Applications of carbonaceous adsorbents in the remediation of polycyclic aromatic hydrocarbon-contaminated sediments: A review

Journal Pre-proof Applications of carbonaceous adsorbents in remediation of polycyclic aromatic hydrocarbons -contaminated sediments: A review

Fang Li, Jianjun Chen, Xin Hu, Feng He, Eban Bean, Daniel C.W. Tsang, Yong Sik Ok, Bin Gao PII:

S0959-6526(20)30310-3

DOI:

https://doi.org/10.1016/j.jclepro.2020.120263

Reference:

JCLP 120263

To appear in:

Journal of Cleaner Production

Received Date:

01 October 2019

Accepted Date:

24 January 2020

Please cite this article as: Fang Li, Jianjun Chen, Xin Hu, Feng He, Eban Bean, Daniel C.W. Tsang, Yong Sik Ok, Bin Gao, Applications of carbonaceous adsorbents in remediation of polycyclic aromatic hydrocarbons -contaminated sediments: A review, Journal of Cleaner Production (2020), https://doi.org/10.1016/j.jclepro.2020.120263

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Journal Pre-proof Amount of words：11909 Applications of carbonaceous adsorbents in remediation of polycyclic aromatic hydrocarbons -contaminated sediments: A review

Fang Li a,b, Jianjun Chenc, Xin Hud, Feng Hee, Eban Beanb, Daniel C.W. Tsangf, Yong Sik Okg, Bin Gaob,*

a College

of economics and management, Shandong Agricultural University, Tai'an 271018,

China bDepartment

of Agricultural and Biological Engineering, University of Florida, Gainesville,

FL, USA c Mid-Florida d

Research & Education Center, University of Florida, Apopka, FL 32703, USA

Center of Material Analysis, Nanjing University, Nanjing 210093, PR China

e College

of Environment, Zhejiang University of Technology, Hangzhou 310014, China

f Department

of Civil and Environmental Engineering, The Hong Kong Polytechnic University,

Hong Kong, China g Korea

Biochar Research Centre & Division of Environmental Science and Ecological

Engineering, Korea University, Seoul, South Korea

*Corresponding

author, phone: (352) 392-1864 ext. 285, Fax: (352) 392-4092, email:

[email protected] 1

Journal Pre-proof Abstract Polycyclic aromatic hydrocarbons (PAHs) are widely accumulated in sediments and thus impose great risks to the ecosystem and public health. There is increasing effort on the development of technologies for remediation of PAH-contaminated sediments. Adsorption is one of the most promising remediation technologies is to PAH-contaminated sediments. Carbonaceous adsorbents (CAs) have been considered desirable for adsorbing PAHs due to their cost-effective and eco-friendly nature. The effectiveness of CAs in removal of PAHs, however, can be greatly affected by adsorbent properties, sediment characteristics, and environmental conditions. An accumulating evidence largely from laboratory experiments has shown that CAs can be used for immobilization and decontamination of PAHs in sediments, however, their full potential has not been demonstrated, particularly with respect to field applications. The present review aims to summarize the current progress on CA physical and chemical properties, and mechanisms of the stabilization of PAHs in sediments as well as key factors controlling field applications. In addition, potential environmental effects associated with CA-mediated PAH stabilization is also discussed. Among various CAs, activated carbon and biochar are the most commonly used adsorbents for sediment remediation and thus are the focuses of this review. Keywords: Biochar; Activated Carbon; Carbon Nanoparticles; Sediment Remediation; Organic Pollutants

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Journal Pre-proof 1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are a group of ubiquitous, highly toxic and persistent pollutants that are comprised of at least two linear, angular or cluster fused aromatic rings. PAHs are formed from either natural or manmade combustion sources and ultimately sink in the soils and sediments, which pose a serious threat to both human health and the environment because of their well-recognized carcinogenic, teratogenetic, and genotoxic characteristics (Boffetta et al., 1997; Li et al., 2017). PAH contamination in sediments can be abatable by physical, chemical, thermal and biological technologies, traditionally including in situ cappings, incineration, thermal conduction, solvent extraction/soil washing, chemical oxidation, bioaugmentation, biostimulation, phytoremediation, composting/biopiles, and bioreactors (Kuppusamy et al., 2017). These methods are largely focused on removal of PAHs, which either are rather costly and difficult to realize in situ remediation or require a long remediation cycle. Another approach is to reduce the bioavailability of PAHs, minimizing their effects to human health and the environment. The sorption method is considered to be an economical and environmentalfriendly technique to reduce the bioavailability of PAHs (Funada et al., 2018; Yin et al., 2019). PAHs are released to the atmosphere and enter into the water, soil, or sediment through wet and dry deposition processes controlled by precipitation, turbulence, gravity and other factors (Bandowe et al., 2018; Dat and Chang, 2017; Zhang et al., 2019). PAHs in soils or sediments tend to be sorbed on the particles. In 1992, Weissenfels demonstrated that PAHs are tightly adsorbed onto soil organic matter to form persistent micropollutants, due to their highly hydrophobic feature (Weissenfels et al., 1992). Such adsorbed PAHs are generally non-bioavailable and non-biodegradable (Weissenfels et al., 1992). In 1995, Wild and Jones presented that soil is the ultimate sink of PAHs, with more than 90% of the UK 3

Journal Pre-proof environmental PAH burden present in soils, while about 5.3% is associated with freshwater sediment (Wild and Jones, 1995). The sources, fluxes and sinks of PAHs and their effects on flora, fauna, and environment are summarized in Fig. 1. Due to their large specific surface area (SSA), rich porous structure, and high adsorption capacity, carbonaceous adsorbents (CAs) have been widely used in organic pollutant sorption (Zhang et al., 2017). A variety of CAs such as activated carbon (AC), biochar, carbon nanotube, and their derivatives have shown good effects in laboratory experiments. Among them, AC and biochar are the most commonly used CAs in sediment remediation due to the dual consideration of environmental and economic value (Beesley et al., 2011; Gomez-Eyles et al., 2013; Oleszczuk, Patryk et al., 2012). The in-situ capping technique using CAs is a frequently used engineering measure to treat aqueous sediments contamination, but this method has strict requirement to the environmental and functional condition of the plot (Murphy et al., 2006; Zimmerman et al., 2005). To emphasize the adsorptive function of the CAs and the interactions between PAHs and the adsorbent, this review mainly focuses on the applications of CAs as an adsorbent to immobilize PAHs and to reduce their impacts in sediments. In geology, soil can be considered as a collection of various sediments, generated from parent rock through the interactions of physical, chemical, and biological weathering processes and transported by the natural forces. Since 1900, there have been many studies on the applications of CAs to water and sediments/soils for PAH sorption/remediation. Fig. 2 shows an increasing trend on the investigation of PAH remediation through the uses of CAs (data from Web of Science database). Until now, a number of papers have examined PAH remediation in sediments by the sorption method, particularly at the laboratory scale (Maletic et al., 2019; Reis et al., 2007; Samuelsson et al., 2015). Most of them are concentrating on the adsorptive removal of PAHs from sediments, conducted in laboratory under ideal conditions with the participation 4

Journal Pre-proof of aqueous solutions (Dabrowski et al., 2005; Lamichhane et al., 2016; Smol and Włodarczyk-Makuła, 2016). Choi et al. built a decision-support framework to assess the effectiveness of AC in sediment remediation with considerations of following factors: sitespecific kinetic coefficients, AC dosage and particle size, sediment and AC sorption parameters, and pore-water velocity (Choi et al., 2016a). The importance of sorption process has been highlighted in several studies of the application of AC in the remediation of PAHcontaminated sediments (Choi et al., 2016b; Ghosh et al., 2011; Lin et al., 2015; Thompson et al., 2016). Several reviews have paid attention to the remediation of PAHs in sediments (Gan et al., 2009; Kuppusamy et al., 2017; Lau et al., 2014; Maletic et al., 2019). However, only few have attempted to comprehensively evaluate the mechanisms and the controlling factors of the remediation of PAH-contaminated sediments by CAs, and determine the connections between laboratory test results and filed engineering applications. Many laboratory studies have been published in this field, while the urgent need of sediment remediation prompts demands in summarizing efficient methods for filed applications. The overarching goal of this review, therefore, is to emphasize the current state of the knowledge on how the influencing factors act on the effectiveness of the treatment, to point out the directions for possible future investigations by optimization the key factors controlling or hindering the applications of CAs in the remediation of PAHs in sediments. 2. Adsorption mechanisms of PAHs on CAs For a long time, carbon-based adsorbents (e.g., black carbon, activated carbon, biochar, carbon nano-tube, graphene, and their derivatives) have been widely used for abating organic pollutants, due to their abundant porous structure and large surface area (Tascón, 2012). Organic pollutants adsorbed onto carbon absorbent can even be visualized by scanning electron microscope imagery (Fig. 3). Hence, CAs, such as activated carbon (AC) and 5

Journal Pre-proof biochar have been extensively used for removing PAHs from aqueous solutions or for immobilizing PAHs in contaminated sediments (Lamichhane et al., 2016). Thus a firm understanding of the adsorption mechanisms of PAHs onto carbon materials is critical for remediation of contaminated sediments. The mechanisms underlying the adsorption of PAHs on CAs can be categorized into relatively weak adsorption due to the hydrophobic interaction, Van der Waals force and/or hydrogen bond, and relatively strong adsorption, mainly through π-π interaction or π-electron from the rich region on CAs, electrostatic interaction and/or π complexation (Keiluweit and Kleber, 2009). These adsorption mechanisms of PAHs on CAs are illustrated in Fig. 4. Carbon adsorbents with stronger molecular polarity may have higher PAH adsorption ability, due to the intermolecular dipole-dipole interactions. Hydrophobic interaction can be predicted using these parameters, such as octane-water distribution coefficient (Kow) (Goss and Schwarzenbach, 2001). Van der Waals forces (VDW) are weak intermolecular forces that can facilitate the adsorption process. It was found that the VDW play an important role in stabilizing the adsorbed pyrene (Pyr) molecule onto multiwalled carbon nanotubes (MWCNTs) (Paszkiewicz et al., 2018). The small interaction energy consistent with the observed low thermodynamic stability Pyr-MWCNT → Pyr + MWCNT process (ΔGr 298 = 2.79 kcal/mol). PAH derivatives contain hydroxyl groups (-OH), carboxyl groups (-COOH), and amino groups (-NH2), which can also be adsorbed on the CAs through a hydrogen bond. Multilayer adsorption of these PAH derivatives also can be achieved by hydrogen bonds between surface-adsorbed molecular and the dissolved ones (Lin and Xing, 2008). The stable aromatic rings in PAHs contain several π electrons, and are easy to conjugate with aromatic rings or π bonds (e.g. C=C double bonds) on CAs. π-π electron-donor−acceptor (EDA) interaction is an important interaction between PAHs and CAs (Keiluweit and Kleber, 6

Journal Pre-proof 2009; Zhang et al., 2014). High molecular weight (HMW) PAHs have been shown to be more readily adsorbed on CAs (π-acceptor/ contain aromatic rings, C=C, etc.), due to a large number of π electrons (Sun et al., 2013). Except for π-π interaction, nitrogen and oxygen atom of adsorbent molecules can provide lone pair electron and vacant orbitals, and form n-π interaction with the π electrons in PAHs (Zhang, K. et al., 2018). The adsorption of hydroxyland amino-substituted aromatic chemicals can be enhanced by nitrogen doped multiwall carbon nanotubes (N-MCNT) (Zuo et al., 2016). The enhanced adsorption of the aromatic derivative to N-MCNT was mainly attributed to the favored n-π interaction between the adsorbate molecule (π-donor) and the N-heterocyclic aromatic ring (π-acceptor) on NMCNT. The adsorption through π complexation between adsorbent and metal or metal cation relies on weak chemical bonds. Transferring electrons between the metal ions presumably occurs through the conjugated π-electron system of the bridging group (Halpern and Orgel, 1960). Heterocyclic molecules or functional groups containing metal ions can accept the delocalized π electron provided by aromatic rings of PAHs and thus form π complexation bond. Electrostatic adsorption for the ionizable organic chemical can be achieved through the variation of pH. Low adsorption would be expected at pH (solution pH)> pKa (dissociation constant of PAH derivatives) and pHpzc (point of zero charges of the CAs), because of electrostatic repulsion between negatively charged PAH derivatives and CAs (Pan and Xing, 2008). On the contrary, pHpzc> pH >pKa would show high adsorption for the electrostatic attraction.

7

Journal Pre-proof 3. Effects of carbon properties on sediment remediation 3.1 Morphological characteristics of CAs SSA and pore size distribution of CAs significantly affect their adsorption of PAHs and thus determine the effectiveness of sediment remediation. In general, a proper pore size distribution offers mass transfer paths for PAHs and large surface areas provide more active sites for adsorption, thus enhancing the sorption capacity and sorption rate. However, surface functional group of CAs and chemical characteristics of the adsorbents play crucial roles in remediation of specific PAHs. Zhang et al. compared removal efficiencies of organic contaminants from bio-treated coking wastewater by two ACs and found that large SSA of CAs can improve their remediation performance (Zhang, C. et al., 2018). Similarly, CAs show the same trend for reducing the PAH porewater concentration in sediments (Choi et al., 2013; El-Shahawi et al., 2017). The surface area of CAs is related to the feedstocks, production processes, and modification processes (de Jesus et al., 2019; Kumar et al., 2019). Common modification methods for increasing carbon surface areas include treatments with alkali, heat, acid, microwave, ozone, and plasma (Xiao et al., 2015; Zhang et al., 2017). Regarding the PAH presence in sediment, Choi et al. indicated that the particle size of AC is one of the key factors for the reduction of the sediment PAHs pore-water concentrations (Choi et al., 2014), which may be related to surface area. Table 1 summarizes the performance of different CAs in aqueous solutions from the reviewed literature and a comparison of different CAs in removal of PAHs in from aqueous solutions. A large SSA, however, is not always correlated with a high adsorption capacity. In comparing with AC, carbon nanotubes have a smaller surface area but a higher adsorption capacity (Table 1). This is mainly due to the different pore geometries (Pan and Xing, 2008). Generally, micropores (pore diameter < 2nm) on the CAs provide principal sorption sites, 8

Journal Pre-proof mesopores (2 nm < pore diameter < 50 nm) provide both active sorption sites and intraparticle diffusion pathways, while macropores (pore diameter > 50 nm) mainly affect diffusion pathways (Nybom, 2015). AC possesses a similar pore volume as DMSCNs but much larger surface area and smaller pore size (<1 nm) than DMSCNs (comparing row 2 with row 3 in Table 1), which significantly restricts the remediation of Ant. Similarly, pore volume and pore size distribution of CAs can be substantially influenced by preparation conditions, preparation methods, and modifications as well as the nature of stock materials (Muniandy et al., 2014; Yakout and Sharaf El-Deen, 2016). Besides the adsorption capacity, the adsorption rate which is generally evaluated based on the adsorption kinetic constant is another important parameter used for assessing the performance of CAs in PAH decontamination. For example, rape straw porous carbon (RC) with a pore diameter of 2.25 nm and total pore volume of 0.72 cm3/g and corn cob porous carbon (CC) with a pore diameter 2.04 nm and total pore volume 0.55 cm3/g were fabricated in the same way (Cheng et al., 2019). The RC contained micropores, mesopores and macropores, while the CC pore distribution was concentrated in the range of micropores. The adsorption equilibrium capacities of naphthalene (Nap) on RC and CC were similar (370 and 340 mg/g, respectively), while the pseudo-first-order adsorption rate constants were 13.332 and 1.412 (/h), respectively (Cheng et al., 2019). The existence of mesopores and macropores on RC led to a fast PAH adsorption rate, which is primarily due to the excellent mass transfer routes provided by the large pores (de Jesus et al., 2019; Yahya et al., 2015). What’s more, the small pore size, narrow passage of CAs and long crossing channel in sediment, hinder efficient mass transfer and diffusion of PAH molecules to utilize all available adsorption sites. In addition, the impact of large pore size distribution on CA performance has been magnified under the influence of sediment. Pignatello, Kwon and Lu provided clear evidence of the blocking of micropores due to the condensation of humic 9

Journal Pre-proof substances on black carbon (BC) surfaces (Pignatello et al., 2006). Large molecule organic materials in sediment cannot penetrate deeply into pore networks and may be continuous or localized in clumps on the external adsorbent particle surface. The condensed organic strands will occupy the throats of many micropores open to the external surface. Narrower pores are clogged more easily by natural organic matter in sediment than wider pores (Jośko et al., 2013). The adsorbents with more micropores will face larger degree of attenuation of target pollutants sorption capacity in sediment than in aqueous solution. Future methods to efficiently enhance CA surface area, total pore volume, and optimize pore-size distributions promise new developments.

3.2 Chemical characteristics of CAs Chemical characteristics of the carbon materials also play important roles in controlling their applications in the remediation of PAH-contaminated sediments. To enhance PAH adsorption onto CAs, appropriate modifications of the adsorbents are needed to either add or eliminate the surface functional groups (Koltowski et al., 2017). The heteroatoms (mainly including oxygen, nitrogen, hydrogen, sulfur, phosphorus, and halogens) in surface functional groups are major determinants of adsorbents chemical properties, which can be determined by elemental analyzer, scanning electron microscope-energy dispersive spectrometer, and the H-O, C=C, C-O chemical bond of CAs can be analyzed by Fourier-transform infrared spectroscopy and nuclear magnetic resonance spectroscopy (Wang and Wang, 2019). Low H/C and O/C ratios indicate high aromatization and high hydrophobicity (Yahya et al., 2015). The degree of graphitization, defect status of carbonaceous materials can be identified by Raman spectra, X-ray diffraction analysis. (Leng et al., 2019; Ouyang et al., 2019). It has been reported that high carbon content is an important parameter to enhance the Ant adsorption on to DMSCNs (Yang et al., 2014). The oxygen-containing functional group 10

Journal Pre-proof (carboxylic, lactonic) mainly presents acidic in the liquid phase, for the reason of hydroxyl and carbonyl groups generation. Oxygen-containing functional groups on CAs provide the possibility of hydrogen bond with PAH derivatives. Meanwhile, the high oxygen content decreases the hydrophobicity of CAs, and inhibits the remediation of PAHs. The nitrogencontaining functional group is likely to form π complexation adsorption under the bridging effect of metal or metal cation. The existential form and the content of the functional groups in CAs depend on the precursor and the activation treatment. For instances, thermal treatment or alkali modification are the common methods to eliminate the oxygen-containing groups, while ozone and acid modification are two common oxygen introduction methods for CAs (Wang and Wang, 2019).

3.3 Carbonaceous adsorbent dosage The reduction of the chemical and biological availability of PAHs in sediment pore water depends on the repartition between sediment and added adsorbents. Hence, the adsorbent dose or the sediment to adsorbent ratio are controlling factors for PAH removal within limits. This results because the obstructive factors, like natural organic carbon fouling and competing adsorbing of ionic substance can be counteracted by increasing the quantities of CAs. The PAHs mass reduction measured by semipermeable membrane devices increased from 44% to 87% after sorbent dose increased from 0.34% to 3.4% (Zimmerman et al., 2005). The pore water concentrations reduction of Flu, Phe, and Ant are proportional to the dose of adsorbents, when dosing with less than 4% adsorbents. When the amount of adsorbent added exceeds a certain range, the removal efficiency of PAHs is not significantly improved (Rakowska et al., 2013). Many laboratory study results have demonstrated that a dosage range of 2% to 4% AC is sufficiently responsible for more than 90% reduction of PAH pore water concentration in well-mixed sediments (Brandli et al., 2008; Cho et al., 11

Journal Pre-proof 2012; Cornelissen et al., 2006b; Hale et al., 2012; Kupryianchyk et al., 2012a; Kupryianchyk et al., 2011; Rakowska et al., 2012). Although a number of laboratory studies have demonstrated PAH bioavailability reductions in sediment, no field study has been conducted in the sediment phase. Through comparing with the few references to polychlorinated biphenyls (PCBs), the minimum dosage of AC to effectively reduce contaminant bioavailability in the field is generally consistent with the dosage in the laboratory (Abel and Akkanen, 2019; Kupryianchyk et al., 2012a; Zimmerman et al., 2004).

4. Effects of PAH properties on sediment remediation Adsorbate molecular sizes and shapes affect their removal efficiencies by CAs and thus are critical to the remediation of contaminated sediments. The mass transfer from sediments to adsorbent particles can be faster for the smaller PAHs than that for the larger ones. This is because since smaller PAHs are generally more polar, they can quickly access the adsorption site of CAs (Abdelrady et al., 2019). In addition, HMW PAHs are retarded by stronger sorption to pore walls that inhibit their desorption from the sediment matrix and diffusion into the adsorbent pores (Hale et al., 2012). Several studies have shown that light molecular weight (LMW) PAHs respond more readily to CA amendment due to faster kinetics in the sediment desorption process and carbon uptake, while HMW PAHs in sediment are more likely to condense on natural organic matters, and need stronger desorption energy to dissolve into pore water (Choi et al., 2013; Hale and Werner, 2010; Zimmerman et al., 2004). Further, as the molar volume increases, the penetration of PAHs into the micropores of adsorbents becomes difficult. Generally, LMW PAHs can easily access the micropores and bind to adsorption sites, while HMW PAHs may not enter into the micropore, or block the entrance of narrow micropores and decrease intra-particle diffusion. Biochar fabricated from enteromorpha prolifera with optimal PAH remediation efficiency in the sorption of Pyr and 12

Journal Pre-proof benzo[a]pyrene (BaP) from aqueous solution is characterized with large surface area (205 m2/g), high adsorption rate during the first two hours and high equilibrium adsorption capacity (187.27 μg/g and 80.00 μg/g for Pyr and BaP, respectively) (Qiao et al., 2018). In addition to the molecule size, the spatial configuration of the PAH molecules also affects the availability of carbon adsorption sites in sediment remediation. Straight-chain or planar PAHs are often adsorbed better than curved-chain or non-planar PAHs (Liphard et al., 1980; Vollhardt et al., 1993). PAHs released from contaminated sediments into solutions can be completely removed by CAs. Granular AC (GAC, SSA 1218 m2/g, pore volume 0.53-0.6 cm3/g, pore size 3.2-3.4 nm) shows different equilibrium times (4–5 h for Nap, 5–6 h for Acenaphthylene (Acy), 8 h for Acenaphthene (Ace), 16 h for Fluorene (Flu) and 24 h for Phenanthrene (Phe)) for PAHs in solutions (Eeshwarasinghe et al., 2018). The PAH removal efficiencies of Flu, Phe, Ant, Pyr and benzo(a)anthracene (B[a]A) from a contaminated urban soil are 96, 92.4, 90.9, 79.6, and 64.6%, respectively, after 6 week shaking extractions by GAC (SSA 1050–1200 m2/g) (Brandli et al., 2008). The remediation efficiencies of linear molecules are significantly higher than that of non-linear molecules, both in aqueous solutions and in sediments (Gotovac et al., 2007). The molecular structures of PAHs commonly used in the literature are summarized in Table 2. Molecule hydrophobicity and weight of PAHs can also affect sediment remediation efficiencies of CAs. Generally, high hydrophobicity PAHs are more likely to adsorb onto adsorbents without polar functional groups, due to the dipole-dipole interactions between PAHs and adsorbents as well as VDW forces among PAH molecules. The sorption Sips isothermal parameters ks and n of Nap, Ace, and Phe onto two kinds of porous carbon are in the order of Nap < Ace < Phe, and in line with the order of their molecule hydrophobicity (Cheng et al., 2019). The molecule hydrophobicity was directly demonstrated to influence the Nap, Phe and Pyr adsorption on MCM-41, due to no hydrogen bonds or other strong polar 13

Journal Pre-proof forces present (Yang, X. et al., 2019). The contribution of the hydrophobic effect is 50%-85% for Phe, while the contribution is lower than 30% for 9-phenanthrol and 9, 10-phenanthrene quinone on functionalized MWCNTs, due to the higher hydrophobicity of Phe (Peng et al., 2017). The adsorption efficiency decreases with the increase of PAH molecular weight because of the relative low diffusion coefficients of higher molecular weight compounds (Zhang et al., 2011). The octanol-water partition coefficient (Kow) and weight of PAHs commonly used in the literature are shown in Table 2.

5. Effects of environmental conditions on sediment remediation BC or hard carbon has sorption coefficients of PAHs in the sediments up to one to two orders of magnitude higher than amorphous organic matter (Rakowska et al., 2012). Due to the complexity of sediments, the remediation of PAHs by CAs is greatly affected by the environmental conditions of contaminated sediments.

5.1 Initial PAH concentration The removal efficiency of sorbents is directly correlated with the level of contaminants in sediments (Caniani et al., 2018). In a moderately contaminated urban soil, the freely dissolved aqueous concentration of native PAHs in a soil/water suspensions was reduced from 64% to 99% by adding small amounts of powder AC (PAC) and GAC (2%) (Brandli et al., 2008). While only median reductions were achieved for a heavily contaminated creosote soil (ranged from 4% to 63%) due to saturation of AC sorption sites by the high levels of PAHs (Brandli et al., 2008). On the contrary, the negative influence of initially high PAH concentrations for CA performance in sediment amendments can be counteracted by increased dosage of adsorbents. A 5% amendment of AC to well-mixed petroleum-impacted sediments slurry with total PAHs contents at 125 mg/kg or 2,550 mg/kg resulted in 99% or 14

Journal Pre-proof 98% reduction of PAHs (Choi et al., 2013). Additionally, the initial PAH concentration also contributed to the amendment rate. For example, available fractions for fast desorption rates of Phe from the soil are 79%, 46%, 40%, 39%, and 35% for initial Phe concentrations of 100, 400, 800, 1200, and 1600 mg/kg respectively (Gharibzadeh et al., 2019).

5. 2 pH and ionic strength Variation in sediment pH can result in a chemical speciation change of the natural organic materials and the added-CAs. Generally, increased pH can lead to increased ionization and polarity by changing the functional groups, therefore, decreasing the remediation performance for PAHs (Qiao et al., 2018; Sun et al., 2013). pH and ionic strength together play complicated roles in affecting PAH desorption from sediment and adsorption onto CAs. Electron donor-acceptor interaction and bridging interaction of metal ions are the dominant mechanisms that pH and ionic strength are attribute to influencing PAH adsorption and desorption. The CAs would show high PAH adsorption from aqueous solution when pHPZC (point of zero charges of CAs)>pH (aqueous solution pH)>pKa (ionization equilibrium constant of PAHs), under the condition of same ionic strength (Zhou et al., 2013). No consistent results have been obtained for the ionic strength and pH influence on HOC-DOM (hydrophobic organic contaminants-dissolved organic matter) interactions (Pan et al., 2008). Ionic strength is a complicated factor that has opposite effects (negative or positive relationships) on KDOC (carbon-normalized sorption coefficients) (Jones and Tiller, 1999; Marschner et al., 2004). In addition, the ion categories also have effects on the mechanisms for adsorption (Chen et al., 2007). Metal ions may enhance the adsorption efficiency for PAHs by bridging interaction, such as PAH salting-out effect. On the contrary, the large hydration shell of metal cations may intrude or shield the hydrophobic sites and indirectly compete with PAHs for surface sites, leading to the inhibition of remediation (Chen et al., 15

Journal Pre-proof 2008; Oleszczuk, P. et al., 2012; Wang et al., 2015). The Phe degradation was enhanced by 1.9 times, with the addition of nitrate in the biochar-river sediment system, due to the electron acceptor role that nitrate played (Yang et al., 2018). The binding coefficient for Phe is not independent of pH and ionic strength, it also depends on the dissolved organic matter concentration (Pan et al., 2007). However, many questions remain unanswered and further research on how the pH and ionic strength alter the CA removal efficiency for PAHs, especially in sediments, is warranted.

5. 3 Natural organic matters Previous studies have well demonstrated that natural organic matter of sediment plays a key role in governing the remediation of PAHs by CAs (Lamichhane et al., 2016). Limited mobility of organic contaminants in the highly polluted San Jacinto superfund site (waste pit) in the Houston Ship Channel, Texas, suggests that the strong sorbing pyrogenic and petrogenic residues in the waste pit keeps PAHs strongly sorbed to particles (Louchouarn et al., 2018). It further supports that aromatic domains of total organic matter in sediments is the main sorption sites for hydrophobic organic contaminant (Choi et al., 2016c). A “dual-mode” sorption paradigm was proposed initially by Huang et al. (1997) where the soft, amorphous materials of organic matter showed linear and noncompetitive absorption, while the hard, condensed, and aromatic materials showed nonlinear, extensive, and competitive adsorption (Cornelissen et al., 2006a). The PAH sorption coefficient based on amorphous organic carbon partitioning alone is 1-2 orders of magnitude above predicted result by total organic matter (Brandli et al., 2008; Cornelissen et al., 2006a; Lasota and Blonska, 2018). To determine the divergent binding capabilities and dissimilar binding strengths of PAHs with different organic matter fractions, soil organic matter fractions including fulvic acids, humic acids, and humin and BC in relation to PAH accumulation levels in agricultural soils were evaluated (Ukalska-Jaruga et al., 2019). 16

Journal Pre-proof The correlation coefficients between total organic matter, humin or BC with PAHs are 0.73, 0.71, and 0.86, respectively. Furthermore, the BC faction showed the highest correlations with (HMW) PAH (r = 0.92). The CAs amended to sediment may be fouled by dissolved organic matters or colloidal matters like amorphous organic carbon, mineral particles, oil, and weathered oil residues, due to site competition and pore blockage (Cornelissen and Gustafsson, 2006). Pyr sorption to biochar can be attenuated by up to 80% in the presence of soils and composts (Kah et al., 2018). Until now, most of the reported CAs (AC (Rasheed et al., 2015), porous carbon (Cheng et al., 2019; Yuan et al., 2010), carbon nanotubes (Paszkiewicz et al., 2018) and graphene (Huang et al., 2018; Wang et al., 2014) show relatively high adsorption kinetics and removal rates for PAHs. However, sediment PAH decontamination has significantly lower removal efficiency and adsorption kinetics (days to months vs. hours to days) than aqueous PAH decontamination (Nybom, 2015; Patmont et al., 2015; Rakowska et al., 2014a). Funada et al. (2018) compared the removal efficiency of a magnetic AC to 12 PAHs in glass beads and road-side sedimentsuna. The removal rate was up to 100% after 3h of remediation in the glass beads system, while the removal rates were just 54% and 62% after 7 days and 14 days, respectively, of remediation in a road-side sediment system. The low adsorption kinetics of PAHs from sediment can be attributed to the following: PAHs desorb out of the sediment matrix, dissolve in pore water, mass transfer to the CA-pore water interface and diffuses inside the particle until it reaches the adsorption site (Choi et al., 2013; Hale et al., 2011). In general, it takes a longer time to release bound PAHs from the sediment particles, due to the inhomogeneous desorption energy from the heterogeneity of sediments. Choi et al. (2013) revealed the kinetics of contaminant desorption was a more significant factor than the pollutants’ concentration for the performance of AC treatment of PAH-contaminated sediments. 17

Journal Pre-proof In order to quantify the impact of natural organic matter, Brandli et al. (2008) proposed a prediction model (Eq. (1) of attenuation of AC sorption in the presence of sediments. n

AC CS  fTOC  KTOC  CW , free  f AC  K F , AC  CWf,,free

(1)

n AC where CS , CW , free and CWf,,free are the PAHs total soil contents, freely dissolved

concentration on the basis of pure AC sorption and the amended sample, respectively. fTOC and f AC are the content of total organic carbon and AC, respectively. KTOC and K F , AC are the organic carbon–water distribution coefficient determined in the non-amended system and Freundlich coefficients for sorption of Phe to AC, respectively. By comparing the apparent sorption coefficients determined in a low-density polyethylene-water system and a low-density polyethylene-water-sediment system, it was demonstrated that the presence of dissolved organic matters or colloidal matters interferes with PAH sorption onto sorbents (Choi et al., 2016c; Devault and Gourlay-Francé, 2017).

5.4 Contact time Extraction quantities of PAHs from sediments by CAs are strongly affected by the extraction time (El-Shahawi et al., 2017; Wu et al., 2017). AC moderation of hydrophobic organic contaminant sequestration effectively increased over time for the progressive mass transfer in sediments (Patmont et al., 2015). Oleszczuk et al (2017) reported that the most rapid reductions of bioavailable PAHs in soils may take 6 months after the addition of the AC or biochar, compared with unamended groups. Hale et al. (2012) conducted a long-term AC amendment study for 28 months. They found PAC and GAC amendment resulted in 93% and 56%, respectively, reduction of PAH concentrations in soil drainage water after 12 months of amendment, but the reduction decreased to 67% and 17%, respectively, after 20 months (Hale et al., 2012). The free aqueous PAH concentration significantly declined from 93% and 84% 18

Journal Pre-proof at 17 months to 76% and 69% at 28 months (Hale et al., 2012). In another study, Choi et al. (2013) reported that total PAHs decreased from 98% at 1 month to 95% at 12 months, due to the replacement of a fraction of LMW PAHs adsorbed onto AC by HMW PAHs. In addition to this, other speculative explanations for the lack of sustained AC effectiveness including natural organic matter fouling and biodegradation retarded by the presence of AC over time, etc. (Hale et al., 2012; Oen et al., 2012). It is worth noting that long-term effectiveness of field-scale AC amendment in reduction of PCBs has been observed by passive samplers (73% reduction after 5 years, 3.7 wt.% AC dose addition) (Cho et al., 2012).

5.5 Mixing The mixing methods of adsorbent with contaminants in sediments can significantly affect the remediation efficiencies. Most lab studies have been conducted with the mixing of shaken end-over-end (Brandli et al., 2008; Cornelissen et al., 2006a; Hale et al., 2012), and free aqueous PAH reductions were largely evaluated in a continuously well-mixed CA treatment for several weeks (Brandli et al., 2008; Oleszczuk and Koltowski, 2017). Mixing methods have been shown to play an important role in the sediment PAH sequestration kinetics for CA amendments (Berardi et al., 2019; Werner et al., 2006). But it is difficult to achieve homogeneous distribution in large field scale application, especially for in-situ remediation. Lower PAH remediation efficiencies of AC amendment observed in field trials than measured in continuously well-mixed laboratory experiments is likely due to the limited contact between adsorbents and PAHs caused by the heterogeneity of sediments (Choi et al., 2014; Hale and Werner, 2010; Rakowska et al., 2014b). In-situ adsorbent treatments of PAH-contaminated sediments mainly involve capping and mixing (Cornelissen et al., 2012; Hale et al., 2012; Samuelsson et al., 2015). There are several reports on capping of marine sediments with AC (Cornelissen et al., 2012; 19

Journal Pre-proof Cornelissen et al., 2011; Samuelsson et al., 2015; Samuelsson et al., 2017). AC + clay capping can reduce PAH bioaccumulation in worms and clams by 40% and 67%, respectively (Samuelsson et al., 2015). However, these are far less effective than the results from laboratory studies, likely due to the homogenization of sediments, cap integrity, new contaminated particle deposition, and bioturbation (Cornelissen et al., 2012). More importantly, the capping layer may alter surroundings, as AC+clay capping has been suspected as the cause of reducing the abundance, biomass, and number of species of marine benthic fauna by up to 90% (Samuelsson et al., 2017). Adsorbents mechanically mixed into PAH-contaminated sediments can achieve better remediation efficiency due to the close contact between adsorbents and contaminants (Oleszczuk et al., 2019). However, there are still technical challenges in mixing devices that incorporating CAs into contaminated sediment. Cho et al. conducted a pilot-scale study to incorporate AC into a tidal mudflat plot using a shallow-draft barge with a rotovator attachment (Cho et al., 2007). The results showed a reduction of aqueous equilibrium PCB concentration in a field trial after 7 months and AC treatment was slightly lower than wellmixed (6 months on a roller), (62% vs. 77%). Barge-mounted rotovator system for direct mixing and crawler-mounted AC-slurry injector were innovatively used in the field trial at a tidal region, San Francisco (Cho et al., 2009). The results indicate that additional mixing during or after AC deployment for in-situ stabilization of PCBs will likely improve AC sediment contact and overall effectiveness. Although the mechanism and trend for removal of PAH and PCB from the sediment are similar (Luthy et al., 2004; Zimmerman et al., 2004), until now, only a few systematic studies of AC amendment to sequester PAHs in field-scale trails have been reported (Hale et al., 2012; Jakob et al., 2012).

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Journal Pre-proof 6. Potential environmental impacts and the recovery of CA amended 6.1 Potential environmental impacts of CAs CA amendments have proven to be an effective way to remove PAHs from sediments. A summary of previous studies on the effectiveness of CAs to reduce the PAH concentrations is presented in Table 3. It is clear that PAHs bind with CAs, which reduces their bioavailability (Bucheli and Gustafsson, 2001; Ghosh et al., 2003; Herrchen et al., 1996; Liu et al., 2019; Spacie and Hamelink, 1985). The decrease in biota-to-sediment accumulation coefficient after CA amendment can reach up to one to two orders of magnitude (Moermond et al., 2005; Rakowska et al., 2012; Zimmerman et al., 2004). For example, the PAH bioaccumulation in plants decreased by about 50% after the sediment was amended with 2% AC (Jakob et al., 2012). The differences of adsorbent characteristics, adsorbent dosage, particle size, contact time, mixing and many other factors can all affect the PAH bio-uptake (Beckingham and Ghosh, 2011). In addition to the decrease of PAH bioavailable concentrations, amended CAs may have some side effects to the sediment environment (Burgess et al., 2009; Kupryianchyk et al., 2011). As an exogenous additive, CA inevitably changes sediment characteristics. If this remediation technology is not used properly, it may bring negative effects to sediment biocenosis (Jonker et al., 2004). It has been reported that dissolved organic carbon in sediments can also be significantly decreased due to PAC and GAC amendment (Hale et al., 2012). The addition of CAs more or less affects the habitat quality, and is detrimental to sediment organisms. CA amendments can stimulate nitrifying bacteria in sediments to increase nitrification activities (Berglund et al., 2004; Gundale and DeLuca, 2007). Furthermore, the CA addition may affect the bioremediation of PAHs in sediments by changing the growth and activities of vegetation (Wu et al., 2018; Zhang and Fan, 2016), fauna (Rodriguez-Campos et al., 2014), and microbes (Azah et al., 2017). For example, the 21

Journal Pre-proof addition of biochar into sediments can enrich the population of fungi and microorganisms, due to the improvement of sediment permeability, water retention, and nutrient content (Lee et al., 2017; Song et al., 2017). Several reports have shown that when the dosage of AC or biochar is less than 5%, the amendment has little or no effect on the sediment environment (Janssen et al., 2012; Kupryianchyk et al., 2012a, b; Kupryianchyk et al., 2011). When the adsorbent dosage exceeds a threshold though, the positive effect of the amendment can be nullified by the inducing toxicity (Bielska et al., 2018). This toxic effect depends on various factors related to the CA properties including size, pore structure, surface area, dosage, and surface chemistry. For example, AC of different particle size has different biological toxicity (Hale et al., 2013; Jakob et al., 2012). Further investigations are needed to determine relationships between CA properties and their toxicity in the remediation of PAH-contaminated sediments.

6.2 CA recovery CA recovery is a means of eliminating the negative effects of added adsorbents and removed PAHs to avoid secondary release into the environment. One of the traditional ways of recovery is to separate the adsorbents from the sediment by sieving and centrifugation (Rakowska et al., 2014b). However, this method is costly, disturbs the soil, and the particle sizes of adsorbent and sediment must be distinguishing, which is practically infeasible. A novel magnetic CA has shown to be separated more effectively (Choi et al., 2016c; Funada et al., 2018; Han et al., 2015; Hao et al., 2018; Shahriman et al., 2018; Zhou et al., 2019). This adsorbent can be easily separated from the adsorption media using magnetism (Zhang et al., 2013). However, the low magnetic particle collection efficiency (50-60%), is a new challenge for the promotion of this new material (Choi et al., 2016c). A textile form of AC is another innovative way of separation and recovery. However, the textile form of activated carbon has 22

Journal Pre-proof exhibited low PAH sorption kinetics resulting from its thickness (Choi et al., 2016c). By comparison, the magnetized AC has as high of a PAH sorption coefficient as non-magnetized AC. 7. Conclusions and perspectives This review summarizes physical and chemical mechanisms of CAs used for PAH remediation in sediments and the major factors influencing remediation efficiencies, including the characteristics of CAs and PAHs, and remediation conditions. Furthermore, potential negative effects of CAs amended into PAH-polluted sediments exist and should be considered. The following conclusions can be drawn from the review: 1) CAs possess appropriate physical and chemical adsorption properties, such as large SSA and favorable pore size distribution, and compatible functional groups for adsorption of PAHs through intermolecular dipole-dipole interaction, VDW forces, hydrogen bonds, EDA interaction, π complexation, and electrostatic adsorption. 2) PAH morphological characteristics, including molecular size, spatial configuration, hydrophobicity, and weight, not only influence the adsorption capacity of CAs but also affect CA adsorption kinetics to a great extent. 3) The environmental conditions, including initial PAH concentration, pH value, ionic strength, natural organic matters, CA dosage, contact time, and mixing method influence PAH remediation. High initial PAH concentration can be counteracted by abundant dosage of adsorbents. pH value, ionic strength, and contact time are nonsignificant factors in the sediment remediation, while natural organic matters and mixing technology are the main barriers hindering the application of this method in the field. 4) The negative effects associated with CAs are a shortcoming of this sediment adsorption amendment. The induced biotoxicity can be controlled by strictly limiting adsorbent dosages. New adsorbent activation and modification methods to reduce the biotoxicity or enhance the recovery rate are needed further study. 23

Journal Pre-proof The complexity of sediment environments can significantly influence the adsorption efficiency in multiiple dimensions. To ensure that CAs are viable means for PAH remediation, further research is recommended: 1) New or modified carbon materials with high PAH adsorption capacity and low cost should be developed. 2) Methods for minimizing interference from natural organic carbon on CA application should be improved. 3) Strategies for increasing the adsorption kinetics in the sediment should be implemented. 4) New and innovative technologies for capping or mixing CAs should be applied to polluted sediments. 5) Technologies for reducing the buoyancy and increasing stability of CAs should be advanced. 6) Integrated approaches for decreasing the negative effect on sediment ecological environment should be developed for remediating PAHs while sustaining environmental quality.

Acknowledgments F.L. would like to acknowledge the support of the National Natural Science Foundation of China (Grant No. 41807124) and the China Scholarship Council (File No. 201809135012).

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Sorbents

to

Reduce

PCB

and

PAH

Bioavailability

in

Marine

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43

1

Tables

2

Table 1. Summary of the results of PAH removal by carbonaceous adsorbents from aqueous solutions Adsorbent

SSA (m2/g)

Pore volume (cm3/g)

Pore size (nm)

C%

N%

O%

Adsorbate

DMCNs

1218

2.10 and 18.40

-

404

-

AC

1202

<1

91(XPS) 73 (EA) 51(XPS) 35 (EA) -

-

DMSCNs

Total 2.23, Mesopore 1.48 Total 0.72, Mesopore 0.41 0.67

PC (rape straw)

1281

0.72

2.25

90.0

PC (corn cob)

1069

0.55

2.04

AC (a)

1156

0.65

AC (b)

1048

AC (c)

1003

AC (d) GNS

Ant

Capacity (Langmuir, mg/g) 948

Kinetic constant (PSO, g /mg·min) 3.50*10-9

-

Ant

348

1.70*10-7

-

-

Ant

113

-

1.6

8.4

86.8

1.4

11.8

-

96.8

0.7

2.0

Nap Ace Phe Nap Ace Phe Nap

525 448 808 358 357 656 563

1.85*10-3 8.17*10-4 1.00*10-4 1.17*10-4 9.83*10-4 1.83*10-4 -

0.56

-

86.4

0.5

12.5

Nap

378

-

0.53

-

86.9

1.3

11.3

Nap

382

-

843

0.45

-

79.7

1.1

18.6

Nap

185

-

392

-

-

76.2

5.2

17.6

Nap

71

2.24*10-3

14.10

GO

236

-

-

46.2

0.5

50.9

Nap

1

3.57*10-1

GAC

1218

0.53

3.40

-

-

-

AC

1770

0.99

2.82

-

-

-

AC (e)

258

0.21

1.99

60.3

0.7

35.0

Nap Acy Ace Flu Phe Nap Phe Pyr Nap

33 77 41 46 48 89 116 118 85

6.17*10-4 5.00*10-4 5.33*10-4 5.17*10-4 6.00*10-4 0.02 0.04 0.06 -

AC (f)

1580

0.81

1.47

83.7

0.5

14.1

Nap

300

-

Reference (Kalantari et al., 2019)

(Cheng et al., 2019)

(Ania et al., 2007)

(Wang et al., 2014) (Eeshwarasi nghe et al., 2018) (Xiao et al., 2015) (Cabal et al., 2009)

3

Note: Abbreviation: Specific surface area (SSA), Activated carbon (AC), Anthracene (Ant), Naphthalene (Nap), Acenaphthylene (Acy), Acenaphthene (Ace), Fluorene (Flu), Phenanthrene (Phe), Benzo(a)anthracene

4

(BaA), Benzo(b)fluoranthene(BbF), Benzo(k)fluoranthene (BkF), Dibenzo(a,h)anthracene (DahA), Dendritic mesoporous carbon nanoparticles (DMCNs), Dendritic mesoporous silica-carbon nanoparticles(DMSCNs),

5

Pseudo-second-order kinetic model (PSO), Porous carbons (PC), Graphene nanosheets (GNS), Graphene oxide nanosheets (GO)

44

Journal Pre-proof 6

Table 2. Characteristics of PAHs commonly used in the literature as target pollutants

7

(Awoyemi, 2011; Eeshwarasinghe et al., 2018; El-Shahawi et al., 2017; Walters and

8

Luthy, 1984; Yang, Xiong et al., 2019)

9 10

PAH

Molecule formula

Naphthalene (Nap) Acenaphthene (Ace)

C10H8 C12H10

Acenaphthylene (Acy)

Molecule structure

Weight (g/mol)

Log KOW (L/L) 3.30 3.92

Molar Molar volume dimensions

128 154

Aqueous solubility (25°C, mg/L) 31.7 3.9

148 173

9.1*7.3*3.8 9.1*8.3*4.2

Toxicity Equivalency Factor 0.001 0.001

C12H8

152

16.1

3.94

168*

-

0.001

Fluorene (Flu) Phenanthrene (Phe)

C13H10 C14H10

166 178

1.9 1.15

4.18 4.46

188 199

11.4*7.3*4.2 0.001 11.6*7.9*3.8 0.001

Anthracene (Ant) Pyrene (Pyr)

C14H10 C16H10

178 202

0.0434 0.135

4.45 4.88

170.3 186

-

0.01 0.001

Fluoranthene (Flua)

C16H10

202

0.26

5.16

187.7

-

0.001

Benzao(a)anthracene (B[a]A) Chrysene (Chr)

C18H12

228

0.0094

5.76

-

-

0.1

C18H12

228

0.002

5.81

-

-

0.01

Benzo(k)fluoranthene (B[k]F) Benzo(b)fluoranthene (B[b]F)

C20H12

252

0.0015

5.78

-

-

0.1

C20H12

252

0.0008

6.11

-

-

0.1

Benzo(a)pyrene (B[a]P)

C20H12

252

0.00162

6.13

-

-

1

Dibenzo(a,h)anthracene (D [a,h] A) Indeno(1,2,3-cd)pyrene (I[1,2,3-cd] P) Benzo(g,h,i)perylene (B[g,h,i]P)

C22H14

278

0.00249

6.75

-

-

1

C22H12

276

0.00019

6.7

-

-

0.1

C22H12

276

0.00026

6.63

-

-

0.01

Note: Kow, n-octanol-water partition coefficients

45

11

Table 3. Summary of the results of PAH removal by carbonaceous adsorbents from sediments Adsorbent GAC

SSA, (m2/g) 1050– 1200

Particle size (um) 75-150

Dose (% of DW ) 5%

Contact time 1month 12 months

PAC

1300

<32

5%

1month 12 months

12

PAC

1300

GAC Coke

1050– 1200 3

AC

940

50% particles <15 um 3% particles >150 um 1.7-0.43

2%

6 weeks

2%

6 weeks

63-105 105-250 75-300

3.4%

1 month 6 months 1 month 6 months

3.4%

Initial concentration (mg/kg) 2500 125 2500 125 2500 125 2500 125 5500 38 5500 38 8 8

Reduction in Cw (%)

Mixing

TOC

75% 97% 97% 97% 98% 99% 95% 99% 63% 99% 4% 64% 64% 74% 84%

2 rpm

4%

Water content 95%

2 rpm

4%

95%

2 rpm

4%

95%

2 rpm

4%

95%

End over end End over end 3 rpm

7.2 2.5 7.2 2.5 1.7%

89%-96%

3 rpm

1.7%

91%

-

6.2

99%

1% 90% 3% 99% 50% particles <15 um AC 1300 6% 2 days 1107 >99% 3% particles >150 um 15% >99% 30% >99% Note: Granular activated carbon (GAC), Powdered activated carbon (PAC), Sample dry weight (Dw), Activated carbon (AC);

Reference (Choi et al., 2013)

(Brandli et al., 2008)

89%-96% 91%

(Zimmerm an et al., 2004; Zimmerma n et al., 2005) (Kupryian chyk et al., 2011)

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Figure captions

15

Figure 1. Sources, transport, and sinks of PAHs and their impacts on flora, fauna, and the environment

16

Figure 2. (a) Increasing trend of researches on PAHs remediation applying carbon adsorbents and (b)

17

Discipline distribution of the research

18

Figure 3. Scanning electron microscope image (SEM) analysis of a carbon adsorbent before (a) and after

19

(b) adsorption of PAHs (Kumar et al., 2019)

20

Figure 4. The adsorption mechanisms of PAHs on carbon adsorbents ( The π-π interaction diagram is from

21

Yang et al. (2013))

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Declaration of interests ☒ The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. ☐The authors declare the following financial interests/personal relationships which may be considered as potential competing interests:

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Highlights:  Applications of CAs in immobilization and decontamination of PAHs in sediments are reviewed.  Influences of characteristics of CAs and PAHs, and remediation conditions are discussed.  Physical and chemical mechanisms of CAs for PAH remediation in sediments are summarized.  Negative effects associated with CA remediation of PAH-contaminated sediments are discussed.