals [180]. The RI is based on the elemental abundance and release capacity data, and was introduced to assess the degree of ecological risks posed by metals in topsoil [181]. The I_geo determines and defines metal contamination in sediments by comparing current levels with pre-industrial levels in order to measure the effects of human activities on the environment. The PI_Nemerow is widely applied to reflect the total pollution level and evaluate environmental quality. Due to its universal formula, the EF is a relatively simple and easy tool for assessing the enrichment degree and comparing the contamination of different environmental media. Toxicity Characteristic Leaching Procedure (TCLP) is widely used to determine the mobility of both organic and inorganic analytes present in liquid, solid, and multiphasic wastes. The method [EPA Method 1311 [182]] was developed by the United States Environmental Protection Agency in order to simulate leaching through landfills. The ecotoxicity impacts can be evaluated using the USEtox® models [183], which simulate the release of chemicals from a source to the environment via mass flows from a succession of homogenous compartments. The use of different pollution indices, each assessing different sets of criteria, requires careful interpretation of the information. For example, the mass concentration of an element may be higher, but in order to access the potential risk, the bioavailability of the metals should also be taken into account.

Over the past few years, there has been growing acknowledgment of the need to include bioavailability in risk assessment frameworks, as protecting an organism from a toxic chemical means that the bioavailability of the chemical for that organism should be known. In this research, it was observed that, of the extensive body of published literature on e-waste’s impact on the health and the environment, relatively few publications measured the bioavailability and/or bioaccessibility of the pollutants (∼1 %). A comprehensive approach taking bioavailability and bioaccessibility of an e-waste pollutant into consideration is needed to help improve the evaluation and to understand its impact on health and the environment.

7 The role of chemistry in tackling the e-waste challenge

The e-waste issue is multi-faceted; in order to tackle this grand challenge successfully, a multidisciplinary approach and a shift in paradigm involving multiple stakeholders is needed. A white paper published by the American Chemical Society (ACS) examined the green chemistry developments in industrial applications and environmental impact reduction across a number of manufacturing and industry sectors [194]. It highlighted how chemistry could underpin a number of potential solutions, from the development of more sustainable and ‘greener’ raw materials (e. g., organic electronics) to the use of novel treatment processes to reduce harmful chemicals and/or recover valuable resources.

Prioritising potentially problematic chemicals substance in EEE manufacturing can be an effective strategy to facilitate the development and use of safer materials in the industry. An example of industry initiative and good practice can be found in a strategy document published in 2018 on ‘A Protocol for Prioritizing Chemicals of Concern in the Electronic Industry’ by Apple Inc. [195]. The document presented the design and application of a protocol to systematically evaluate chemicals of interest to make products and materials safer for workers, customers, recyclers, and the environment.

On the other end of the spectrum, there is a need to support knowledge transfer to developing countries; they are often the most affected by health and environmental issues related to e-waste. In addition to capability building for safer recycling and environmental clean-up, an efficient monitoring and screening of environmental pollutants will also be of benefit. This project presents a case study to examine the concept of the use of certain metals as tracers for the contamination of BFR or other hazardous e-waste compounds that may be beneficial in developing countries, where facilities for organic analysis may be less available.

8 Case study: can metals be used as tracers for organic contaminants in potentially e-waste-polluted environmental media?

The scientific literature reports on the contamination of the ambient environment of e-waste dismantling and recycling facilities (section 4). Typical compounds emitted during e-waste recycling-related activities are toxic BFR (e. g., PBDEs) and metals (including toxic metals, such as cadmium or lead). PBDEs and metals can be detected, e. g., in the dust emitted from e-waste facilities and in soil near such sites.

E-waste recycling facilities are often operated in less developed countries, where the capabilities for environmental monitoring are limited. Environmental monitoring in order to protect the health of people living proximal to such sites is often not implemented.

Since organic (trace) analysis is expensive and sophisticated analytical equipment may not be commonly available in smaller laboratories, the use of tracers for e-waste contaminations as a screening tool would be beneficial. Here it is evaluated if one or several metal(s) could serve as tracer(s) for organic contaminants of e-waste. Analysis of most of the relevant metals is cheap and relatively easy to perform. At the least, a pre-selection of samples may be possible by this means, so that only a few suspicious samples (high content of tracer(s) indicating possible e-waste impact and presence of e-waste-related pollutants) have to be analysed for organic pollutants to assess the potential risk.

For the evaluation several questions are considered and discussed.

9 Conclusions

Rapid advancements in consumer technology are leading to a continual increase in a challenging, heterogenic, and chemically complex e-waste stream that is difficult to manage. East Asia and West Africa are key destinations for e-waste globally and the informal recycling of e-waste provides significant income-generating activities in these countries. Our societies’ current depletion of critical elements in the Periodic Table for use in high tech EEE is unsustainable, which is inadvertently driving backyard, uncontrolled recycling. To recover the valuable resources, e-waste is often dismantled manually, acid stripped, and burnt in an open fire. Metal(loid)s, POPs, and BFRs are frequently detected in many environmental compartments associated with e-waste recycling sites and surrounding areas. Moreover, as new alternative flame retardant compounds replace the banned substances, NBFRs have emerged as contaminants in environmental samples. In general, pollutant levels increase with the following order: control sites < repair sites < dismantling sites < burning sites. Currently, multiple pollution indices or risk models are used to evaluate the ecological impact of e-waste, making comparison of different studies challenging. Nevertheless, it is evident that pollutants related to e-wastes recycling are having an impact in multiple trophic levels (from microbial communities to migratory and resident bird populations and even humans). Future studies should take the bioavailability and bioaccessibility of e-waste pollutants, e. g., in water, soil, or dust, into consideration. Currently, these aspects seem not sufficiently considered in e-waste site risk assessments.

It is noteworthy that countries such as China have introduced policies and legislation to regulate informal e-waste recycling, but many of the POPs (e. g., PCBs and the PBDEs) remain for long periods in the environment and the food chain. To tackle the e-waste challenge, multiple strategies are needed, bringing together environmental, economic, and societal stakeholders (Fig. 5). These measures include: increase investment in the environment remediation of contaminated sites; improve monitoring, evaluation, and reporting of e-waste; develop better technologies for resource recovery and greener and sustainable manufacturing practices; offer incentives to valorise e-waste and promote responsible circular economy; encourage users to ‘reduce, reuse and recycle’ e-waste and increase education and awareness programmes. One strategy can involve building the whole life cost into the original product by applying an appropriate tax, such that its proceeds are directed towards recycling operations. In addition, stronger governance is required to implement policies to prevent pollution of the environment, co-ordinate better e-waste collection, recycling, and reuse; to provide clearer definitions of the responsibility of producers of EEE and its importers; and to promote capacity building in developing countries to safeguard e-waste workers and the environment.

Fig. 5:

Integrated strategies to manage e-waste.

10 Membership of the sponsoring body

Membership of the IUPAC Chemistry and the Environment Division Committee for the 2020 was as follows:

President: Hemda Garelick (UK); Past President: Rai Kookana (Australia); Vice President: Roberto Terzano (Italy); Secretary: Annemieke Farenhorst (Canada); Titular Members: Doo Soo Chung (South Korea), Petr S. Fedotov (Russia), Bradley Miller (USA), Diane Purchase (UK), Fani L. Sakellariadou (Greece), Weiguo Song (China); Associate Members: Nadia Kandile (Egypt), Yong-Chien Ling (Taiwan), Irina Perminova (Russia), Bipul Behari Saha (India), John B. Unsworth (UK), Baoshan Xing (USA); National Representatives: Vladimir Beškoski (Serbia); Cristina Delerue-Matos (Portugal); Michal Galamboš (Slovakia), Joon Ching Juan (Malaysia), Bulent Mertoglu (Turkey), Oluseun Elizabeth Popoola (Nigeria), Yehuda Shevah (Israel), Tiina Sikanen (Findland), Monthip Sriratana (Thailand); Emeritus member: Laura L. McConnell (USA)