Across the Global South, electrical and electronic waste (e -waste) is recovered using rudimentary and often dangerous methods in informal and unregulated facilities. Although these activities provide a valuable contribution to the global circular economy, their uncon trolled nature results in a risk of potentially hazardous substance emission into the environment from where they may pose considerable risk to both occupational and public health. Here, we focus a systematic PRISMA review on two distinct groups of activit ies undertaken in e-waste management in low - and middle -incom e countries (LIMICs): (i ) Physical decon struction and re clamation, involv ing dismantling assemblies of items and materials to recover value; and , (ii) hyd rometallurgical treatment, involving the dissolution and suspension of precious metals using solvents (cyanide ) and acids ( aqua regia ). For comparison purposes, we consolidate information on (i) and (ii) according to the types of substances evidenced ; and identify , critically assess and rank most prevalent hazard -pathway - receptor (H-P-R) risk combinations experienced by people w orking across the Global South. Despite the proliferation of publications, e vidence to assess risk is comparatively limited. Still, we are confident to highlight the extrem ely hazardous nature of work undertaken, often by children, handling highly hazardous substances without protective equipment to reclaim gold and other precious metals using hydrometallurgical processes. Emissions of hazardous substances, particularly pote ntially toxic elements (PTEs) from physical dismantling also represent a serious risk to health . Numerous sources speculatively link concentrations in the environment (a significant risk to children who have a tendency to eat soil) to e-waste dismantling processed. However, many of the sources that identify elevated substance concentrations in environmental media face difficulties in unambiguously and convincingly linking emissions from specific activities to the environmental concen trations , i.e. establishing causality . This key limitation presents us with a challenge for designing and implementing interventions to target, control and replace such highly risky resource recovery methods. Yet, such insufficient information cannot be us ed as an excuse for inaction, especially as our generalised H -P-R inferences here provide for sufficient interlinkages. Keywords: Health and s afety; E-waste; WEEE; Waste; Informal recycling sector; Recycling; Resource recovery; Circular economy; Global South; Risk; Hazardous waste; Hydrometallurgical processing; Potentially toxic elements; Low- and middle -income countries; Poverty alleviation; SDGs; Systematic review. 3 Graphical abstract 4 Abbreviations ADI ‘acceptable daily intake’ As arsenic BDE brominated phenyl ether BFR brominated flame retardants Cd cadmium Co cobalt CP chlorinated paraffins CR Cancer risk Cr chromium CRT cathode ray tube Cu copper dis. dismantling DP Dechlorane Plus DRC dioxins and related compounds EU European Union Fe iron Geog. geographical context Haz. hazard Hg mercury HIC high income countries HQ hazard quotient HS Harmonized Commodity Description and Coding System HSE Health and Safety Executive IARC International Agency for Research on Cancer ICT Information and Communication Technologies In Indium I-TEQ international toxic equivalency factor L likelihood LCCPs long chain paraffins LIC low income countries LIMIC low income and middle income countries LMC lower middle income countries MCCP medium chain paraffins Mn manganese Mo Molybdenum MSW municipal solid waste n number of samples Ni nickel OECD Organisation for Economic Co -operation and Development PAH polycyclic aromatic hydrocarbons Pb lead PBDEs polybrominated diphenyl ethers PCB polychlorinated biphenyls PCDD polychlorinated dibenzo -p-dioxins PCDD/Fs polychlorinated dibenzo -p-dioxins and polychlorinated dibenzo -p-furans PCDF polychlorinated dibenzofurans PM particulate matter PM 10 particulate matter < 10 µm PM 2.5 particulate matter < 2.5 µm POPs persistent organic pollutants PPE personal protective equipment PPP purchasing power parity PRISMA preferred reporting items for systematic reviews and meta -analyses 5 PTEs potentially toxic elements PVC polyvinyl chloride R risk rec. recycling Ref. reference RQ research question S severity Sb antimony SCCP short chain paraffins TBBPA tetrabromobisphenol A TEQ toxic equivalency Tl thallium UMC upper middle income countries US United States USD United States dollars WEEE waste electrical and electronic equipment wt. Weight (i.e. on a weight reporting basis) Zn zinc 6 1. Introduction Electronic and electrical waste (hereafter e -waste) has become a topic of increasing concern in recent decades (Salehabadi, 2013) as stories of its mismanagement and illegal transboundary trade have emerged (Robinson, 2009) . E-waste is not inherently damaging to the environment, however many electrical and electroni c products contain a bewildering array of components, some of which contain potentially hazardous substances and materials (Townsend, 2011) . If appropriate risk measures are not put in place to manage the deconstruction, reuse, recycling, recovery or disposal of ‘after -use’ items and their components, then potentially hazardous substances and mate rials may be at risk of emission into environmental media and/or human receptors (Tsydenova and Bengtsson, 2011) . The treatment and reclamation of valuable components (parts), materials and chemical compounds from e -waste in low income and middle income countries (LIMICs) is often unregulated, and informal (Duan et al., 2013; Tsydenova and Bengtsson, 2011) . Therefore, it is challenging to determine the number of people involved , and the mass of material being processed , because informal and unregulated bu sinesses inherently avoid or have no reason to report their activities (Nnorom and Osibanjo, 2008b) . Once discarded, e -waste flows t hrough society via complex and sometimes geographically diverse pathways, some of which may cross international boundaries as summarised in Figure 1. In the event that e -waste is collected for reclamation, it is likely to be processed via three broad groups of treatment types (Kaya, 2016; Ongondo et al., 2011; Tsydenova and Bengtsson, 2011) : (1) Physical processing: including the manual disassembly or comminution (s ize reduction) to recover components and materials. (2) Hydrometallurgical treatment: whereby leaching solutions (lixiviant) such as nitric acid, hydrochloric acid or cyanide are used to dissolve metals bonded to electrical and electronic assemblies after whic h they are recovered from the solution through a variety of methods including precipitation, electrowinning and solvent extraction. (3) Thermal processing: including the heating or combustion of electrical and electronic assemblies to melt solder or plastic bonding agents and housings to recover components and metals. Each of these processes is controllable with sufficient safeguards in place. H owever , in the absence of regulation, enforcement and a positive health and safety culture (Glendon and Stanton, 2000) , these processes may pose significant risk t o those who are engaged in the 7 activities, resulting in the release of a range of potentially hazardous substances putting at risk of exposure huma n beings and/or environmental media . Figure 1: Material flow system for e -waste management through society . Waste generation Collection for recycling (Formal) Mechanical sorting and disassembly (Formal) Illegal export Open loop mechanical recycling Incineration & gasification Landfill Electrical & electronic product manufacturing Consumer/ commercial use System boundary Closed loop mechanical recycling Recovery and disposal Material production Collection for recovery or disposal (Formal) Diffuse dump Uncollected Collected Collection for recycling (Informal) Transport Chemical recycling Import Dumpsites Dumpsite open burning Atmosphere Dumpsite Collection for recycling (Informal) Manual sorting & disassembly (Informal) Materials recycling Component reuse Reuse Thermal treatment (open burning &smelting) Hydro - metallurgical treatment (acid/alkali washing) 9 In response to prior transgressions, the Basel Convention on the Control of Transboundary Movements of Hazardous Wastes and their Disposal (hereafter the Basel Convention) has prohibited its parties from e xporting broken electrical equipment from Organisation for Economic Co -operation and Development (OECD ) countrie s to non –OECD countries since 2002 (Secretariat of the Basel Convention, 2002) . The objective of this ban, is to prevent high income countries (HIC) outsourcing e -waste t reatment to countries that have limited regulation and enforcement to ensure they are managed responsibly. Whilst this has been effective at curbing the legal trade in e -waste, there are concerns, and mounting evidence that the trade has been pushed underg round (Hopson and Puckett, 2016) and many exporters now exploit a loophole whereby ite ms are exported as functional used goods rather than waste, thereby potentially avoiding the gaze of inspectors (Lee et al., 2018; Lepawsky and McNabb, 2010) . The topic of e -waste is wel l researched in the literature compared to other waste materials such as construction and demolition waste (Cook and Velis, 2020a) or medical waste (Cook et al., 2020c) . Several articles over the last decade have reviewed practices carried out to reclaim and reprocess e -waste , including studies with a scope at global level (Ongondo et al., 2011) , regional level - Africa (Bimir, 2020) and Asia (Herat and Agamuthu, 2012) , and socio economic level - developing countries (Nnorom and Osibanjo, 2008b) . A recent and very detailed study by Kaya (2016) provides the most comprehensive overview of physical and ch emical e -waste reclamation practices, but without any detailed evidence for the potential or realised health effects of the practices described . Other studies have provided more general review s of e-waste policy, flow, and practices alongside assessment of solutions and prospects for l ong-term management of e -waste in various geographical areas including : Asian countries (Herat and Agamuthu, 2012) , Pakistan (Iqbal et al., 2015) , India (Borthakur and Govind, 2018) , China (Lu et al., 2015) , India and China (Awasthi and Li, 2017) , Bangladesh (Bruce -Vanderpuije et al., 2019) , Malaysia (Ismail and Hanafiah, 2019) , Botswana (Mmereki et al., 2015) , Ghana (Daum et al., 2017) , and one global study with a Nigerian focus by Nnorom and Osibanjo (2008a) . There are also many primary data gathering studies relating to specific substances of concern emitted from e -waste processing activities, and some of these have been consolidated for groups such as for neurotoxicants (Chen et al., 2010) , organic flame retardants (Gravel et al., 10 2019a) and tetrabromobisphenol -A (TBBPA) (Malkoske et al., 2016) . More comprehensive reviews have consolidated evidence for substance emission from e -waste activities and described the pathways through which receptors may be exposed. With the exception of a global review by Townsend (2011) , these have had a regional , national or socio -economic category scope , for instance for: China (Xu et al., 2015) , India (Awasthi et al., 2018; Awasthi et al., 2016) , India and China (Brigden et al., 2 005), Ghana (Awere et al., 2020; Brigden et al., 2008; Bruce -Vanderpuije et al., 2019) , and for developing countries (Ackah, 2017) . Three studies have reviewed the impa ct of e -waste on public and occupational safety. Tsydenova and Bengtsson (2011) comprehensively summarised the st ate of knowledge for chemical hazard emissions from e -waste processing activities for ‘developed’ and ‘developing’ countries, providing extensive narrative to describe the link between emissions and processes in each of the two socio -economic groups . Grant et al. (2013) , provided a thorough review of epidemiological studies investigating the association betw een e -waste and a range of health effects including physical health, neurodevelopmental disorders, violence, criminality and education. I n a more recent study , Vaccari et al. (2019) systematically reviewed the impact of e -waste on environmental pollution and human health with a focus on the informal sector, listing and describing the current state of research with a focus on hazardous substance emissions and identification of substan ces in environmental media including a particularly strong focus on potentially toxic elements, with further reference to evidence for environmental concentrations of polybrominated diphenyl ethers (PBDEs ), polychlorinated biphenyls (PCBs ) and dioxin like PCBs , polychlorinated dibenzo -p-dioxins and polychlorinated dibenzo -p-furans ( PCDD/Fs ). Vaccari et al. (2019) was also able to compare emissions of various metals from different e -waste activities such as open burning, hydrometallurgical treatment and physical processing, finding a non -statistical inferred correlation between open burning and metal concentration in soils and between physical processing and higher atmospheric concentrations. Whilst Tsydenova and Bengtsson (2011) , Grant et al. (2013) and Vaccari e t al. (2019) have provided comprehensive contributions the f irst was not a systematic review, and there has been a great deal of new research since publication of Tsydenova and Bengtsson (2011) ; and Grant et al. (2013) , and the la st does not provide detail on hazard exposure or quantified risk. To address this gap following a systematic a nd structured approach we have adapted 11 preferred r eporting items for systematic reviews and meta -analyses (PRISMA ) guidelines to identify, select and review sources that evidence occupational and public health risks associated with the processing of e -waste across global geographies and socio -economic cultures. Initially , we assessed the three broad groups of activities : 1) Physica l processing, 2) Hydrometallurgical treatment and 3) Thermal processing. However, for the sake of clarity and brevity, we have chosen to present here only two of these thematic areas, described as Challenge 1: Physical processing (Section 2.1) and Challenge 2: Hydrometallurgical treatment (Section 4), leaving the thermal treatment for presentation in another publication (Cook et al., 2020b) . For each Challenge, we summarise evidence for hazards emission, followed by sections that present the evidence of quantifie d risk of non -carcinogenic and carcinogenic effects. Whereas our review is global, a strong focus is placed on activities in LIMICs where the most concerning safety challenges are evident. Deliberately, we have excluded from our scope, specific articles th at cover lithium batteries, a new and specialist subject for which there are already two very recent reviews (Asadi Dalini et al., 2020; Siqi et al., 2019) . 2. Methods 2.1. Systematic review This study explored three research questions ( RQ) via a systematic review PRISMA guidelines (Moher et al., 2009) , adapted as detailed by Cook et al. (2020a) :  RQ1: What evidence exists to in dicate risk to public and occupational safety posed by the physical and hydrometallurgical processing of e -waste?  RQ2: What are the comparative risks to public and occupational safety that arise from the physical and hydrometallurgical processing of e -waste?  RQ3: What research could be carried out that would have the greatest impact on harm reduction in the sector that processes e -waste using physical and hydrometallurgical processes ? Scopus, Web of Science and Google Scholar were searched using Boolean search queries that were optimised using one-at-a-time sensitivity analysis to ensure that the maximum number 12 of relevant articles were returned without an overburden of non -relevant art icles (Section S.1.1) . Further supplementary search was carried out using snowballing and citation searching techniques (Cooper et al., 2018) . Literature was screened according to the inclusion and exclusion criteria detailed in Section S.1.2 to achieve the results detailed in Section S.1.3. Further sources were obtained by searching through the websites of international and national organisations such as the United States Environmental Protection Agency (2020) , Occupational Safety and Health Administration (2020) , The World Bank (2020) and World Health Organization (2020) . Risks, hazards, pathways and receptors were identified in literature and combined into scenarios that we re either reported directly or could be conceived to be feasible from the descriptions in each source as described by Cook et al. (2020a) . These combinations enabled the production of conceptual diagrams as shown in Figure 2. 13 Figure 2: Summary of the main sources, pathways and receptors for hazards associated with e-waste (purple arrow delineates crossing arrow and has no further meaning) . Dust Inhalation Adult citizens Occupational (formal) Soil Groundwater Rivers streams & lakes Surface water Sediments Deposition Mouthing Child citizens Ingestion Uptake in crops Food Bonding agents Substances in components Lixiviants (acids, alkalis) Liquid substance Atmosphere Dermal contact Mucous membranes Occupational (informal) Directly discarded Plastic additives Resid. monomers, oligomers, catalysts Leaching activities Ash Released during comminution Released during sorting & dismantling Receptor Route of human exposure Environmental compartment Substance or material Transfer process Legend Sources Pathways Receptors 14 2.1. Risk based approach In order to rank an d compare the relative risks, an approach adapted from Kaya et al. (2018) , Burns et al. (2019) , World Health Organization (2012) , and Hunte r et al. (2003) was used to indicatively score each of the hazard -pathway -receptor combinations on the basis of the likelihood and severity of each occurring using the matrices in Table 1 and Table 2. It is important to note that this method, reported first by Cook et al. ( 2020a) , was not intended to constitute or substitute for a quantitative assessment of risk, but to be used to support decision making by indicat ing the relative harm from each scenario and assist with directing future research a genda. The ranked, aggreg ated results are shown in Section S.2 . Table 1: Matrix used to calculate the relative risk of each hazard -pathway -receptor scenario; after Cook et al. (2020a) . Consequence Very slight Slight Moderate Severe Very severe 1 2 3 4 5 Likelihood Very unlikely 1 1 2 3 4 5 Unlikely 2 2 4 6 8 10 Likely 3 3 6 9 12 15 Very likely 4 4 8 12 16 20 Inevitable 5 5 10 15 20 25 Table 2: Colour coding used to rank hazard potential qualitatively in each category; after Cook et al. (2020a) . Red (R) High harm potential Amber (A) Medium/high harm potential Yellow (Y) Medium/low harm potential Green (G) Low harm potential Grey Insufficient data 3. Challenge 1: Sorting, dismantling and physical processing of e -waste E-waste that has been collected for recycling or refurbishment/reuse is often subject to a multiplicity of processes that are intended to maximise the value from the constituent product components, each of which may contain many subcomponents, substances and materials (Tzoraki et al., 2019) . E-waste has a varying material and chemical composition, which includes a variety of potentially hazardous substances and materials that may negatively affect the health of receptors if a pathway is created. 15 In the following section s, evidence is reviewed for emissions of potentially hazardous substances from the handling, dismantling and physical processing of e -waste. Comparisons are made between the concentrations of s ubstances present in occupational environments in which the formal and informal sector operate, as well as the levels observed in nearby environmental compartments and the blood urine and hair of those who participate in e -waste dismantling activities. The potential hazards posed by e -waste sorting, dismantling and physical processing are simplified in Figure 3. In isolation, these hazards pose no threat to receptors, unless one of the pathways illustrated allows them to encounter them. Figure 3: Hazard exposure conceptual model (source –pathway –receptor) associated with sorting, dismantling and physical processing of e -waste. 3.1. Potentially toxic elements (PTEs) In the context of interaction with biota, metals can be categorised as those which are essential for organisms to function and those which are not (Egorova and Ananikov, 2017) . Popular sentiment incorrectly assumes that the latter group are those which are more likely to result in harmful effects, however metals that are essential for life may also be toxic if the dose is sufficient. In this section, studies are compared that report the concentration of metals and Solder Plastic additives Resid. monomers, oligomers, catalysts Sources Released during comminution Dust Atmosphere Food Inhalation Child citizens Adult citizens Occupational (formal) Mouthing Pathways Receptors Soil Ingestion Groundwater Rivers streams & lakes Surface water Sediments Uptake in crops Receptor Route of human exposure Environmental compartment Substance or material Transfer process Legend Substances in e - components Released during sorting & dismantling Deposition 16 other elements as they have been observed in the human body and several environmental compartments. These eleme nts are typically described in literature and throughout as ‘heavy metals’; a non -specific category and one which excludes metalloid and non -metal elements. Instead, the present study will use the term potentially toxic elements (PTEs), following the recom mendation of Pourret and Hursthouse (2019) . The present research identified 22 studies that provided evidence for PTE emissions from e - waste recycling and treatment activities carried out b y the formal and informal sector . Of these, several provided specific data, or an inference, that the concentration s of PTEs identified in people or environmental compartments originated as a consequence of sorting, dismantling and physical processing. Of course, other activities are carried out by e -waste recyclers which result in emissions of PTEs, such as open burni ng, heating and smelting to recover solder and components as described by Cook et al. (2020b) and hyd rometallurgy to recover metals as described in Section 4. It is therefore challenging to disaggregate emissions from these sources as they often take place in the same locations. Nonetheless, it is important to understand the contribution of specifi c practices to emissions of potentially hazardous substances so that interventions can be implemented efficiently and that the most damaging activities receive attention first. E-waste dismantling in HICs make a useful context for compari son with studies in a LIMIC context, as practices such as open burning are unlikely to take place in HICs and heating and hydrometallurgy are likely to be carried out with engineering control measures in place, such as local exhaust ventilation. Five studi es summarised in Table 3 measured atmospheric PTE concentrations in or around e -waste recycling areas in different contexts, two of which were in HICs. 17 Table 3: Selected elements detected in air samples (μg m-3) close to e -waste dismantling, sorting and physical processing sites. Ref. Geog. Receptor Activity context n Pb Cd Cr Cu Mn Mo In Hg Julander et al. (2014) SWE Formal recycling workers All activities 77f 7i 0.18i 0.45i 2.2i 2.2i 0.050 i 0.018 0.011i Dismantling 34f 8i 0.3g i 0.58i 2.9g i 0.064g 0.26g h 0.012i Indoors 30f 6.8i 0.15i 0.43i 2i 0.050i 0.018 i 0.012i Outdoors 13f 5.2i 0.027i 0.27i 1.3i 0.002i 0.070 i 0.0001i Office workers Office 3f 0.4 0.0053 0.036 0.093 0.11 0.003 0.0004 0.00039 Ceballos et al. (2017) USA Formal recycling workers Shredder sorting (A) 9 1.6-67†*¥ ND-0.84 CRT buffing and grinding (A) 5 9.8-27 0.18-10†¥ CRT operators (A) 6 6.1-16 0.09-0.34 CRT dismantling (A) 8 2.1-5.3 ND-0.84 Baler, battery bulb sorting (A) 5 0.9-3.9 ND-0.065 Shipping, receiving, shredding (C) 12 0.33-3.3 ND-0.08 Battery sorters (C) 13 0.2-3.3 ND-0.065 Dismantling (C) 9 0.37-1.2 ND-0.14 Huang et al. (2016) CHN Informal recyclers Recyclingb 165 0.17 0.006 0.023 0.09 0.076 0.003 Zeng et al. (2016) CHN Informal recyclers Recycling 300a 0.153c 0.006 0.006 0.022 Ref. 170a 0.080c 0.003 0.008 0.023 Tang et al. (2015) CHN Informal recyclers Recycling 15 0.0307d Residents Village 3 kme from recycling area 15 0.0154d Ref. 15 0.0072 Air concentrations benchmarked against UK and US occupational exposure limits ( Section S.4.1 ) except for ACGIH® 2019 TLV® time weighted averages which are proprietary, as follows: ♯ exceeds eight -hour time weighted average ( TWA ) indoor air reference value set by the UK Health and Safety Executive ( HSE); ‡ exceeds eight -hour TWA permissible exposure limit set by US OSHA; † exceeds eight -hour TWA permissible exposure limit set by Californian OSHA; * exceeds 10 -hour TWA recommended exposure limit set by NIOSH (US); ¥ exceed s eight -hour TWA threshold limit value set by the ACGIH® 2019 TLV®; a In PM2.5 air samples; b average of three heights above the ground in e -recycling zone; c significant difference between values using T -test based on Ln -transform (p < 0.000); d signific ant difference between values using Pearson's correlation coefficient (p > 0.05); e village was 3 km distan ce from an area where significant recycling of e -waste was taking place; f inhalable fraction; g dismantling workers significantly more exposed compa red to outdoor workers ( p < 0.05); h dismantling workers significantly more exposed compared to indoor workers ( p < 0.01); i dismantling workers significantly more exposed compared to office workers ( p < 0.05 ). (A) and (C) denote different facilities studied by Ceballos et al. Abbreviations: Geographical context (Geog.) ; number of samples (n); dismantling (dis.); recycling (rec.); cathode ray tube (CRT) ; reference (Ref.) ; not detected (ND) . 18 The first, Julander et al. (2014) , analysed the indoor and outdoor atmosphere at three e -waste recycling facilities in Sweden, comparing the exposure potential between workers car rying out different activities. Although the concentration of the PTEs studied did not exceed guideline limits from any of the UK or US safety institutions, significantly higher concentrations of cadmium ( Cd), copper ( Cu), molybdenum ( Mo) and indium ( In) were experienced by dismantlers in comparison to the workers whose activities were outdoors. In another HIC context, Ceballos et al. (2017) measured Pb and Cd concentrations in air at three e -waste recycling plants in the US, finding comparatively low concentrations in most samples with three exceptions. One out of the nine samples (specific data not shown) collected from workers in the shred ding and sorting area showed a concentration of Pb that exceeded US guidelines (Section Error! Reference source not found. S.4.1) . Further, two out of five samples collected from cathode ray tube (CRT ) buffing and grinding exceeded US guidelines for Cd concentrations in air. It is noteworthy that none of the samples for lead ( Pb) or Cd exceeded the UK guidelines. Occupational exposure limits in the UK a nd other parts of Europe have been shown to be higher for comparable substances compared to the US or Australia for instance Schenk et al. (2008) , reflecting different attitudes and interpretation of evidence in different societal attitudes toward s risk. Whereas it is beyond the scope of the present research to compare guidelines of different national agencies, the fact that the levels identified fell between the two limits and not above, provides an inference that the concentrations were unlikely to result in serious adverse effects. Concentrations of PTEs measured at e -waste dismantling plants in China reported by Huang et al. (2016) , Zeng et al. (2016) and Tang et al. (2015) were all below US and UK limits (Table 3). While this may be unexpe cted, it is suggested here that the level of mechanisation is much greater at the HIC context facilities, increasing the likelihood of particle transmission to the atmosphere. The differences between exposed and reference groups for Pb and mercury (Hg) concentrations measured in air reported by Zeng et al. (2016) , Tang et al. (2015) and Julander et al. (2014) , were all significant and for Julander et al. (2014) , they were significant for all other metals except for In. Whereas the results of the studies by Tang et al. (2015) and Zeng et al. (2016) may be confounded by potential, open burning a ctivities in the sampling area, 19 this is unlikely to be the case for the facilities investigated by J Julander et al. (2014) , providing a justifiable link between elemen tal releases and physical, non -thermal processing. One other study by Fang et al. (2013) reported concentrations of PTEs in mg g-1 of particulate matter (PM) making it non -comparable with the other studies ( Table 4). The difference between concentrations in the two workshops were not compared statistically, however approximately two orders of magnitude more Pb can be observed in the dismantling workshop than the mechanical work shop, approximately 3 –4 times more Cd, twice the concentration of Cr, a similar quantity of Cu and 1,000 times more Ni. Table 4: Element concentrations in air samples at e -waste dismantling workshops in Shanghai, China observed by Fang et al. (2013) . Activ ity location Units n Pb Cd Cr Cu Ni Mechanical workshop mg g-1 PM 10 2x3 12.34 0.108 0.554 27.76 0.472 mg g-1 PM 2.5 2x3 20.46 0.033 1,202 3,753 0.744 Dismantling workshop mg g-1 PM 10 2x3 2,043 0.398 0.436 31.80 0.459 mg g-1 PM 2.5 2x3 6,935 0.094 2,875 1,205 1,148 Abbreviations: number of samples (n) ; particulate matter < 10 µm ( PM 10); particulate matter < 2.5 µm ( PM 2.5). 3.1.1. Soil and dust In addition to measuring atmospheric concentrations of PTEs at dismantling, sorting and physical processing plants, investigations into soil and dust concentrations also provide evidence that indicates the potential hazard exposure. Nine studies reviewed here reported concentrations of 18 PTEs ( Table 5 and Table 6) at formal and informal e -waste processing facilities in seven countries. For context, each reported value has been benchmarked against United States Environmental Protection Agency (USEPA) So il Screening Levels (United States Environmental Protection Agency, 2019b) (Section S.4.2 ). While these levels provide an indication as to the level of contamination, the reported values have not been put into context with background values, which may partly or entirely ex plain the presence of the various PTEs identified. 20 Table 5: Element concentrations observed in soil in and around e -waste dismantling, sorting and physical processing plants mg kg-1). Ref. Geog. Activity context Media n Ag As Cd Co Cr Cu Fe Ga g Hg Fujimori et al. (2012) PHL E-recycling site ( formal) Soil 6 a b 1.5 2.6† 0.3 21¥ 350¥ 31,000¥ E-recycling site ( informal) 7 a b 4 3.2†‡ 2.9 26¥ 810¥ 32,000¥ E-recycl ing site ( formal) Dust 11 a 180¥ 5.1†¥‡ 2.9 53¥♯ 23,000¥♯ 42,000¥ E-recycl ing site ( informal) 6 a 130¥ 7.6†¥‡ 3 17¥ 6,600¥♯ 52,000¥ Isimekhai et al. (2017) NGA Dis. area (D season) Soil 10 c 10.29¥ 36.78†¥‡ 3,165¥ Dis. area (W season) 10 c 8.67¥ 49.6†¥‡ 5,580¥♯ Ackah (2019) GHA Dis. Area Soil 41 b 16 271†¥‡♯ 11¥ 103†¥‡ 11,200¥♯ 56,800¥ 37 d 12 5†¥‡ 8¥ 60†¥‡ 1,800¥ 33,400¥ Ohajinwa et al. (2018) NGA Dis. area Soil 29 b 80¥ 25†¥‡ 2.4 3.4¥ 118†¥‡ 3,199¥ 34,140¥ 12.4 10.2¥♯ Repair area 29 b 1 3.6†¥‡ 2.6 1.5 49†¥‡ 28 11,000¥ 4 0.5 Dis. are a 29 b 19 22†¥‡ 5 15.5¥ 117†¥‡ 3,399¥ 62,896¥ 9 6¥♯ 29 b 15 40†¥‡ 24¥ 1.5 197†¥‡ 7,880¥♯ 70,090¥ 14.5 0.5 Dust 32 38.5 8.4†¥‡ 2.6 7¥ 36†¥‡ 2,062¥ 30,620¥ 8.6 3.3¥ 32 40¥ 71†¥‡ 30¥ 1.5 188†¥‡ 13,580¥♯ 97,260¥♯ 15 0.5 Dis. area (control soil) Floor dust n/a 34.5 5.7†¥‡ 2.55 1.5 78†¥‡ 766¥ 7.8 2.3¥ Dis. sites Roadside dust n/a 7.7 13†¥‡ 2.55 2.8¥ 62.5†¥‡ 234 5.8 0.18 Dis. sites Direct electronic dust n/a 1.35 1.2† 2.55 72.3¥♯ 14.3†¥‡ 42 21 0.5 Dis. area (control soil) Floor dust n/a 40¥ 71†¥‡♯ 30¥ 1.5 188†¥‡ 13,580¥♯ 15 0.5 Tang et al. (2015) CHN E-recycling site Soil 15 e 3.1¥ Village 3 km away 15 e 1.3¥ Ref. area 15 e 0.1 E-recycl ing site Dust 15 37.6¥♯ Damrongsiri et al. (2016) THA Adjacent residential Soil 4 f 58-143 16,285 -19,124¥ Dis. area 11 f 214-12,986¥♯ 26,968 -102,580¥♯ Chakraborty et al. (2019) IND Dis. w’shop Soil 5 22†¥‡ 3 7¥ 63†¥‡ 273 3¥ Sorting w’shop 4 0.2 1 3¥ 110†¥‡ 533¥ 1 Tzoraki et al. (2019) GRC Around the plant Soil 25 0.057 -0.721 6.7-21.7†¥‡ 0.13-2.42 4.2-15.4¥ 21.9-114.4†¥‡ 20.62 -254.23 0.9-4.2 0.033 -0.238 Fang et al. (2013) CHN Floor dust mech. w’shop Dust n/a 92¥ 174†¥‡ 947¥ Floor dust dis. w’shop n/a 59¥ 152†¥‡ 2,160¥ a mg/kg air -dry-base; b depth 030 cm; c depth 0 -10 cm; d depth 30 -100 cm; e depth 5 -10 cm; f depth 10 cm; g no guidelines were published by USEPA and hence not benchmarked; † exceeds USEPA carcinogenic screening level (TR=1E -06) for residential soils; ¥ exceeds USEPA child non -carcinogenic screening level (THI=0.1) for residential soils (Section S.4.2 ); ‡ exceeds USEPA carcinogenic screening level (TR=1E - 06) for industrial soils; ♯ exceeds USEPA non -carcinogenic screening level (THI=0.1) for industrial soils; specific screening level elements were compared to: Cd to ‘cadmium (Diet)’; Cr to ‘chromium (VI)’; Hg to ‘elemental mercury’ . Abbreviations: Geographical context (Geog.); number of samples (n) ; dismantling (dis.) ; mechanical (mech. ); not available (n/a) 21 Table 6: Element concentrations observed in soil in and around e -waste dismantling, sorting and physical processing plants (mg kg -1). Ref. Geog. Activity context Media n In Mn Mo Ni Pb Sb Se V Zn Fujimori et al. (2012) PHL E-recycling site ( formal) Soil 6 a b 1 800† 16 53 140 E-recycling site ( informal) 7 a b <0.5 900† 64 650¥ 1,000 E-recycling site ( formal) Dust 11 a 19 820† 2,600¥♯ 9,000¥♯ 3,000¥ E-recycling site ( informal) 6 a <0.5 2,100† 200¥ 1,400¥♯ 2,800¥ Isimekhai et al. (2017) NGA Dis. area (D season) Soil 10 c 254.9† 77.4 911¥♯ 22.5¥ 862.6 Dis. area (W season) 10 c 120.8 23.91 1,823¥♯ 58.4¥♯ 1,921 Ackah (2019) GHA Dis. area Soil 41 b 2,380¥♯ 5 1,820 37 d 846¥♯ 3 866 Ohajinwa et al. (2018) NGA Dis. area Soil 29 b 672† 5.6 103¥ 1,723¥♯ 523¥♯ 2.4 51¥ 2,534¥ Repair area 29 b 81 1.5 100¥ 12 1.5 0.3 10 25 Dis. area 29 b 859† 6.5 95¥ 11,757¥♯ 78¥♯ 14 52¥ 2,724¥ 29 b 540† 17 153¥ 2,271¥♯ 204¥♯ 12 46¥ 5,650¥ Dust 32 17,094†♯ 2.1 131¥ 370 84.5¥♯ 0.7 36 1,616 32 20,265†♯ 18 149¥ 3,770¥♯ 382¥♯ 4 26 5,401¥ Dis. area (control soil) Floor dust n/a 680† 0.5 99.4¥ 277 43¥ 0.65 1,342 Dis. sites Roadside dust n/a 426† 0.5 83 374 14¥ 0.6 924 Dis. sites Direct electronic dust n/a 1,607† 1.3 320¥ 0.5 1.5 0.25 122 Dis. area (control soil) Floor dust n/a 654† 17.5 149¥ 3,770¥♯ 382¥♯ 4 5,401¥ Damrongsiri et al. (2016) THA Adjacent residential Soil 4 f 350-405† 12-23 40-92 158-316 Dis. area 11 f 466-674† 16-183¥ 86-4,556¥♯ 182-4,258¥ Chakraborty et al. ( 2019) IND Dis. w’shop Soil 5 75 197 Sorting w’shop 4 234¥ 87 Tzoraki et al. (2019) GRC Around the plant Soil 25 <0.02 -0.03 201-972† 0.27-1.72 14.2-99.7¥ <0.010 -0.083 0.17-45.42¥ <0.1-0.5 2-55¥ 30.9-1,089.2 Fang et al. (2013) CHN Floor dust mech. w’shop Dust n/a 1,225,000†¥‡♯ 13,880¥♯ Floor dust dis. w’shop n/a 318¥ 17,830¥♯ a mg/kg air -dry-base; b depth 0 -30 cm; c depth 0 -10 cm; d depth 30 -100 cm; e depth 5 -10 cm; f depth 10 cm; † exceeds USEPA carcinogenic screening level (TR=1E -06) for residential soils (Section S.4.2 ); ¥ exceeds USEPA child non -carcinogenic screening level (THI=0.1) for residential soils; ‡ exceeds USEPA carcinogenic screening level (TR=1E -06) for industrial soils; ♯ exceeds USEPA non -carcinogenic screening level (T HI=0.1) for industrial soils; specific screening levels elements were compared to: Mn as Manganese ‘non -dietary’; Ni as ‘Nickel oxide’; Sb as metallic antimony . Abbreviations: Geographical context (Geog.); number of samples (n) ; dismantling (dis.) ; mechani cal (mech) ; not available (n/a) . 22 Many of the values reported exceeded Soil Screening Levels set for residential soils which are set to the lowest common denominator that assumes that children will ingest soil. While several sources review ed provide eviden ce that children live on e -waste sites, it is possibly more relevant to consider the concentrations that exceed the levels set for industrial premises which are deliberately set higher as it is considered that workers are less exposed to soils in compariso n to a domestic context. Two elements, As and Cr , exceeded carcinogenic industrial soil screening level at 16 and 18 sites respectively. Pb was also prevalent, exceeding the non -carcinogeni c industrial level at 14 sites. Two sites reported by Fang et al. (2013) showed very high leve ls of nickel ( Ni) and Pb in dust collected from the mechanical workshop floors of e -waste dismantling plants in Shanghai, China. Levels of Ni here were more than 1,000 times the non -carcinogenic industrial soil screening level set by USEPA and levels of Pb were 17 and 22 times the level. Other high levels of Pb were reported by Fujimori et al. (201 2) who observed levels 11 times the non -carcinogenic industrial soil screening level set by USEPA in formal e -waste recycling sites in the Philippines where levels were higher than for the informal sector sites investigated. 3.1.2. Water PTEs have also been observed in drinking and environmental freshwater in and around e - waste dismantling facilities ( Table 7). Tang et al. (2015) observed Hg and Zhang et al. (2019) observed Cd, Cu, Pb and zinc ( Zn) all at concentrations below USEPA guidelines for tap water, freshwater aquatic life and freshwater human consumption (Section S.4.3 ). Tzoraki et al. (2019) analysed well water, stream water and water from a storage tank at an e - waste dismantling site, finding values generally below thresholds for drinking water safety. A notable exception was one sample which exceeded the limit for arsenic ( As), possibly because of local volcanic activity rather than the e -waste dismantling activities. Several other samples exceeded limits for drinking water by a small margin for some elements, notably Cd, cobalt ( Co), chromium ( Cr), Cu, Pb, manganese ( Mn), antimony ( Sb), and thallium ( Tl). Chronic freshwater (aquatic life) limits were exceeded for Pb in 50% of cases, with one exceeding by 13 times, and Cd in just one case out of eight by a small margin. The paper found a statistically significant correlation between the compa ny’s activities and the concentrations on the basis of elemental composition of the processed feedstock, indicating a clear link between its activities and local pollution of the soil and water. 23 Table 7: Elemental concentrations ob served in water in and around e -waste dismantling sites (µg L-1). Analyte Tang et al. (2015) Zhang et al. (2019) Tzoraki et al. (2019) Residents in e -waste recycling (CHN) Residents in dismantling area (CHN) Ref. area (CHN) Residents and workers (GRE) Drinking water (n=15) Drinking water (n=25) Drinking water (n=3) Streams (n=6) Storage tank (n=1) Well (n=1) Ag <0.05 -0.11 Al 24-717 As 2.9-12.2 ¥†*₸ Be <0.05 -0.12 Cd 0.048 <0.01 <0.05 -1.02 ¥♯ Co 0.09-1.23 ¥ Cr 0.07-10.4 ¥† Cu 66.5 65.3 3.5-81.1 ¥ Ga <0.05 -0.19 Hg 0.014 - Mn 3.97-177 ¥ Mo 0.3-6 Ni 0.2-19.2 Pb 1.91 1.5 1.1-43.5 ¥♯ Pt <0.01 -0.01 Sb 0.15-2.91 ¥ Se <0.5-0.8 Tl <0.01 -0.03 ¥ W 0.06-0.26 Zn 80.7 75.1 4.9-328.2 † Exceeds USEPA carcinogenic screening level (TR=1E -06) for residential drinking water (Section S.4.3 ); ¥ exceeds USEPA child non -carcinogenic screening level (THI=0.1) for residential drinking water; ‡ exceeds USEPA acute quality criteria for freshwater aquatic life; ♯ exceeds USEPA chronic quality criteria for freshwater aquatic life; * exceeds USEPA quality criteria for freshwater consumption of water and organisms; ₸ exceeds USEPA quality criteria for freshwater consumption of organisms only. Abbreviations: number of samples (n) ; reference (Ref.) 3.1.3. Food Two authors also reported concentrations of PTEs in foodstuffs near to e -waste dismantling areas in China ( Table 8). The analysis by Tang et al. (2015) tested samples for Hg finding that they exceeded thresholds set by the Chinese National Standard Agency in just a few cases. Limits are commonly provided for Hg in fish, which is thought to be a significant source of human intake. Therefore few guidelines are available to compare levels with other foodstuffs; however, none of the mean concentrations exceeded the threshold for fish set in European Commission Regulation 1881 /2006 (European Union, 2006; European Union, 2015) (Section S.4.4 ). Zhang et al. (2019) analysed sam ples for Pb, Cd, Cu and Zn, finding levels of Pb that equalled or exceeded by a small margin the limits stated in Commission Regulation 1881/2006, with higher concentrations detected in e -waste dismantling areas compared to the reference samples. Similarly , levels of Cd were notably higher in the e -waste dismantling sites , however none exceeded Commission Regulation thresholds. 24 Table 8: Elemental concentrations observed in food and crops (µg g-1 foodstuff) around e - waste dismantling sites. Ref. Geog. Activity context Food n Pb Cd Cu* Zn* Hg Tang et al. (2015) CHN Village 3 km away Rice* 15 0.0505 Corn* 15 0.068 Soybean* 15 0.043 Cole* 15 0.025 Grain* 15 0.056 Rice (ref)* 0.012 Chicken & pork* 15 0.0601 Fish 15 0.2298 Milk* 15 0.0029 Food oil* 15 0.0054 Table salt* 15 0.0045 Zhang et al. (2019) CHN Residents living e -waste dismantling area Vegetables 46 0.35† 0.096 1.37 3.89 Residents ref. area Vegetables 4 0.1† 0.023 1.75 6.23 Residents living e -waste dismantling area Rice 32 0.2† 0.15 12.3 27.7 Residents ref. area Rice 5 0.18 0.041 9.38 24.6 Residents living e -waste dismantling area Egg* 28 0.071 0.006 1.01 16.2 Residents ref. area Egg* 10 0.014 0.0002 0.63 12.2 †Exceeds limits on concertation as per Commission Regulation 1881/2006 (European Union, 2006; European Union, 2015) (Section S.4.4 ); *no threshold comparison stated in Commis sion Regulation 1881/2006. Abbreviations: Geographical context (Geog.); number of samples (n) ; reference (ref.) . 3.1.4. Human exposure Whereas elemental concentrations in various environmental compartments are an indicator of environmental emissions, observations of these substances in blood, urine and hair evidence exposure and absorption, which can occur through a variety of mechanisms (exposure pathways). Many PTEs which have not accumulated in the body are excreted through urine, indicating exposure over recen t hours or days (Zhang et al., 2019) . Two studies, Zhang et al. (2019) and Julander et al. (2014) reported concentrations of elemental mass per volume of urine for e -waste dism antlers in China and Sweden respectively. In Table 9, the results of these studies are shown alongside concentrations of non -occupationally exposed UK residents’ urine for comparison; though these should be treated with caution as dietary intake and environmental conditions can significantly influence concentrations of some elements (Asante et al., 2012) . 25 Table 9: Element concentrations observed in urine of people working and living around e - waste dismantling facilities (μg L-1 urine). Element Zhang et al. (2019) (CHN) Julander et al. (2014) (SWE) Morton et al. (2014) (GBR) Residents near e - waste (n=139) Residents ref. area (n=26) E-recycling workers (n=52) Office workers (n=10) Non-occupationally exposed adults All (n=132) Al 17 19.1 3.82 As 46.6 62 13 19 10.48 Cd 2.12* 1.33 0.37 0.27 0.13 Co 0.44 0.32 0.25 0.24 0.22 Cr 0.74 0.71 0.35 Cu 22.2 16.9 8.75 Hg 0.5 0.42 1.4 0.66 0.43 In 0.0068 0.0047 <LOQ (0.031) Mn 0.77 0.61 <LOQ (0.092) Pb 4.98*** 1.23 1.8** 0.66 0.47 Sb 0.2 0.11 <LOQ (0.092) Se 30.5 26.3 13.4 Tl 0.57 0.59 0.17 Zn 530 493 n/a *Significantly higher ( p < 0.05) in e -waste recycling area compared to reference (Mann –Whitney U -test); ** significantly higher ( p < 0.01) in e -waste recycling area compared to reference (Mann –Whitney U -test); *** significantly higher ( p < 0.001) in e -waste recycling area compared to reference (Mann –Whitney U -test). Abbreviations: Number of samples (n) ; reference (ref.); limit of quantification (LOQ) ; not available (n/a) . Urinary concentrations of most elements were higher in the study of exposed and non - exposed Chinese residents compared to those in the UK or Sweden. An interesting exception is the concentrations of Hg in the uri ne of both groups of Chinese residents that were similar to UK non -exposed residents. In contrast, the levels of Hg in e -waste workers in Sweden were almost double in the e -waste workers and slightly higher in the office workers at the e - waste recycling fa cility. In the case of Cd and Pb, levels were significantly higher for e -waste workers in China compared to the reference group ( p < 0.05 for Cd; p < 0.001 for Pb), clearly indicating occupational exposure to these elements. A similar correlation between e-waste dismantling and recycling activities and exposure to PTEs is indicated by three studies that analysed the blood of people working in the sector (Table 10). For instance , Julander et al. (2014) observed concentrations of Cr and Pb to be higher in occupationally exposed workers compared to office workers in the s ame premises with a significance of p < 0.05 and p < 0.01 respectively. 26 Table 10: Blood concentrations of selected elements compared for adults working near and in e-waste recycling areas (μg L-1). Ref. Geog. Activity context n Cd Co Cr Hg In Pb Julander et al. (2014) SWE E-recycling workers 50 0.081 1.4†* 1.4 0.0057 32** 48a 0.81†* 0.0043* Office workers 10 0.064 1.1† 1.2 0.0030 15 10a 0.30 0.0023 Ceballos et al. (2017) USA Baler, shredde r, battery (A) 5 ND-12† ND-88† CRT shredders (B) 7 ND-17† ND-46† Dismantling (B) 13 ND-9† ND-13 Shipping, receiving, shredding (C) 5 5-10† ND Dismantling (C) 3 6-15† ND Battery sorters (C) 2 5-6† ND Amankwaa et al. (2017) GHA E-waste dismantlers 28 36.3† a As concentration in plasma/serum fraction; † exceeds reference value for blood on non -occupationally exposed Canadian adult citizens (20 to 79 years of age) (Section S.4.5 ); ♯ exceeds reference value for blood on non -occupationally exposed Canadian citizens (3 to 79 years of age); * significantly higher ( p < 0.05) in e -waste recycling area compared to reference (Mann –Whitney U -test); ** significantly higher ( p < 0.01) in e -waste recycling area compared to reference (Mann –Whitney U-test). Abbreviations: Geographical context (Geog.); number of samples (n); ca thode ray tube (CRT); Not detected (ND). Ceballos et al. (2017) did not analyse blood from the control group or statistically compare variance between the workers, however comparison with average concentrations in the blood of Canadian adults between 20 and 79 years old ( Section S.4.5 ) shows lev els of Cd were six to twenty times higher in workers where it was detected. The highest recorded blood Cd levels were in the CRT and shredding workers who also showed higher Pb concentrations. Ceballos et al. suggest s that the shredder is the cause, despit e operating only three days per month. In an interesting finding, Julander et al. (2014) reported concentration of In approximately twice as high in the serum of e -waste recycling workers in comparison to the office workers in the same building, ( p < 0.05) ( Table 10). In is used increasingly in electronics; mainly as indium -tin oxide in flat -screens. According to Julander et al. (2014) , information on toxicity and carcinogenicity is scant and no reference concentration was available in the study of the Canadian population. The study cautions that with the increase in the number of flat -screens entering the after -use phase, that monitoring exposure in recycling workers in the context of their health is increasingly important. Many workers in LIMICs, particularly those who are self -employed, run businesses from home and e -waste dismantlers are no exception according to the present study. Consequently, children may be exposed to potentially hazardous substances that are emitted d uring recovery of valuable materials whether the children are occupationally involved or otherwise. Five studies of children and one of pre -school children living in e -waste dismantling sites 27 measured blood concentrations of elements and compared them to n on-exposed children (Table 11). Levels of Pb were significantly higher in exposed children compared to the control groups in all six studies, providing a strong indication that e -waste recycling operations were the cause. Compared to the Canadian child population (Saravanabhavan et al., 2 017) (detailed in Section S.4.5 ), blood Pb levels were more than double for pre -school children and between double and triple the level for all children in the other five studies. The United States Centers for Disease Control and Prevention stipulates a relative safety Pb concentration of 50 μg L-1 (Centers for Disease Control and Prevention, 2019) which it states is far below the 450 μg L-1 after which chelation therapy is required. Four of the studies reported values greater than 50 μg L-1, one of which, Zhang et al. (2017) , reported an average of more than double. Table 11: Blood concentrations of selected elements compared for children living in e -waste recycling sites in China compared with reference groups (μg L-1). Ref. Receptors Activity context n Cd Cr Mn Pb Cai et al. (2019) Preschool children Living e -recycling area 358 48.8¥*** Living non -e-recycling area 216 34.7¥ Liu et al. (2018) Children Living e -recycling site 146 49.4‡*** Living non -e-recycling site 88 38.5‡ Zhang et al. (2017) Children Living e -recycling site 153 8.3‡ 103.4‡*** Living non-e-recycling site 141 1.79‡ 23.9‡ Zheng et al. (2019) Children Living e -recycling site 104 72.3‡*** Living non -e-recycling site 96 39.1‡ Zeng et al. (2017) Children Living e -recycling site 100 3.57‡ 55.3‡*** Living non -e-recycling site 106 0.57‡ 5.8 Zeng et al. (2016) Children Living e -recycling site 300 0.576‡*** 7.65† 28.18 62.4‡*** Living non -e-recycling site 170 0.5‡ 7.49† 20.09 47.5‡ The following reference values are applied as appropriate to the age group ( Section S.4.5 ): ¥exceeds reference value for blood on non -occupationally exposed Canadian child citizens ( three to five years of age); ‡exceeds reference value for blood on non -occupationally exposed Canadian child citizens ( six to nineteen years of age); *** significantly ( p<0.001) higher in e - waste recycling area compared to reference using independ ent sample t -test. or Mann -Whitney U test in the case of Zheng et al. (2019) or two sample T -Test in the case of Cai et al. (2019) . Abbreviat ions: Number of samples (n). Levels of blood Cd were tested in three of the six studies, and were found to be between two and thirteen times higher than the Canadian average (0.27 for six to nineteen year olds (Saravanabhavan et al., 2017) ) in both the reference groups and the exposed groups . However , a significant difference between the exposed and control groups was only found by Zeng et a l. (2016) . Cd is widely used in electronic g oods in batteries and printed circuit boards and is considered to be highly hazardous with reported effects including teratogenicity; endocrine and reproductive toxicities; nephrotoxicity, immunotoxicity , neurotoxicity, and cardiovascular disease (Zhang et al., 2017) . Clearly, any indication that children are being ex posed to elevated levels of Cd should be treated with concern. 28 3.1.5. Non-carcinogenic risk To interpret concentrations of PTEs measured in human and environmental media, human health risk assessment models have been developed to evaluate the adverse health effects in receptors who may be occupationally or environmentally exposed via different pathways. Non-carcinogenic and carcinogenic effects are reported separately by calculating the average daily intake for a specific pathway/receptor combination; requiri ng the following parameters: (1) Exposure frequency (day/year), exposure duration (year), the body weight and the average daily exposure time are common factors and are used for all exposure routes (ingestion, inhalation and dermal exposure); (2) Ingestion rate an d concentration of PTEs in soil, dust and/or food needed for the calculation of average daily intake via ingestion; (3) Inhalation rate and concentration of PTEs in air needed for the calculation of ‘acceptable daily intake’ ( ADI) via inhalation; (4) Exposed ski n surface area, skin adherence factor, dermal absorption factor and concentration of PTEs in soil, dust and/or air needed for the calculation of average daily intake via dermal exposure. Recommended values for these exposure factors and further methodolog y description is provided by the United States Environmental Protection Agency (2001) . Non-cancer risk effects are expressed using either a ‘hazard index’, a ‘hazard quotient’ or both. The hazard quotients are obtained by calculating the ratio of the average daily intake to the corresponding reference dose. If the ratio is greater than one, then the receptor is likely to experience adverse health effects; if i t is greater than 10, the risk is considered high and the receptor may experience chronic health effects (Fang et al., 2013; Fujimori et al., 2012) . When exposure to more than one substance is being assessed, the hazard quotients can be summed to calculate a hazard index, thus indicating the risk from multiple hazards (Ohajinwa et al., 2019a) . As with the hazard quotients, if the hazard index is greater than one, the potential for non -cancer risk is significant, if it is less than one, then the non-cancer risk is considered non -significant (Cao et al., 2020) . Two studies , Fang et al. (2013) and Huang et al. (2016) determined the non -carcinogenic risk through air inhalation for different receptors ( Table 12 and Table 13). Specifically, Huang et al. (2016) estimated the hazard index for non -occupationally exposed receptors (children and adults), while Fang et al. (2013) focused on occ upationally exposed receptors. Instinctively, occupational exposure may be expected to result in higher risks compared to non - occupational environments. However the results showed higher non -cancer risk effects for children and residents living near e -recycling sites compared to the workers. This unexpected 29 discrepancy may be explained by differences in the two study approaches. Fang et al. (2013) measured the concentration of PTEs in ambient PM 2.5 air samples whereas Huang et al. (2016) measured whole air exposure from buildings very close to e -waste recycling workshops . Fang et al. (2013) also tested far fewer element types compared to Huang et al. (2016) with five and thirteen elements investigated respectively. 30 Table 12: Non-carcinogenic hazard index for several pathway/receptor combinations for receptors expos ed to potentially toxic elements ( PTEs ) from e-waste dismantling activities. Pathway Media Ref. Geog. Media description Activity context Receptor n Hazard index PTEs of concern (HQ>1) PTEs included Inhalation Air Huang et al. (2016) CHN From three heights above the ground E-recycling site Adults 165 2.7† - Pb, Cd, Cr, Cu, Fe, Mn, Ni, Zn, As, Sb, V, Co, Mo Children 165 8† Ni Fang et al. (2013) CHN PM 2.5 air samples Mech. sorting area Workers 2 <1 - Pb, Cd, Cr, Cu, Ni Dis. area Workers 2 <1 - Ingestion Soil Ackah (2019) GHA Topsoil (d=0 -30 cm) Dis. area Adults 41 1.9† Pb Pb, Cd, Cr, Cu, Fe, Zn, Ag, As, Se Children 41 14.9† Pb, Cu, Fe Subsoil (d=30 -100 cm) Dis. area Adults 37 <1 - Children 37 4.9† Pb Fujimori et al. (2012) PHL Topsoil (d=0 -30 cm) E-recycling site ( formal) Adults 6 <1 - Pb, Cd, Cu , Mn, Ni, Zn, Ag, Co Children 6 <1 - E-recycling site ( informal) Adults 7 <1 - Children 7 3.3† - Ohajinwa et al. (2019a) NGA Topsoil (d=0 -10 cm) Dis. Area * Workers n/a <1 - Pb, Cd, Cr, Cu, Fe, Mn, Ni, Zn, Ag, As, Sb, V, Co, Mo, Ga, Hg, Se Repair area Workers n/a <1 - Ingestion Dust Fujimori et al. (2012) PHL Surface dust E-recycling site ( formal) Adults 11 4.6† Pb Pb, Cd, Cu, Mn, Ni, Zn, As, Ag, Co, In Children 11 37† Pb E-recycling site ( informal) Adults 6 1.4† Pb Children 6 11† Pb Fang et al. (2013) CHN Floor dust Mech. sorting area Workers n/a 7.28† Pb Dis. area Workers n/a 2.81† Pb Pb, Cd, Cr, Cu, Ni Zhang et al. (2019) CHN Indoor and outdoor dust E-recycling site Children (2 -7 yrs) 21 2.02† Pb Pb, Cd, Cu, Zn Children (8 -19 yrs) 30 <1† - Ohajinwa et al. (2019a) NGA Floor dust (control soil) Dis. area (control soil)** Workers n/a <1 - Pb, Cd, Cr, Cu, Fe, Mn, Ni, Zn, Ag, As, Sb, V, Co, Mo, Ga, Hg, Se Repair area (control soil)** Workers n/a <1 - Roadside dust Dis. area Workers n/a <1 - Repair area Workers n/a 1† - Direct dust from e -waste Dis. area Workers n/a <1 - Repair area Workers n/a 1.5† - * From three different e -recycling sites; ** from two different e -recycling sites. Abbreviations: Geographical context (Geog.); number of samples (n); particulate matter < 2.5 µm (PM 2.5); hazard quotient ( HQ); potentially toxic elements (PTEs) ; dismantling (dis.); mechanical (mech.) ; not available (n/a) . 31 Table 13: Non-carcinogenic hazard index for several pathway/receptor combina tions for receptors exposed to potentially toxic elements ( PTEs ) from e-waste dismantling activities. Pathway Media Ref. Geog. Media description Activity context Receptor n Hazard index PTEs of concern (HQ>1) PTEs included Dermal exposure Soil Ohajinwa et al. (2019a) NGA Topsoil (d=0 -10 cm) Dis. area * Workers n/a 2,300† – 4,000† Cd, Cr, Cu, Fe, Mn, Ni, Zn, Ag, As, Sb, V, Hg Pb, Cd, Cr, Cu, Fe, Mn, Ni, Zn, Ag, As, Sb, V, Co, Mo, Ga, Hg, Se Repair area Workers n/a 1,500† Cr, Fe, Ni, Ag, As, Sb, V, Hg Ackah (2019) GHA Topsoil (d=0 -30 cm) Dis. area Adults 41 <1 - Pb, Cd, Cr, Cu, Fe, Zn, Ag, As, Se Children 41 <1 - Subsoil (d=30 -100 cm) Dis. area Adults 37 <1 - Children 37 <1 - Dermal exposure Dust Ohajinwa et al. (2019a) NGA Floor dust (control soil) Dis. area (control soil) ** Workers n/a 1,700†- 5,500† Cd, Cr, Cu, Fe, Mn, Ni, Zn, Ag, As, Sb, V, Hg Pb, Cd, Cr, Cu, Fe, Mn, Ni, Zn, Ag, As, Sb, V, Co, Mo, Ga, Hg, Se Repair area (control soil) * Workers n/a 1,400†- 2,300† Cr, Cu, Fe, Mn, Ni, Ag, Sb, V, Hg Roadside dust Dis. area Workers n/a 2,000† Cr, Cu, Fe, Mn, Ni, Zn, Ag, As, Sb, V, Hg Dis. area Workers n/a 5,400† Cr, Fe, Mn, Ni, Ag, V, Hg, Co Direct dust from electronics Repair area Workers n/a 4,000† Cr, Cu, Fe, Mn, Zn, As, Sb, V, Hg Roadside dust Repair area Workers n/a 1,800† Cr, Fe, Mn, Ni, Ag, Sb, V, Hg Dermal exposure Air Fang et al. (2013) CHN PM 2.5 air samples Mech. sorting area Workers n/a <1 - Pb, Cd, Cr, Cu, Ni Dis. area Workers n/a <1 - * From three different e -recycling sites; ** From two different e -recycling sites . Abbreviations: Geographical context (Geog.); number of samples (n); particulate matter < 2.5 µm (PM 2.5); hazard quotient ( HQ); potentially t oxic elements (PTEs) ; dismantling (dis.); mechanical (mech.) ; not available (n/a) . 32 Five studies calculated the hazard quotients as a result of soil and dust ingestion ( Table 12 and Table 13). The pathway for exposure to soil and dust is through deposition on food, hand -to-mouth activity and/or direct soil consumption, usually by children (Ohajinwa et al., 2019a) . Two studies reported the non -cancer risk induced by soil ingestion for residents living in e -recycling sites indicating significant non -cancer risk for children and low risk for adults living near informal e -recycling areas (Ackah, 2019; Fujimori et al., 2012) . For example, Ackah (2019) reported that hazard quotient of Pb was 7.5 times greater for children than for adults indicating the exposure to Pb via soil ingestion was the greatest contributor to non-cancer risk effect for children. In addition, Ohajinwa et al. (2019a) reported the hazard index through soil ingestion for workers in informal e -recycling as inconsiderable. These findings indicate that the specific exposure pathway (soil ingestion) is negligible for adults, including occupationally and non -occupationally receptors, but it is considerable for children due to their soil -pica behaviour and low bo dy weight leading to higher soil ingestion rates (Fujimori et al., 2012; United States Environmental Protection Agency, 2017) . Zhang et al. (2019) reported a high hazard quotient from ingestion of Pb through dust for preschool children. These results are similar to Fujimori et al. (2012) who calculated the hazard index and quotients different PTEs caused by dust ingestion for adults and children living nearby formal and informal e -recycling operations. Interestingly, the hazard index was higher at formal e -recycling areas than informal sites attributed to the accumulation of PTEs in indoor surface dust of formal e -recycling sites, absence of any natural dilution effect such as rain and wind, and higher concentration of specific elements (for example, Pb, Cu, Ni and In) (Fujimori et al., 2012) . In addition, Fang et al. (2013) and Ohajinwa et al. (20 19a) determined the non -cancer risk for occupationally exposed receptors through dust ingestion in and around different e -recycling workshops related to dismantling. Fang et al. (2013) found high hazard index for workers in mechanical sorting and dismantling workshops caused by d ust ingestion with Pb being the greatest contributor in both workshops (80% to 92% of hazard index). Ohajinwa et al . (2019a) found negligible non -cancer risks around the e-recycling workshops, except for workers in the repair area where the hazard index was higher than one due to dust ingestion directly from e -waste. Substances present in soil and dust can adhere t o exposed skin leading to dermal absorption and therefore dermal exposure to PTEs via soil and dust can occur (Ohajinwa et al., 2 019a) . However, limited information exists making it difficult to compare and assess the non -cancer 33 risk for this exposure pathway. Ackah (2019) found that dermal exposure induced negligible risk for children and adults living nearby e -recycling sites. However, Ohajinwa et al. (2019a) reported that dermal contact with soil and dust were the main pathways of exposure to PTEs for workers finding extremely high hazard index’ , which were attribute d to the absence of PPE for the majority of workers (82% of total workers) (Ohajinwa et al., 2019a) . The difference between Ackah (2019) and Ohajinwa et al. (2019a) for dermal exposure via soil is related to the selection of different dermal absorption factors, but still further scrutiny is required to obtain confide nce about the contribution of dermal exposure to non -cancer risk effects. Food consumption is also a pathway of exposure via ingestion, reported here as ‘dietary intake’ to distinguish from incidental dust and soil ingestion. The exposure to PTEs through dietary intake for different age group populations nearby e -recycling sites was investigated by Zhang et al. (2019) (Table 14), indicating that the co nsumption of vegetables and rice is likely to cause adverse non -cancer health effects. Hazard quotients higher than one found for Pb in vegetable consumption, whereas in rice consumption Cd and Cu had the greatest contribution to the hazard index (Zhang et al., 2019) . Table 14: Non -carcinogenic hazard index (HI) for different age group population living in China nearby e -waste dismantling activities exposed to potentially toxic elements ( PTEs ) via dietary intake reported by Zhang et al. (2019) . Food type Receptor n Hazard index PTEs of concern (HQ>1) PTEs included Vegetable Children (2 to 7 years) 21 2.51† Pb Children / teenager s (8 to 19 years) 30 1.7† - Pb, Cd, Cu, Zn Adults ( ≥20) 88 1.47† - Rice Children (2 to 7 years) 21 7.74† Cd, Cu, Zn Children / teenagers (8 to 19 years) 30 5.94† Cd, Cu Adults ( ≥20) 88 4.76† Cd, Cu Egg Children (2 to 7 yr) 21 <1 - Children / teenagers (8 to 19 years ) 30 <1 - Adults ( ≥20) 88 <1 - Drinking water Children (2 to 7 years ) 21 <1 - Children / teenagers (8 to 19 years ) 30 <1 - Adults ( ≥20) 88 <1 - Fish Children (2 to 7 years ) 21 <1 - Children / teenagers (8 to 19 years ) 30 <1 - Adults ( ≥20) 88 <1 - Meat Children (2 to 7 years ) 21 <1 - Children / teenagers (8 to 19 years ) 30 <1 - Adults ( ≥20) 88 <1 - Abbreviations: Number of samples (n); hazard quotient (HQ); potentially toxic elements (PTEs). 34 The determination of hazard index reported by different authors for each exposure pathway was variable and appears inconsistent in places. It was beyond the resources available in the present study to compare the methods employed in detail, but there are clearly differences in the number and type of elements investigated a nd the absorption factors used, for example, Ackah (2019) and Ohajinwa et al. (2019a) . Most of the researchers determined the total concentrations of PTEs in different sources (for example, soil, dust, air and food) for the calculation of non -cancer risk indices (hazard index and hazard quotients) leading to a potential overestimation of human health risk effects. An alternative approach would have been to determine the concentration of bio -accessible PTEs which has been reported to provide a more accurate representation of risk (Cao et al., 2020; Oguri et al., 2018) . Cao et al. (2020) reported t hat the exposure risk can be assessed with higher accuracy by using bio - accessible concentrations of PTEs and bio -accessibility -corrected human health risk assessment was recommended. 3.1.6. Carcinogenic risk The carcinogenic risk is the incremental probability of a receptor, exposed to carcinogenic hazards, to develop cancer over a lifetime. The carcinogenic risk is determined by average daily intake and the cancer slope factor, for which indicative values are provided by US Environmental Protection Agency (United States Environmental Protectio n Agency, 2019a) . Tolerable cancer risk levels range between 10-6 and 10-4 depending on the situation and circumstances of exposure (NHMRC, 2010; United States Environmental Protection Agency, 1991) . Levels less than 10-6 indicate that cancer risk is negligible, while values greater than 10-4 are unacceptable (NHMRC, 2010) . As with the non -cancer risk assessments, cance r risks resulting from PTE exposure for specific pathway/receptor combinations in e -waste dismantling activities were also investigated ( Table 15). 35 Table 15: Cancer risk (hazard index) from potentially toxic element (PTE) exposure reported for several pathway -receptor scenarios in e -waste disman tling zones . Pathway Media Ref. Geog. Media description Activity context Receptor n Hazard index PTEs included Inhalation Air Huang et al. (2016) CHN From three heights above the ground E-recycling site Adults 165 1.3E -03† Children 165 3.9E -03† Cr, Co, Ni, As, Cd Adults 165 5.4E -05 Children 165 1.6E -04† Co, Ni, As, Cd Fang et al. (2013) CHN PM 2.5 air samples Mech. sorting area Workers 2 3.5E -04† Cr 2 9.4E -06 Ni 2 1.5E -06 Cd 2 5.6E -06 Pb PM 2.5 air samples Dis. area 2 1.4E -04† Cr 2 6.6E -06 Ni 2 2.0E -06 Cd 2 9.9E -06 Pb PM 2.5 air samples Southeast of workshops Residents 2 1.0E -04† Cr 2 3.7E -06 Ni 2 3.3E -07 Cd 2 3.3E -07 Pb Ingestion Soil Ackah (2019) GHA Topsoil (d=0 -30cm) Dis. area Children 41 2.6E -04† As Adults 41 3.6E -05 Subsoil (d=30 -100cm) Children 37 8.0E -06 Adults 37 1.1E -06 Ohajinwa et al. (2019a) NGA Topsoil (d=0 -10 cm) Dis. area* Workers n/a 2.3E -05 - 6.8E -05 Cr, Co, Ni, As, Cd, Hg, Pb Repair area n/a 1.2E -05 Ingest ion Dust Ohajinwa et al. (2019a) NGA Floor dust (control soil) Dis. area** Workers n/a 1.9E -05 - 8.9E -05 Cr, Co , Ni, As, Cd, Hg, Pb Repair area* n/a 1.1E -05 - 8.7E -05 Roadside dust Dis. area n/a 2.1E -05 Repair area n/a 6.0E -06 Direct dust from e -waste Dis. area n/a 3.4E -06 Repair area n/a 3.1E -05 Dermal exposure Soil Ohajinwa et al. (2019a) NGA Topsoil (d=0 -10 cm) Dis. area** Workers n/a 2.8E -03† - 21E-03† Cr, Co, Ni, As, Cd, Hg, Pb Repair area n/a 1.2E -03† Floor dust (control soil) Dis. area** Workers n/a 3.1E -03† - 35E-03† Cr, Co, Ni, As, Cd, Hg, Pb 36 Pathway Media Ref. Geog. Media description Activity context Receptor n Hazard index PTEs included Dermal exposure Dust Ohajinwa et al. (2019a) NGA Repair area* n/a 7E-04 - 52E-04† Roadside dust Dis. area n/a 5.2E -03† Repair area n/a 4.1E -04† Direct dust from e -waste Dis. area n/a 3.9E -04† Repair area n/a 4.2E -03† † Exceeds target value of 1.0E-04 indicating unacceptable ri sk. Abbreviations: Geographical context (Geog.); number of samples (n); particulate matter < 2.5 µm (PM2.5); potentially toxic el ements (PTEs) ; dismantling (dis.); mechanical (mech.) ; not available (n/a) . 37 The cancer risk induced by air inhalation for non -occup ationally exposed receptors (adults and children) was determined by Huang et al. (2016) and for occupationally exposed receptors by Fang et al. (2013) . Huang et al. (2016) found high cancer risk for both children and adults living nearby e -waste dismantling activities with Cd being the main contributor. The cancer risk calculated by Huang et al. (2016) approaches almost the threshold value if Cd is excluded, indicating that Cd is the main reason for high cancer risk via air inhalation for both receptors. In addition, Fang et al. (2013) ranked the PTE’s contribution to lifetime cancer ris k in the following order Cr > Ni > Pb > Cd, indicating that Cr poses the highest cancer risk on workers. For the ingestion of PTEs via soil pathway, the cancer risk for different receptors was investigated by Ackah (2019) and Ohajinwa et al. (2019a) . In Nigerian e -waste dismantling sites, Ohajinwa et al. reported no cancer risk through ingestion of soil. Ackah also reported no risk for adults; howeve r, the risk of cancer to children from topsoil was greater than the target value, indicating an unacceptable level of risk. Limited information can be found for the cancer risk as a result of dermal exposure to contamina ted soil or dust by PTEs. Only Ohajinwa et al. (2019a) reported considerably high values for both sources (dust and soil), indicating the need for further investigation on human health risk of dermal exposure to PTEs in e -recycling areas. 3.2. Flame retardants 3.2.1. Context Plastics used in electrical and electronic goods are often modified with additives to inhibit combustion in the event of malfunction. Thes e additives are known collectively as ‘flame retardants’ and comprise several groups of substances , including brominated flame retardants (BFRs) and chlorinated paraffins (CP). Both groups of substances include multiple formulations, some of which are pote ntially hazardous (National Research Council (US) Subcommittee on Flame -Retardant Chemicals, 2000) . The BFR group comprises 209 congeners divided into three formulas: Penta -BDE, Octa - brominated diphenyl ether (BDE ) and Deca -BDE (McGrath et al., 2017) . An increasing body of evidence exists which implicates multiple BFR congeners in serious health effects including cancer, diabetes, reproductive health, neurobehavioral and developmental 38 disorde rs, and alteration in thyroid function (Hong -Gang et al., 2016; Kim et al., 2014; McGrath et al., 2017) . Penta -BDE , Octa -BDE and Deca -BDE formulations are all listed as persistent organic pollutants (POPs) under the Stockholm Convention (Tang et al., 2014; UN Environment, 2017) and production of both Penta -BDE and Octa -BDEs is banned in the US (Venkatesan and Halden, 2014) and Europe, despite continued production in China (Tang et al., 2014) . More than 200 formulations of CPs exist which can be broadly classified by the length of the polymer as follows: short -chain CPs (SCCPs, C 10-13), medium chain CPs (MCCPs, C 14−17) and long -chain (LCCPs, C >17) (National Research Council (US) Subcommittee on Flame - Retardant Chemicals, 2000) . CPs have generally low acute toxicity and a recent health risk assessment carried out by the EFSA Panel on Con taminants in the Food Chain (CONTAM), was inconclusive (Efsa Panel on Contaminants in the Food Chain et al., 2020) . Nonetheless, CPs have high bioaccumulation potential, particularly for lower chlorinated species, and SCCPs are carcinogenic to rats and some toxicity has been observed in aquatic organisms (Iozza et al., 2008) . SCCPs have also been listed as POPs under the Stockholm Convention (Secretariat of the Stockholm Convention, 2017) , as possibly carcinogenic to humans by the International Agency for Research on Cancer (IARC) and as hazardous under the European Water framework Directive. 3.2.2. Air Environmental release rates of BFRs and CPs are only partly understood (Webster et al., 2009) and in this research, incomplete information was iden tified to evidence emissions from sorting, dismantling and physical processing of e -waste. Many authors also report results which are potentially confounded, especially in LIMICs where open burning may take place in proximity to non -thermal processing. The four studies of flame retardant concentrations in air presented in Table 16 all reported data from facilities where, according to the text, open burning did not appear to be taking place. One of these, Muenhor et al. (2010) reported concentrations of PBDEs in the air at five e -waste storage facilities in Thailand. Referring to Table S 14 in Section S.4.6 (supplementary information) many of the concentrations reported by Muenhor et al. (2010) were not particularly high in comparison to mean concentrations reported in outdoor urban air in the USA (0.052, 0.1 ng m-3), Italy (0.106 ng m-3) and China (0.045 ng m-3). However , several were higher than those reported in Spain (0.035, 0.018 ng m-3), the UK (0.018 ng m-3), Sweden (0.0063 ng m-3), Greece (0.026, 0.015 39 ng m-3), Australia (0.0093 ng m-3), and China (0.019, 0.025, 0.045 ng m-3). Notably, concentrations in the indoor air at Facility 5 were approximately an order of magnitude higher than many of three other observed concentrations, in dicating the possibility that migration and passive release is a potentially significant source of BFR emissions . Nonetheless , the concentrations were still an order of magnitude lower than computer laboratories in C alifornia investigated by Cahill et al. (2007) . Table 16: Flame retardant concentrations in air at e -waste storage and processing facilities (ng m-3). Author Geog. Activity context Sample n ΣPBDE a-DP s-DP ΣCP ΣotherBFR Muenhor et al. (2010) THA E-waste storage F1 Indoor 1 0.046 Outdoor upwind 1 0.024 Outdoor downwind 1 0.075 F2 Indoor 1 0.052 Outdoor upwind 1 0.012 Outdoor downwind 1 0.008 F3 Indoor 1 0.14 Outdoor upwind 1 0.061 Outdoor downwind 1 0.11 F4 Indoor 1 0.35 Outdoor upwind 1 0.15 Outdoor downwind 1 0.057 F5 Indoor 1 0.046 Outdoor upwind 1 0.033 Outdoor downwind 1 0.023 Mean 0.079 Median 0.052 Gravel et al. (2019b) CAN Recycling (control)* Facility 15 88 0.71 0.36 1.2 0.2 E-waste recycling Small facility 22 320 3.3 3.1 6.5 98 Medium facility 30 810 10 6.3 16 85 Large facility 36 6,600 27 14 41 150 Julander et al. (2005) SWE E-waste dismantling Respirable dust 3 6.17 Inhalable dust 4 214.27 Total dust 4 33.35 Cahill et al. (2007) USA Computer lab. (control) Computers switched off 2 1.4 Computers switched on 6 1.8 E-waste dismantling Shredding a 4 650 No shredding a 2 93 *Facility reported to recycle commercially sourced glass and a small amount of aluminium and cardboard; a samples were taken at the same plant on days when the sh redder was active and inactive. A bbreviations: facility (F); number of samples (n); Geographical context (Geog.); polybrominated diphenyl ethers (PBDE); chlorinated paraffins (CP); Dechlorane Plus (DP); brominated flame retardants (BFR). As with PTE emissions discussed in Section 3.1, concentrations of flame retardants measured in air at HIC e -waste physical processing plants provide a useful benchmark of the potential 40 emissions from non -thermal processing because open burning and de -soldering activities are highly unlikely to be taking place. The studies by Gravel et al. (2019b) , Julander et al. (2005) and Cahill et al. (2007) provide such a context as they were carried out in HICs, showing a range of values at each of the plants being studied. Air c oncentrations at the control site determined by Gravel et al. (2019b) were comparable with many of the concentrations measured by Muenhor et al. (2010) , however the concentrations in the medium and large plants were considerably higher, with the large facility measuring two orders of magnitude greater than the control. Of the 6,600 ng m-3 mean concentration detected at the large plant, 6,500 ng m-3 (range 5,300 –8,400) was the BDE 209 congener (specific congeners not shown). The study of a US e -waste dismantling plant by Cahill et al. (2007) showed similar concentrations of BFRs when the shredding equipment was switched on compared to the medium plant investigated by Gravel et al. (2019b) . When the shredder was not operational, the BFR concentrations decreased by approximately an order of magnitude. This study provides compelling evidence to support the theory that BFRs are emitted during shredding activities. Gravel et al. (2019b) also determined concentrations of two Dechlorane plusTM stereoisomers, anti-DP and syn -DP. The concentrations are hard to contextualise because although dechloranes have been in production in increasingly large quantities since the 1960s, their detection in the environment wasn’t reported until 2006 (Sverko et al., 2011) . Since then, dechloranes have been identified in multiple global locations, and their detection and potential hazardousness are an ongoing area of research. Due to their low water solubility and high octanol -water partition coefficient, in 2019 a proposal was submitted to the Stockholm Convention to have dechloranes added to Annexes A, B and/or C (Secretariat of the Stockholm Convention on Persistent Organic Pollutants, 2019) . While acknowledging the persistence in the environment of dechloranes, the intersessional working group of the Stockholm Convention indicates acute toxicity is unlikely, although longer term studies h ave not been carried out (Secretariat of the Stockholm Convention, 2020) . 3.2.3. Dust and soil Studies of flame -retardants in soils and dusts in and around e -waste physical dismantling and processing facilitie s are another indication of emissions, evidenced in four studies, Chen et al. 41 (2018) , Ohajinwa et al. (2019a) , Muenhor et al. (2010) and Ma et al. (2009) (Table 17) with three further studies, Wang et al. (2013) , Harrad et al. (2008) and Iwegb ue et al. (2019) shown for comparison. Chen et al. (2018) measured concentrations of CPs in e -recycling workshop dust in China, finding concentrations of SCCPs and MCCPs ten times higher than in residential areas 2 km away and approximately 76 (SCCPs) and 80 (MCCPs) times higher than the reference homes 30 to 40 km away. All the concentrations of SCCPs were considerably higher than those observed by Wang et al. (2013) whose highest measurement at the roadside in China was two orders of magnitude lower than the concentrations identified in reference homes by Chen et al. (2018) . 42 Table 17: Flame retardant concentrations observed in dust soil in and around e -waste dismantling facilities. Author Geog. Activity context Sample n Unit ΣPBDE ΣSCCP ΣMCCP Mean Median Chen et al. (2018) CHN E-recycling workshop Dust 41 µg g-1 3,760 13,000 Residents <2 km 30 370 1,130 Nearby streets 10 359 567 Ref. homes (30 -40 km) 15 49 162 Wang et al. (2013) CHN Farmland Soil 18 µg g-1 0.0012 -0.21 Roadside 6 0.031 -0.42 Woodland 3 0.00042 -0.014 Ohajinwa et al. (2019a) NGA, Lagos Dismantling site s Soil* Soil (n=16) Floor dust (n=29) Roadside dust (n=5) Direct (n=6) ng g-1 1.579 Repair sites 0.00319 Dismantling sites Floor dust (control soil) 3.134 Repair sites 0.452 Roadside at dismantling site s Dust 1.06 NGA, Ibadan Dismantling site s Soil * 0.018 Repair sites Floor dust (control soil) 8.22 From electronics Dust (direct) 3.594 From electronics 1.685 NGA, Aba Repair site s Floor dust 0.55315 Roadside Dust (control RS) 0.1114 Muenhor et al. (2010) THA E-waste storage facilities Dust 25 ng g-1 43,000₸♯ Ma et al. (2009) CHN E-waste dismantling shredding facility Dust 5 ng g-1 1,910 Soil 10 30,700₸♯ Leaf 6 30.6 Urban soil ( Ref.) 3 0.30 Harrad et al. (2008) GBR Homes Dust 30 ng g-1 260,000₸♯ 8,500♯ Offices 18 31,000₸♯ 7,400♯ Cars 20 340,000₸♯ 57,000₸♯ Iwegbue et al. (2019) NGA Computer/photocopier/printer workshops Dust 10 ng g-1 457 366 TV/radio/stereo/video player workshops 20 1,112 958 Telephone/tablet workshops 10 622 704 * d=0 -10 cm; dust (direct) means dust collected directly from the circuit -boards inside electronic equipment; soil and dust concentrations benchmarked against USEPA screening levels (detailed in Section S.4.7 ) as follows: † exceeds USEPA carcinogenic screening level (TR=1E -06) for BDE -209 in industrial soils; ‡ exceeds USEPA non -carcinogenic screening level (TR=1E -06) for BDE -209 in industrial soils; ₸ exceeds USEPA non -carcinogenic screening level (TR=1E -06) for BDE -153 in industrial soils; ♯ exceeds USEPA non -carcinogenic screening level (THI=0.1) for BDE -47 and BDE -99 in industrial soils. Abbreviations : Polybromi nated diphenyl ether (PBDE); SCCP (short chain paraffins); MCCP (medium chain paraffins); number of samples (n) ; Geographical context (Geog.) ; reference (Ref.) . 43 The PBDE concentrations reported by Ohajinwa et al. (2019a) were extremely low in comparison to all other authors other than the reference urban soil samples analysed by Ma et al. (2009) . Even the highest concentrations identified by Ma et al. (2009) and Muenhor et al. (2010) were lower than dusts analysed by Harrad et al. (2008) in UK offices, and lower by an order of magnitude compared to UK cars and homes. While initially surp rising, it is important to note that the likely source of the PBDEs in the study by Harrad et a l. (2008) is soft furnishings , in which plastic fibres have a much higher surface area compared to electrical items. While it is beyond the scope to consider these mechanisms in more detail, it is suggested that dust accumulating on the soft furnishing fibres act as a conduit for PBDE migration. Some evidence for this is also indicated by Ohajinwa et al. (2019a) who found some of the highest concentrations of PBDEs on dust sampled from the surface of electronic circuit boards, indicating that migration into dust is a significant pathway through which PBDEs migrate from plastic electrical components. Whereas many papers reviewed here mentioned non -thermal emissions of BFRs, discussion of the mechanism of release was brief or non -existent, focusing on open burning , which is thought to be one of the main mechanisms of BFR emissions (Hong -Gang et al., 2016) . Non- thermal emissions of BFRs or CPs require their migration from the spaces between the host polymer chains to the surface of the plastic (Stubbings and Harrad, 2014) , which may be hastened through the increased surface area created through comminution, the latter of which was suggested as a potential source of BFR emissions at rural e -waste processing sites in China by Tian et a l. (2011) . Other BFRs , such as TBBPAs, are less likely to be released through migration as they are principally used as a reactant in epoxies, phenolic resins and polycarbonate . 3.2.4. Non-carcinogenic risk Only one study in Nigeria by Ohajinwa et al. (2019a) calculated hazard quotients for non - carcinogenic risk from exposure to BDE congeners fro m e-waste dismantling and repair activities ( Table 18). Direct dust from electronics handled in the repair are resulted in the greatest risk to workers, inside e -waste dismantling and repair areas through dermal contact. One site, Ibadan, returned the highest hazard index , as well as Aba, whereas the Lagos site showed an index of below two in all exposure pathway categories. The ingestion pathway was only calculated to pose a risk from exposure to dust that had been in direct contact with circuitry at the Ibadan dismantling workshop, whereas the pathway did not result in hazard index of greater than one for any other context. 44 Table 18: Non-carcinogenic hazard indices reported by Ohajinwa et al. (2019a) for polybrominated diphenyl ether ( PBDE ) exposure to workers in Nigerian e - waste dismantling and repair workshops. Pathway Sample Geog. Sample source location Hazard quotients Hazard index PBDE -47 PBDE -99 PBDE -153 PBDE -209 Dermal exposure Dust NGA, Ibadan Direct dust from e -waste Dismantling area 340†† 810†† 88† 2,300†† 3,700† Repair area 78†† 140†† 33† 110† 360† NGA, Lagos Floor dust Dismantling area <1 <1 <1 <1 <1 Repair area <1 <1 <1 <1 <1 NGA, Ibadan Floor dust (control soil) Repair area (control soil) 22†† 57†† 20†† 610†† 710† NGA, Aba Repair area (control soil) 7.6† 14†† 4.3† 41† 66† NGA, Lagos Roadside dust Dismantling area <1 <1 <1 <1 <1 NGA, Aba Repair area 21† 31† 21† 5.2† 78† Soil NGA, Lagos Topsoil (d=0 -10 cm) Dismantling area <1 <1 <1 <1 <1 Repair area <1 <1 <1 <1 <1 NGA, Ibadan Dismantling area 1.1† 2.2† 1.3† <1 5.6† Ingestion Dust NGA, Ibadan Direct dust from e -waste Dismantling area <1 <1 <1 1.74† 2.64† Repair area <1 <1 <1 <1 <1 NGA, Lagos Floor dust (control soil) Dismantling area <1 <1 <1 <1 <1 Repair area <1 <1 <1 <1 <1 NGA, Ibadan <1 <1 <1 <1 <1 NGA, Aba <1 <1 <1 <1 <1 NGA, Lagos Roadside dust Dismantling area <1 <1 <1 <1 <1 NGA, Aba Repair area <1 <1 <1 <1 <1 Soil NGA, Lagos Topsoil (d=0 -10 cm) Dismantling area <1 <1 <1 <1 <1 Repair area <1 <1 <1 <1 <1 NGA, Ibadan Dismantling area <1 <1 <1 <1 <1 † Exceeds target value of one, indicating unacceptable risk; †† exceeds target value of 10 , indicating very high risk. Abbreviations: polybrominated diphenyl ether ( PBDE ); number of samples (n) ; Geographical context (Geog.). 45 3.2.5. Carcinogenic risk As with the non -carcinogenic risks, the carcinogenic risks from e -waste dismantling activities were highest from exposure to dust on circuit boards at the e -waste dismantling workshop in Nigeria ( Table 18). Table 19: Cancer risk for several pathway/receptor combinations induced by e-waste dismantling activities (specific congener BDE -209); after Ohajinwa et al. (2019a) . Pathway Sample Geog . Media Activity context Receptor Cancer risk (CR) Dermal exposure Soil NGA, Lagos Topsoil (d=0 -10 cm) Dismantling area Workers 5.6E -07 Repair area 3.2E -10 NGA, Ibadan Dismantling area 4.8E -06 Dust NGA, Lagos Floor dust (control soil) Dismantling area Workers 1.1E -06 Repair area 1.6E -07 NGA, Lagos Roadside dust Dismantling area 3.6E -07 NGA, Ibadan Floor dust (control soil) Repair area 3.0E -03† NGA, Ibadan Direct dust from e -waste Dismantling area 1.2E -02† Repair area 5.5E -04† Aba, Nigeria Floor dust (control soil) Repair area 2.1E -04† Roadside dust 2.6E -05 Ingestion Soil NGA, Ibadan Topsoil (d=0 -10 cm) Dismantling area Workers 3.5E -09 Dust NGA, Lagos Floor dust (control soil) Dismantling area Workers 8.1E -10 Repair area 1.2E -10 NGA, Lagos Roadside dust Dismantling area 2.6E -10 NGA, Ibadan Floor dust (control soil) Repair area 2.2E -06 NGA, Ibadan Direct dust from e -waste Dismantling area 8.5E -06 Repair area 3.9E -07 Aba, Nigeria Floor dust (control soil) Repair area 1.4E -07 Roadside dust 1.9E -08 † Exceeds target value of 1.0E -04 for BDE -209 or 1.0E -06 for polycyclic aromatic hydrocarbons ( PAH ) indicating unacceptable risk . Abbreviations : Geographical context (Geog.) . 3.3. Risk characterisation for sorting, dismantling and physical processing of e - waste The qualitative risk assessment score s for sorting, dismantling and physical processing of e - waste ar e shown in Table 20. Informal waste workers were scored with very high risk for dermal con tact with PTEs and medium to high risk for dermal contact with BFRs and inhalation of PTEs during physical processing of e -waste. Formal workers also scored medium to high risk for inhalation of PTEs while working in proximity to mechanical processing as d espite better standards of protective equipment and safety procedures in general, it is suggested that the more highly mechanised systems result in considerably greater emissions compared to the less mechanised informal sector operations. However, the carcinogenic risk of exposure to PTEs has been reported to exceed the acceptable level (10-6- 46 10-4) for informal workers, as a result of limited accessibility to protective equipment risk of ingestion, high level of dermal contact and inhal ation. Children living on e -waste sites are also at high risk from ingestion of PTEs because they eat more soil than adults and at medium to high risk from PTE s uptake in food. Specifically, the levels of Pb and Cd in blood of children living nearby e -waste disma ntling activities were found to be concededly high er compared to non -exposed children . The risk for population nearby e -waste recycling activities is medium to high due to e levated levels of PTEs identified in dust and soil of surrounding areas, but still the risk is higher for children compared to adults. Apart from the very high risk to informal workers identified, the risk from BFRs to the general population was low for children and adults as the detection of these substances in environmental media was not sufficient t o pose a significant threat. 47 Table 20: Risk characterisation summary for sorting, dismantling and physical processing of e -waste (non -thermal). Haz. Pathway Receptor Geog. Evidence and justif ication for risk assessment Uncertainty (aleatoric and epistemic) Receptor vulnerability L S R Global receptor context PTE Atmosphere/ Inhalation E-waste workers (formal) CHN, SWE, USA, GHA  Clear evidence (Ceballos et al., 2017; Julander et al., 2014) of emissions to air of several PTEs through physical processing in HICs, potentially greater than in LIMICs due to more intensive mechanised processing.  Dismantling workers significantly more exposed compared to office workers for most elements (Ceballos et al., 2017) .  Emissions of Pb and Cd exceeded gu idelines ( Section S.4.1) in two examples relating to shredding and CRT processing (Ceballos et al., 2017) .  Significantly elevated blood concentrations of Cr and Pb for those engaged in HIC physical processing activities (Julander et al., 20 14) and exceeding reference value for non -exposed Canadian citizens (Ceballos et al., 2017) .  Elevated Pb in dismantlers urine significantly higher than office workers (Julander et al., 2014) .  Limited information on specific processing activities creates challenge to establish causality from particular processes.  Use of PPE sometimes optional and not always used (Ceballos et al., 2017; Julander et al., 2014) . 3 4 12 HIC E-waste workers (informal)  Evidence for informal sector indicates concentrations low in areas where physical processing takes place (Huang et al., 2016; Tang et al., 2015; Zeng et al., 2016) , however evidence of elevated blood Pb in one study (Amankwaa et al., 2017) .  Non-carcinogenic hazard index indicates low risk to informal waste workers in two scenarios studied (Fang et al., 2013) .  Carcinogenic risk indicates slightly above 1.3E -03 in one case (Fang et al., 2013) .  Results may be confounded with open burning and smelting that may take place in the sa me location.  Informal workers operate without respiratory protective equipment. 3 4 12 LIMIC Population  Non-carcinogenic risk to adults and children high in two examples (Huang et al., 2016) .  Carcinogenic risk to population (Huang et al., 2016) indicated in four examples.  Significantly higher Cd and Pb for residents living near to e -waste dismantl ing (Zhang et al., 2019) .  Children living in e -waste areas have significantly higher Pb and Cd in blood compared to non -exposed in multiple studies (Cai et al., 2019; Liu et al., 2018; Zeng et al., 2017; Zeng et al., 2016; Zhang et al., 2017; Zheng et al., 2019) , however inhalation is not considered a very prevalent source of exposure to children.  Blood levels confounded with other exposure pathways, particularly ingestion.  Adults and children have no choice to avoid exposure if they live around e -waste dismantling activities. 2 4 8 LIMIC Children  Children have no choice to avoid exposure if they live around e -waste dismantling activities. 3 4 12 LIMIC PTE Ingestion of dust, soil and water E-waste workers (informal) PHL, GHA, NGA, THA, IND, GRC, CHN  Significant non carcinogenic risk identified in one case, specifically dust but low in all others (Ohajinwa et al., 2019a) .  Results may be confounded with open burning and smelting that may take place in the same location.  Low hygiene means informal workers at risk of ingesting, particularly dust that they come into contact with. 1 4 4 LIMIC Population  Elevated levels of PTEs identified in multiple dust and soil samples close to e -waste dismantling activities (Ackah, 2019; Chakraborty et al., 2019; Damrongsiri et al.,  Adults and children have no choice to avoid exposure if they live 3 4 12 LIMIC 48 Haz. Pathway Receptor Geog. Evidence and justif ication for risk assessment Uncertainty (aleatoric and epistemic) Receptor vulnerability L S R Global receptor context 2016; Fang et al., 2013; Fujimori et al., 2012; Isimekhai et al., 2017; Ohajinwa et al., 2018; Tang et al., 2015; Tzoraki et al., 2019) in LIMICs and one in an HIC.  Cd, Pb, Cr and Mn with multiple example s of exceeding USEPA guidelines.  Levels of PTEs low in few samples of fresh water and very low in tap water analysed (Tang et al., 2015; Zhang et al., 2019) in LIMICs but elevated level s of As, Cd, Cr, Pb in one HIC context (Tzoraki et al., 2019) .  Significantly higher Cd and Pb for residents living near to e -waste dismantling (Zhang et al., 2019) .  Children living in e -waste areas have significantly higher Pb and Cd in blood compared to non -exposed in multiple studies (Cai et al., 2019; Liu et al., 2018; Zeng et al., 2017; Zeng et al., 2016; Zhang et al., 2017; Zheng et al., 2019) .  Significant non -carcinogenic risk through ingestion of dust and soil to adults and children in several cases (Ackah, 2019; Fujimori et al., 2012; Zhang et al., 2019) .  Carcinogenic risk only identified in one case for children (Ackah, 2019) . around e -waste dismantling activities. Children  Children living in e-waste areas have significantly higher Pb and Cd in blood compared to non -exposed in multiple studies (Cai et al., 2019; Liu et al., 2018; Zeng et al., 2017; Zeng et al., 2016; Zhang et al., 2017; Zheng et al., 2019) .  High risk to children living in and near e -waste dismantling activities (Ackah, 2019; Fujimori et al., 2012) .  Carcinogenic risk to children identified in one case (Ackah, 2019) .  Children ingest soil and have no choice to avoid exposure if they live around e -waste dismantling activities. 4 4 16 LIMIC Uptake in food Population CHN  Some evidence (Tang et al., 2015; Zhang et al., 2019) provided for concentrations in food near e -waste physical processing, with three examples showing slightly higher Pb concentration than EU regulatory limits (European Union, 2006; European Union, 2015) , however evidence is not strong enough to correlate with the e-waste processing activities.  High non -carcinogenic risk through dietary intake highlighted in one study particularly through dust settling on vegetables and rice (Zhang et al., 2019) .  Limited information indicates that many PTEs are not easily up -taken by rice plants (Wu et al., 2019) , however further sources were not reviewed.  Assessment is based on a single estimate of non - carcinogenic risk in China (Zhang et al., 20 19).  Local population may have limited ability to choose whether to eat food from local area or not.  3 4 12 LIMIC Children  Non-carcinogenic risk through dietary intake highlighted in one study, particularly through dust settling on vegetables and rice (Zhang et al., 2019) , indicated slightly higher for children.  3 4 12 PTE Dermal contact with soil and dust Population PHL, GHA, NGA, THA, IND, GRC, CHN  Multiple examples of PTE concentrations in soils and dusts (Ackah, 2019; Chakraborty et al., 2019; Damrongsiri et al., 2016; Fang et al., 2013; Fujimori et al., 2012; Isimekhai et al., 2017; Ohajinwa et al., 2018; Tang et al., 2015; Tzoraki et al., 2019) in local environment.  Non-carcinogenic risk low according to one study (Ackah, 2019) because contact with dusts limited in comparison to occupational exposure.  Results may be confounded with open burning and smelting that may take place in the same location.  Adults and children have no choice to avoid exposure if they live around e -waste dismantling activities. 1 4 4 LIMIC 49 Haz. Pathway Receptor Geog. Evidence and justif ication for risk assessment Uncertainty (aleatoric and epistemic) Receptor vulnerability L S R Global receptor context E-waste workers (informal)  Multiple examples of PTE concentrations in soils and dusts (Ackah, 2019; Chakraborty et al., 2019; Damrongsiri et al., 2016; Fang et al., 2013; Fujimori et al., 2012; Isimekhai et al., 2017; Ohajinwa et al., 2018; Tang et al., 2015; Tzoraki et al., 2019) , particularly dust on electronic components and circuitry that are frequently handled by informal waste wo rkers (Ohajinwa et al., 2018) .  Non-carcinogenic and carcinogenic risk considered extremely high for both soils and dusts (Ohajinwa et al., 2019a) .  One study (Fang et al., 2013) reports low risk through derm al contact with PM2.5 deposited from atmosphere, though this may be a minor source in comparison to direct electronic dusts.  Informal workers operate without dermal protective equipment . 4 4 16 LIMIC BFR Atmosphere/ Inhalation E-waste workers (formal) CAN, THA, SWE, USA  Evidence for release into atmosphere as a consequence of dismantling and physical processing in HICs is robust, including an inferred link between increased physical processing, such as shredding and great emissions (Cahill et al., 2007; Gravel et al., 2019b; Julander et al., 2005) .  Carcinogenic and non -carcinogenic risk not i nvestigated through this pathway.  Results may be confounded with open burning and smelting that may take place in the same location.  Limited occupational exposure data in LIMICs  Use of PPE sometimes optional and not always used (Ceballos et al., 2017; Julander et al., 2014) . 1 4 4 HIC E-waste workers (informal)  Informal workers operate without respiratory protective equipment. 1 4 4 LIMIC BFR Ingestion of dust, soil and water Population NGA, CHN, THA  Though concentrations have been determined in soils and dusts in surrounding e - waste dismantling areas (Chen et al., 2018; Ohajinwa et al., 2019b; Wang et al., 2013) , that are very high in some examples (Ma et al., 2009; Muenhor et al., 2010) , the levels are low in compariso n to indoor dusts in the UK, homes and offices and cars which residents are exposed to every day (Harrad et a l., 2008) .  Carcinogenic and non -carcinogenic risk not investigated through this pathway.  Results may be confounded with open burning and smelting that may take place in the same location.  Limited occupational exposure data in LIMICs  Adults and children have no choice to avoid exposure if they live around e -waste dismantling activities. 1 4 4 LIMIC E-waste workers (informal)  Non-carcinogenic risk determined by one study (Ohajinwa et al., 2019b) and considered low in all but one case involvin g dust.  Carcinogenic risk low in all cases reported in one study (Ohajinwa et al., 2019b) .  Low hygiene means informal workers at risk of ingesting, particularly dust that they come into contact with. 1 4 4 LIMIC BFR Dermal contact with soil and dust Population NGA, CHN, THA  Though concentrations have been determined in soils and dusts in surrounding e - waste dismantling areas (Chen et al., 2018; Ohajinwa et al., 2019b; Wang et al., 2013) , that are very high in some examples (Ma et al., 2009; Muenhor et al., 2010) , the levels are low in comparison to indoor dusts in the UK, homes and offices and cars which residents are exposed to every day (Harrad et al., 2008) .  Carcinogenic and non -carcinogenic r isk not investigated through this pathway.  Results may be confounded with open burning and smelting that may take place in the same location.  Limited occupational exposure data in LIMICs  Adults and children have no choice to avoid exposure if they live around e -waste dismantling activities . 1 4 4 LIMIC E-waste workers (informal)  Though concentrations have been determined in soils and dusts in surrounding e - waste dismantling areas (Chen et al., 2018; Ohajinwa et al., 2019b; Wang et al., 2013) , that are very high in some examples (Ma et al., 2009; Muenhor et al., 2010) ,  Informal workers operate without dermal protective equipment. 3 4 12 LIMIC 50 Haz. Pathway Receptor Geog. Evidence and justif ication for risk assessment Uncertainty (aleatoric and epistemic) Receptor vulnerability L S R Global receptor context the levels are low in comparison to indoor dusts in the UK, homes and offices and cars which residents are exposed to every day (Harrad et al., 2008) .  Non-carcinogenic risk calculated to be high in most cases (Ohajinwa et al., 2019b) .  Carcinogenic risk considered high in some cases of exposure to dust (Ohajinwa et al., 2019b) . Abbreviations: Abbreviations: likelihood (L); severity (S); risk (R); hazard being assessed (Haz.); low income and middle inc ome countries (LIMIC); high income countries (HICs); geographical context (Geog.); potentially toxic elements (PTE); Brominated flame retardants (BFR); personal protective equipment (PPE) ; low income and middle income countries (LIMICs); cathode ray tube (CRT). 51 4. Challenge 2: Hydrometallurgical treatment to recover metals 4.1. Conte xt As well as the major constituent metals , such as steel and copper, e -waste contains several precious metals , including : platinum (Pt), selenium (Se), gallium (Ga), gold (Au), palladium (Pd), tellurium (Te), silver (Ag), germanium (Ge), and tantalum (Ta ) (Abdelbasir et al., 2018) . These metals exist in electronic products and components as substances that are coated onto, or bonded to other substances and materials whose combined purpose is to conduct, resist or store electricity. The bonded state of these metals within e -waste presents a challenge for those who wish to recover them. They cannot be simply prised out or physically abraded , meaning that thermal (pyrometallurgical) or liquid (hydrometallurgical) techniques must be used for separation. Pyrometallurgical treatment (heating) has been discussed by Cook et al. (2020b) and therefore , this section will focus on hydrometallurgical treatment which involves the treatment of e -waste with acids and alkalis to leach and recover metals. A furthe r process, biometallurgical treatment , is also emerging as a potential technique, however at the time of writing this is thought to be at laboratory level of technological readiness (Abdelbasir et al., 2018) and is therefore not discussed here. 4.2. Types of hydrometallurgical process Compared to pyrometallurgical treatment, hydrometallurgy is reported to be a more ‘exact’ method of e -waste metal recovery, as it is more readily controllable and doesn’t require emission control (Hsu et al., 2019) . However, this assertion implies that the chemicals used in hydrometallurgical treatment are disposed of in a way which does not involve their uncontrolled release into soil or water -bodies . There are many different types of hydrometallurgical treatment, many of which have been reviewed comprehensively for formal sector operations by Khaliq et al. (2014) , Cui and Zhang (2008) and Abdelbasir et al. (2018) . Simply , hydrometa llurgy takes place in two steps: (1) Comminuted circuitry is immersed in a leaching solution (lixiviant) that may include: nitric acid, hydrochloric acid, aqua regia (nitric and hydrochloric acids mixed at a molar ratio of about 1:3), sulphuric acid, cyanide, or alkalis such as ammonia. Metal are dissolved into the solution which can be strained from the solids where the non - metals and non -target metals remain (Abdelbasir et al., 2018) . 52 (2) Metals are recovered from the dissolved solution by precipitation, electrorefining, electrowinning , ion exchange, solvent extraction, adsorption and cementation. The newly recovered solids then need to be dried out and are often melted into ingots for onward sale (Abdelbasir et al., 2018) . Looking at the list of lixiviants described in 1, it should be obvious that extreme ca re is required to prevent these liquids from coming into contact with humans and also to control the risk of discharge into the environment; each has the potential for toxicity and corrosion. In HICs, the use of these substances is generally tightly contro lled by robust and strictly enforced legislation, however the following sections reveal evidence of activity in LIMICs where the health and wellbeing of people and protection of the environment are lacking in some hydrometallurgical treatment processes. If managed poorly or not managed at all, there a several pathways through which receptors may be exposed to three substances and residues from hydrometallurgical treatment which are summarise in the conceptual model in Figure 4. Figure 4: Hazard exposure conceptual model (source –pathway –receptor) associated with the hydrometallurgical treatment of e-waste. There are tens, possibly hundreds of potential hydrometallurgical processes available to remove metals from e -waste (Abdelbasir et al., 2018; Cui and Zhang, 2008) , though this Sources Dust Inhalation Adult citizens Occupational (formal) Pathways Receptors Soil Groundwater Rivers streams & lakes Surface water Sediments Receptor Route of human exposure Environmental compartment Substance or material Transfer process Legend Deposition Mouthing Child citizens Ingestion Uptake in crops Food Bonding agents Substances in components Lixiviants (acids, alkalis) Atmosphere Dermal contact Mucous membranes Occupational (informal) Discarded Leaching activities 53 review finds little information on the specific types of proces ses used by the informal waste reclamation sector. In many publications reviewed here,