Environmental Implications, conclusions and perspectives

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Contaminations by metals

Contaminations of metals in e-waste sites are summarized as table 1.3. As presented, high contents of heavy metals, mainly Cu, Zn, Pb, Ni, Cr and Cd, were detected in soils, sediments and also dust. Elevated levels of contamination were identified in comparison to the Chinese national maximum allowable concentrations. Based on available data (table 1.3), Cu is usually of highest concentration among all heavy metals, and Zn is of the second highest [4, 5]. Open burning or combustion process is the most polluted approach among all, while acid washing and dismantling could lead to the release of metal-containing leachate and dust, respectively.
Despite the abovementioned commons, comparison among different regions and various processing approaches suggested diverse contamination patterns. The regions of e-waste sites seemed to be a less significant influential factor on the contaminating patterns compared to varied processing approaches. The levels of contamination are more related to the processes. For instance, the contents of Cu and Zn were comparable in both combustion sites of Qingyuan and Guiyu [26, 57], while survey conducted in Taizhou indicated that soil contamination pattern is related to the scale of workshops and that small-scale simple workshops caused more severe pollution with higher percentage of Pb and Hg compared to large-scale factories [62].
Soils and sediments in the vicinity of e-waste recycling sites were inevitably contaminated due to these activities. Paddy soils receiving the dust and leachate from combustion sites were recognized to be contaminated mainly by Cd [26, 60]. And 24 of the 29 soil samples from the vicinity of dismantling sites were found to contain Cd at a very high content, which even surpasses the Cd concentration (4.01 mg/kg) detected in Japan Itai-Itai disease area. Contaminations by Zn and Cd around workshops were demonstrated more severe than that in open burning soils [66].
Besides soils, sediment is another important sink of pollutants, and may act as a carrier and source of heavy metals in estuary systems [54, 67], as waters across the e-waste sites were often the acceptors for industrial waste, which is thus severely contaminated by heavy metals. A study showed that two mainstreams in Guiyu catchment are contaminated by Cd (n.d. – 10.3 mg/kg), Cu (17.0 – 4540 mg/kg), Ni (12.4 – 543 mg/kg), Pb (28.6 – 590 mg/kg) and Zn (51.3 – 324 mg/kg). Non-residual metals are of higher proportion than that in control area, which imposes a risk on local environment and downstream ecology [53].
To summarize, different processing approaches determine the components and the quantity of the released metals from e-wastes, and as a result processing approaches determine the contamination patterns. Owing to the characteristics of geochemistry, soils and e-waste components different from region to region, contamination patterns of heavy metals vary across different sites. Thus it is important to understand the contamination pattern in a specific area and the goal of treatment before taking any remediation action. Decontamination strategy should be designed accordingly and one approach of an area should not be imposed directly to another one.

Contaminations by POPs

Similar to the situations of metal contaminations, the contaminations by organic pollutants (most of which are POPs as presented in table 1.2) varied diversely among regions and processing approaches. Surveys and reviews on the issue have been extensively and intensively conducted for the last decades [4].
One family of the typical contaminants is PAHs, which is primarily originated from the combustion process. The contaminations varied largely among surveys and are highly dependent on the sampling sites and methods. Commonly, elevated levels of PAHs (usually in the sum of the 16 PAHs in priority suggested by USEPA) in soils were detected at and near combustion sites, with the contents from 899.9 to 40000 µg/kg (dw) [68, 69]. In addition, soils near dismantling workshops and recycling facilities [6, 7, 61] were also contaminated by high contents of PAHs from 190.8 to 9156 µg/kg (dw). Based on the Dutch list: the optimal cumulative concentration of 10 PAHs, most of the soil samples near e-waste sites did not exceed the optimal values except for several sites near combustion or recycling facilities [4].
Besides PAHs, brominated flame retardants (BFRs) are also typical contaminants in these sites as many materials of electric and electronic products are flame retarded and most of the BFRs are not chemically bound to the materials [70]. Dismantling, shredding, combustion and other disposal treatments are able to accelerate the release of BFRs to the environment. The concerned species of BFRs are listed in table 1.2.
Among all the BFRs, PBDEs are one of the most concerned species as penta-BDEs and octa-BDEs have been officially listed as persistent organic pollutants (POPs) [28]. The total concentrations of PBDEs from e-waste sites were high in soils and dusts. The two soil samples most burdened by PBDEs (sum of 37 congeners) were reported in Guiyu that 25478.84 µg/kg (dw) in the surface soil near an abandoned dump/burying site for unsalvageable materials [71] and 36215 µg/kg (dw) near an opening burning site [72]. Contamination by PBDEs in Qingyuan was similar to that in Guiyu but less severe with 1140 – 1169 µg/kg d of ∑21PBDEs, which is 10 – 30 times of the amount in other regions [34]. The indoor and outdoor dust samples from Qingyuan e-waste sites were all found to be PBDEs contaminated. The total PBDE (17 congeners) concentrations for indoor dusts ranged from 230 – 157500 µg/kg (mean = 9400 µg/kg) , and 212 – 25880 µg/kg (mean value = 3311 µg/kg) for outdoor samples, respectively [73]. The high contents of PBDEs in both soils and dusts cast a threat to the health of local people and ecological system [8].

Chemical and physical properties of PCBs

The toxicity and behavior of PCBs in the environment are highly related to their properties. From 1920s to 1970s, PCB congeners (structure presented as figure 1.4) were often synthesized at industrial scale directly through chlorinated reaction on the molecule of biphenyl under high temperature and with metal catalysts. Theoretically, there are 209 congeners with various numbers of chlorine atoms on different substitution positions. Pure PCB compounds are in crystalline state at normal temperature and pressure while a mixture of PCBs is an oily liquid at the same condition. The mobility of a PCB compound decreases with higher number of chlorine atoms in the molecule, with its physical state shifting from mobile viscous liquid to solid resin. PCBs are stable with low aqueous solubility (-logSw=5.26-10.49) and relatively high octanol-water partition coefficient (logKow=4.66-8.02) [83]. The persistence of PCBs increases along with higher number of substituted chlorine atoms. PCBs are recognized as POPs, considering the durability, lipotropy, toxicity, the ability of long-distance migration and low bio-degradation rate [84].

Environmental behavior of PCBs in soils and sediments

The retention of PCBs on soils and sediments by sorption and/or partitioning on organic matters of soils and sediments is a result of their hydrophobicity and lipophilicity. Thus PCBs are more persistent in soils with higher content of organic carbons [92, 93]. It is estimated that in the British environment, 93.1% of PCBs burden is associated with soils and 2.1% with marine sediments [94]. Thus it is essential to study the environmental behavior of PCBs for further remediation strategies. PCBs in soils are mainly from atmospheric particle deposition and landfill leachate, with a low decontamination rate and a half-life of approximately 10-20 yr [95]. In natural soils, the removal of PCBs is limited via microbiological degradation. Other processes as volatilization, leaching, plant uptake and soil erosion, have negligible influences on the content of PCBs in soils [96].
PCBs enter into water bodies through atmospheric deposition or accepting industrial and municipal wastewater. Although a small amount of PCBs is soluble in water, the majority attaches on suspended particles and deposits to sediments at different speeds depending on particle features [97, 98]. PCBs accumulated in sediments can be degraded by anaerobic microorganisms yet at a low degradation rate. The half-life of PCBs varies greatly depending on the properties of different sediments and PCB congeners. Half-life of tetra-CBs and penta-CBs in Hudson River sediments in USA was estimated as approximately 10 yr [99]; half-life of PCB-105, PCB-126, PCB-156 and PCB-169 in the Rhine River sediments was 9 yr; half-life of PCB-31 in New Bedford Harbor sediments was as long as 465 yr while half-life of PCB-105 was only 4.4 yr, that of other congeners ranges from several years to several decades [100].

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Contamination by PCBs in soils and sediments

The large varieties of PCB congeners cause difficulties when making comparisons among different studies. To solve the problem, seven commonly used PCB congeners (PCB-28, -52, – 101, -118, -138, -153 and -180) with different extent of chlorinated substitution are used as representatives to assess the contamination by PCBs [101, 102]. The contaminations by PCBs at global scale are illustrated as table 1.4 with the exemption of the contamination in e-waste sites of mainland China. According to a report on National Priorities List (NPL) sites conducted by USEPA, approximately one third of the polluted areas of superfund were contaminated by PCBs to various extents [103], the most well-known example among which is Hudson River in New York State. The river had been contaminated by industrial activities in vicinity from 1950s to 1970s with up to 1000 mg/kg of PCBs at some sampling sites in 1980s [104]. Many industrial countries in Europe, from as south as the Mediterranean Sea to as north as the Baltic Sea [102] and across a large variety of latitude, from Slovenia to Scotland, have been contaminated by PCBs to various degrees [105]. In China, it is known that e-waste sites are the main PCB contaminated sites, of which contamination levels have been well-studied and discussed in section 1.1.2.2.
Based on the introduction and discussion above, it is of necessity and importance to gain more understandings on the contamination and adverse effects of PCBs. Remediation technologies to remove the contaminations also need to be developed.

Zerovalent iron (ZVI) mediated reduction

In the last two decades, ZVI has been used for the decontamination of ground waters and soils. The process of reactions together with influencing factors, mechanisms and applications have been intensively studied. As a strong reductant, ZVI is ready to be oxidized. The oxidation of ZVI in anaerobic condition could be described as below:
Fe0 2H Fe2H 2 Equation 1.1.
Fe 0 2H 2 OFe 2H 2 2OH Equation 1.2.
The reactivity of ZVI allows its application to the reduction of a variety species of contaminants, including metals, POPs and chlorohydrocarbons [145]. The reduction rate varies according to the surface properties of iron (such as the specific surface area, the number of catalytic sites and oxidative status) and the transfer efficiency of target compounds onto the iron surface as ZVI-mediated reduction is surface reaction.

Mechanisms of halogenated hydrocarbon degradation by zerovalent iron

Zerovalent iron (ZVI) is a frequently used material in permeable reactive barriers (PRBs) during the in situ remediation of contaminated groundwater. ZVI reacts with chlorinated hydrocarbon through beta-elimination reduction via stepwise hydrogenolysis (replacement of halogen by hydrogen) [180-182]. In the hydrogenation reduction, iron and hydrogen are regarded as electron donors while chlorinated pollutant serves as an electron receiver. This reduction is heterogeneous, reacting between between ZVI (solid) and liquid. The final products are mainly hydrocarbons with low-halogenated intermediates.
Research on ZVI degradation has been arisen to seek solutions to contaminations by PBDEs and PCBs. A study indicates that sequential debromination goes through different pathways in ZVI and bimetallic system. The reaction pathways with bimetallic system favor removal of para-halogens first then ortho-halogens on PBDEs and PCBs. While in ZVI system, ortho-halogens are preferentially removed followed by para-halogen removal [152]. Zhuang et al. (2011) inferred that the steric hindrance of ether bond and benzene ring hinders the catalyst (Pd) to insert in C-Br bond at ortho-position, slowing down the catalyzing reaction. While in bimetallic system, removal of para-halogens is favored [183]. Nano-scale ZVI also features in sequential dechlorination of PCBs. Chlorines in the para and meta position are predominantly removed over chlorines in the ortho position [18, 154]. Normally ring-opening reactions are not involved for the dehalogenation of halogenated biphenyls or diphenyl ethers, and the final products remain biphenyls or diphenyl ethers, respectively. Even though halogen removal reactions partially decrease the toxicity and hydrophobicity of pollutants, it is impossible to decompose and mineralize dehalogenated hydrocarbons by microbes in natural environment in a short term. Therefore, the application of ZVI degradation techniques is still restricted.
Along with the chemical reduction mechanisms, ZVI could also promote the degradation of target compounds by microorganisms under anaerobic conditions. The combined use of ZVI and anaerobic microbes for the treatment of various pollutants, including perchlorate, azos, PCBs and nitrates, has been demonstrated effective [184-186]. Two possible mechanisms are suggested to explain the role of ZVI herein: (1) ZVI has a significant selective effect on anaerobic microorganisms and influences their succession [187]; (2) H2 generated during the aqueous ZVI reaction may be a preferred electron donor for many bacteria, thus promoting the growth and activity of microorganisms [185].
The chemical and biological processes of reduction mediated by ZVI are often simultaneous and difficult to distinguish, especially during field applications. Insights into the mechanisms are still needed for comprehensive understandings.

Application of ZVI degradation in remediation

The application of ZVI is flexible since it could be used in situ or ex situ, on waters or soils (as illustrated by figure 1.9) [188]. ZVI was initially used as an active material in permeable reactive barriers (PRBs) for underground water treatment. According to a report from USEPA, ZVI has been applied to in situ remediation of contaminated ground water for more than two decades [189]. A survey in 2002 conducted in US revealed that the effective remediation material in 55% of PRB is ZVI [190].
Ground water treatment systems use ZVI fillings in PRB to degrade the chlorinated hydrocarbons in Denver area, Colorado, as well as to intercept pollutions of chromium VI and chlorinated hydrocarbons in Elizabeth City [191]. In long-term application, PRB surface not only undergoes geochemical changes, but also experiences an increased microbial activities which biologically mediate the dehalogenation [192]. The Monkstown ZVI-PRB in Northern Ireland, the oldest commercially installed ZVI-PRB in Europe, has been used to remove TCE for nearly 20 years. According to the data from 2001 to 2006, TCE was still being reduced effectively, and mineral precipitation and great variety of microbial communities were detected [193].

Table of contents :

Chapter 1. Literature Review
1.1 Contamination in soils and sediments by improper disposal of e-waste
1.1.1 Hazard and harm
1.1.2 Soil and sediment contamination by improper disposal of e-waste
1.2 PCBs in soils and sediments
1.2.1 Chemical and physical properties of PCBs
1.2.2 Toxicity and bioaccumulation of PCBs
1.2.3 Environmental behavior of PCBs in soils and sediments
1.2.4 Contamination by PCBs in soils and sediments
1.3 Remediation trials on e-waste contaminated soils and complex co-contaminated soils
1.3.1 Phytoremediation
1.3.2 Physico-chemical remediation
1.3.3 Removal of multiple contaminants by soil washing
1.4 Zerovalent iron (ZVI) mediated reduction
1.4.1 Factors influencing ZVI degradation
1.4.2 Mechanisms of halogenated hydrocarbon degradation by zerovalent iron
1.4.3 Application of ZVI degradation in remediation
1.5 Conclusions
Chapter 2. Soil contamination by polychlorinated biphenyls in e-waste recycling sites in south China
2.2 Experimental
2.2.1 Studied areas and sample collection
2.2.2 Characterization of soil samples
2.2.3 Analysis of heavy metals
2.2.4 Analysis of PCBs and QA/QC
2.3 Results and discussion
2.3.1 Properties of sampled soils
2.3.2 PCB contaminations
2.3.3 Metal contaminations
2.4 Conclusions
Chapter 3. The effect of Tween-80, humic acid and metal ions on the dechlorination of PCBs by ZVI
3.1 Introduction
3.2 Experimental
3.2.1 Chemicals
3.2.2 Dechlorination experiments
3.2.3 PCBs extraction and detection
3.2.4 Data analysis
3.3 Results and discussion
3.3.1 Kinetics
3.3.2 Dechlorination pathway
3.4 Conclusions
Chapter 4. Factors and processes influencing dechlorination of PCBs by ZVI in aqueous solution
4.1 Introduction
4.2 Experimental
4.2.1 Chemicals
4.2.2 Orthogonal experiments
4.2.3 Extraction and detection of PCB-52
4.2.4 Surfactant detection
4.2.5 Characterizations of ZVI
4.3 Results and discussion
4.3.1 Mass transfer
4.3.2 Iron oxidation
4.4 Conclusions
4.5 Supplementary Information
4.5.1 Data analysis
4.5.2 SEM-EDX
4.5.3 CMC of Envirosurf
Chapter 5. Surfactant-assisted PCBs reduction by ZVI in contaminated sediment
5.1 Introduction
5.2 Experimental
5.2.1 Chemicals
5.2.2 Contaminated sediment
5.2.3 Batch experiments of approach A
5.2.4 Batch experiments of approach B
5.2.5 Analysis and characterization
5.3 Results and discussions
5.3.1 Approach A
5.3.2 Approach B
5.4 Conclusions
5.5 Supplementary Information
5.5.1 Desorption kinetics
5.5.2 Details of contaminated sediments
Chapter 6. Environmental Implications, conclusions and perspectives
6.1 Environmental implications
6.1.1 The issue
6.1.2 Feasibility
6.1.3 Lab-scale application
6.2 Conclusions
6.3 Deficiencies and perspectives
Résumé étendu en français
Reference

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