Zero-valent iron activated persulfate remediation of polycyclic aromatic hydrocarbon-contaminated soils: An in situ pilot-scale study

Accepted Manuscript Zero-valent iron activated persulfate remediation of polycyclic aromatic hydrocarbon-contaminated soils: An in situ pilot-scale study Yue Song, Guodong Fang, Changyin Zhu, Fengxiao Zhu, Song Wu, Ning Chen, Tongliang Wu, Yujun Wang, Juan Gao, Dongmei Zhou PII: DOI: Reference:

S1385-8947(18)31600-0 https://doi.org/10.1016/j.cej.2018.08.126 CEJ 19740

To appear in:

Chemical Engineering Journal

Received Date: Revised Date: Accepted Date:

8 June 2018 7 August 2018 19 August 2018

Please cite this article as: Y. Song, G. Fang, C. Zhu, F. Zhu, S. Wu, N. Chen, T. Wu, Y. Wang, J. Gao, D. Zhou, Zero-valent iron activated persulfate remediation of polycyclic aromatic hydrocarbon-contaminated soils: An in situ pilot-scale study, Chemical Engineering Journal (2018), doi: https://doi.org/10.1016/j.cej.2018.08.126

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Zero-valent iron activated persulfate remediation of polycyclic aromatic hydrocarbon-contaminated soils: An in situ pilot-scale study

Yue Song1, Guodong Fang1*, Changyin Zhu1,2, Fengxiao Zhu1, Song Wu1,2, Ning Chen1,2, Tongliang Wu1,2, Yujun Wang1, Juan Gao1, Dongmei Zhou1*

1.

Key Laboratory of Soil Environment and Pollution Remediation, Institute of Soil Science,

Chinese Academy of Sciences, Nanjing 210008, China. 2.

University of Chinese Academy of Sciences, Beijing 100049, China.

*Corresponding

author.

Tel:

86-25-86881327,

fax:

86-25-86881000,

E-mail

address:

[email protected] (G. D. Fang); Tel: 86-25-86881180, fax: 86-25-86881000, E-mail address: [email protected] (D. M. Zhou)

1

Highlights: 

ZVI activated PS was in situ pilot-scale tested to remediate PAH in soil



Nano-sized ZVI/PS removed 82% of PAHs



Soil pH decline in varying degrees due to different ZVI with PS



ZVI/PS long-term negative effects on soil microbial diversity



The phylum Firmicutes shows a tolerance to PS stress

Keyword: Persulfate, Zero-valent irons, PAH-contaminated soil, Bacterial communities, Pilot-scale

2

Abstract Zero-valent iron activated persulfate (ZVI/PS) is widely used for the degradation of contaminants. However, the applicability of ZVI/PS has rarely been tested for the in situ pilot-scale remediation of organic-contaminated sites. In this study, different types

of

ZVI,

including

micro/nanostructured

ZVI

(nZVI),

stearic-coated

micro/nanostructured ZVI (C-nZVI) and commercial micron-sized ZVI (mZVI), were used to activate PS to remove polycyclic aromatic hydrocarbons (PAHs) from the contaminated site. Three reaction pits were excavated for testing the activation ability of different types of ZVI, each with an area of ~30 m2 and a depth of 4 m. The chemicals (30 g/kg of PS and 3.5 g/kg of ZVI) were added in a soil slurry by in situ stirring. After treatment for 104 days, the PAHs (~17 mg/kg) removal efficiencies were 82.21%, 62.78%, and 69.14% for PS activated by nZVI, C-nZVI and mZVI, respectively. It was found that the soil pH decreases due to the release of H+ from PS decomposition. The catalase activity was enhanced compared to before the chemical application, while soil bacterial communities, reflected by operational taxonomic unit values, decreased markedly from 250 to ~100. In particular, the bacteria of the phylum Chloroflexi almost disappeared after remediation. However, the bacteria of the phylum Firmicutes still dominated after remediation and exhibited a tolerance to PS. The findings of this study provide a useful implementation case when PS activation is used in the remediation of PAH-contaminated soil and groundwater.

3

1. Introduction The soil and groundwater of industrial lands are frequently polluted by organic contaminants, posing a threat to public health and the ecological environment [1,2]. Polycyclic aromatic hydrocarbons (PAHs) are one of the most concerning organic contaminant groups due to their toxic, carcinogenic and mutagenic effects [3–5]. Although thermal remediation or bioremediation strategies have been used to treat PAH-contaminated sites, these methods generally require high energy consumption or long remediation periods [6–9]. Recently, advanced oxidation processes (AOPs), with the advantages of speed, easy application and aggressiveness, are receiving more and more attention for the remediation of contaminated soil and groundwater [10–12]. AOPs are traditionally based on the generation of reactive oxygen species (ROS), such as hydroxyl radical (•OH) and superoxide radical anion (O2•−), which can nonselectively oxidize or mineralize organic contaminants [13–16]. Recently, AOPs based on sulfate radical (SO4•−) have attracted significant attention and have been widely used for the degradation of a wide range of contaminants such aspolychlorinated biphenyls, perfluorooctanesulfonates, phenols, PAHs, ect.,[17–20]. The SO4•− can be formed by activation of persulfate ions (S2O82−, PS) with different ways, including heat, UV radiation and transition metals (Men+), as shown by the following equations [21–25]:

S2O28‒

Heat/UV

→

4

2SO•– 4

(1)

S2O28‒ + Men + → SO•–4 + Me(n + 1) + + SO24‒

(2)

Among the transition metal ions used for activating PS (e.g., Fe2+, Mn2+, Ag+, Co2+ and V3+), Fe2+ ions have been wildly used to remediate various kinds of organiccontaminated groundwater and soil with the obvious advantages of relatively low cost and environmental friendliness [24,26–29]. In the homogeneous Fe2+/PS reaction system, Fe2+ ions can activate PS to generate SO4•− quickly (Eq.3). However, the quenching reactions between Fe2+ ions and SO4•− would reduce the efficiency of target contaminants degradation (Eq.4)[29]. Hence, how to optimally support Fe2+ ions has become one of the key questions for the success of the Fe2+/PS reaction system.

S2O28‒ + Fe2 + →SO•–4 + Fe3 + + SO24‒

(3)

SO•–4 + Fe2 + →Fe3 + + SO24‒

(4)

In recent years, the development of the zero-valent iron (ZVI) activation of PS (ZVI/PS) has provided a possible method to resolve the problem of the quenching of SO4•− by Fe2+ [30–34]. ZVI can not only continuously support Fe2+ ions through acidification or oxidation (Eqs. 5-7), but also directly reacts with PS to generate SO4•− through the heterogeneous transportation of electrons from ZVI to PS (Eq. 8) [30,34].

Fe0→Fe2 + + 2e ‒

5

(5)

2Fe0 + O2 + 2H2Oâ † ’2Fe2 + + 4OH ‒

(6)

2Fe0 + S2O28‒ →2Fe2 + + 2SO24‒

(7)

Fe0 + 2S2O28‒ →2SO•4‒ + 2Fe2 + + 2SO24‒

(8)

According to their different dimensions, ZVI particles can be classed from nanosized ZVI to micron-sized ZVI [35]. In spite of a higher reactivity with PS due to the greater specific surface area, nano-sized ZVI generally requires a higher synthesis cost than micron-sized ZVI [36]. Recently, Kang et al. [37] reported a facile approach to fabricate a micro/nanostructured ZVI via ball-milling the industrially reduced iron powders. This kind of nano-sized ZVI consists of micron-sized (Φ = 2–5 μm) plates with a nano-sized thickness (δ = 35–55 nm). Moreover, when used in the Fenton reaction, this nano-sized ZVI has demonstrated much high degradation performances to dichlorodiphenyltrichloroethane, and the cost could be reduced to only 8 ~ 13 USD/kg from the commercial nano-sized ZVI price which varies from 150 to 350 USD/kg [36], making it possible for large-scale use. However, similar to common nano-sized ZVI, micro/nanostructured ZVI is also easily oxidized to form agglomerates, resulting in a low mobility and weak reactivity [37]. To increase the dispersion in aqueous media and the mobility in porous media, and slow down the oxidation process, it is necessary to modify the surface properties of nano-sized ZVI, for example, though coating by polymers, anionic surface-active agents or other

6

organic coatings [36,38]. In the last decade, the ZVI/PS technique has been widely tested on the lab-scale to remediate organic-contaminated water, groundwater and soil, particularly for PAHs [33,39,40]. To the best of our knowledge, however, only a few scale-up studies have been conducted using this technique to remediate actual PAH-contaminated soil and groundwater [41–43]. Moreover, to evaluate this emerging soil remediation strategy, it is important and necessary to investigate its impacts on the original ecological system. In spring 2017, therefore, we benefited from an occasion of a soil and groundwater remediation project of a coal gas plant relocation site in Nanjing, China to conduct a pilot-scale study of ZVI/PS in situ remediation of actual PAHcontaminated soil. The two main objectives of this research were: (1) to verify the effectiveness of different types of ZVI [micro/nanostructured ZVI (nZVI), steariccoated micro/nanostructured ZVI (C-nZVI) and commercial micron-sized ZVI (mZVI)] activated PS in a large-scale in situ remediation of PAH-contaminated soil; and (2) to evaluate the environmental impact of this method by examining the change in bacterial communities and soil enzyme activity.

2. Method and materials 2.1. Site characteristics The test site was located at Qixia District, Nanjing City, Jiangsu Province, China. A working coal gas plant was at the site from 1985 to 2000, afterwards this site was

7

used for logistics storage. In 2013, to develop a real estate project, a survey of the environment of the site was completed, revealing that some organic contaminants (mainly PAHs) exceeded the residential land standards of the Chinese environmental quality standards of soil (GB15618-2008), shown in Table 1. Some other soil properties were listed in Table S1 of the supporting information (SI).The total remediation area of this project was ~24,000 m2, in which a test site with a dimension of ~18 m (length)  5 m (width)  4 m (depth) was reserved for our in situ pilot-scale remediation research.

2.2. Dosage chemical characterization Tap water was used to prepare the remediation chemical solutions. Sodium persulfate (99%) was purchased from Suzhou Huahang Chemical Technology Co., LTD (Suzhou, China). 5,5-Dimethyl-1-pyrrolidine N-oxide (DMPO, 97%) was purchased from Sigma-Aldrich, Inc. (Shanghai, China). The commercial micron-sized ZVI (mZVI) was obtained from Tapery instrument equipment Co. Ltd. (Nanjing, China), micro/nanostructured ZVI (nZVI) was purchased from Hangda Technology Co. Ltd. (Shenyang, China) and the stearic-coating of micro/nanostructured (C-nZVI) was processed by the Institute of Solid State Physics, Chinese Academy of Sciences, China.

2.3. Chemical application and sampling operations The chemical application process could be generally described in the following

8

steps. After previous land leveling work, three reaction pits were evacuated, each with dimensions of ~6 m (length)  5 m (width) and 4 m (depth), as shown in Figs. 1(a) and (b). Each pit was separated from each other by 50 cm wide (top width) reserved earth dams. Chemical application and original soil backfilling were operated three times. First, the tap water was filled into the pits with a volume of 1/3 the volume of each pit. Second, PS and different types of ZVI (according to different pits) were added to create a ZVI/PS reaction solution with 30 g/L of PS and 3.5 g/L of ZVI [PS(mol):ZVI(mol) = 2:1]. Third, 1/3 of the contaminated soil was backfilled. Then, the pits were stirred well with an excavator and the aforementioned process was repeated three times until each pit was filled up. Some photographs were taken during the chemical application operation and are shown in Fig. S1 of the SI. In addition, the dosage of PS and ZVI was optimized with experiments in lab-scale studies and referenced to the study of Zhao et al. [20]. After the chemical application, we left the test site for a 104-day remediation, during which four sampling operations were performed on the 0th (original soil samples before remediation), 12th, 33th and 104th days in order to study the short-term, medium-term and long-term remediation effects and the ecological impacts. The weather information including cumulative rainfall and average temperature during the remediation was shown in Fig. S2 of the SI.

9

Fig. 1. Schematic of each reaction pit. The top (a) and cross-sectional (b) views and dimensions. The blue blocks show the positioning of the sampling points. Each sample was collected according to a five-point difference method, as detailed in (c).

10

For each sampling operation, sampling points were located in each reaction pit at two opposite sides, 20 cm from each edge, as shown in Fig. 1(c). We only sampled the soil samples at 5–20 cm depth, since soils deeper than 20 cm were muddy and difficult to sample. At each sampling point, we mixed five samples in 0.36 m2 into one sample. Each sample was immediately loaded in an ice box and transported back to the laboratory and then immediately stored at - 80 °C until further chemical and biological analysis.

2.4. Analysis 2.4.1. ZVI characterization The mineral composition of each ZVI sample was determined using an X-ray diffraction powder diffractometer (D8 advance, Bruker, Bremen, Germany). The specific surface area of ZVI was determined using the Brunauer–Emmett–Teller nitrogen-adsorption method at 77 K using a Quadrachrome Adsorption Instrument (Florida, USA). The surface morphology and structure of the different ZVI samples were investigated using a scanning electron microscope (SEM, ZEISS Merlin, Oberkochen, Germany).

2.4.2. Electron paramagnetic resonance (EPR) studies To identify free radicals in the treatment solution, different ZVI samples and PS solutions were firstly mixed according to the previously mentioned concentration.

11

Then, 0.9 mL of the treatment solution was reacted with 0.1 mL DMPO (0.1 M). Furthermore, the free radicals in the suspension of the contaminated soils were also investigated after adding treatment solution at the water/soil ratio 10 : 1, then 0.9 mL of the suspension was reacted with 0.1 mL DMPO and filtered through a 0.45 μm cellulose filter for the EPR analysis. Our previous research demonstrated the radicals generated for PS remained unchanged during the first 10 min after adding activators [30], so that the analysis is completed in 10 min. The free radicals were examined with an EPR spectrometer (Bruker EMX/plus) with a resonance frequency of 9.77 GHz, a microwave power of 20.02 mW, a modulation frequency of 100 kHz, a modulation amplitude of 1.0 G, a sweep width of 100 G, a time constant of 40.96 ms, a sweep time of 83.89 s and a receiver gain of 1.0  103.

2.4.3. Chemical analysis of soil samples Soil samples of ~100 g were taken from -80 °C storage, freeze-dried and then ground through a 20-mesh sieve for chemical analysis. Soil PAHs were extracted three times by sonication using an acetone-hexane mixture (1:1) and analyzed with gas chromatography coupled to mass spectrometry according to EPA Method 8275A for PAHs [44]. The persulfate concentration was determined using a Shimadzu 2700 UV-VIS spectrophotometer according to the method of Liang et al. [45]. Soil pH (soil: distilled water = 1:2.5) was measured with an Orion Star A211 pH analyzer using a pHS-3B electrode (Shanghai REX Instrument Factory, Shanghai, China). Two replicate samples were used for each chemical analysis to ensure accuracy.

12

2.4.4. Analysis of soil bacterial community and soil catalase activity An E.Z.N.A Soil DNA kit (OMEGA) was used for DNA extraction from soil samples. The V3-V4 variable regions of the 16S rRNA gene were amplified using bacterium-specific primers 343F (5'-TACGGRAGGCAGCAG-3') and 798R (5'AGGGTATCTAATCCT-3') [46]. Sequencing was performed using the Illumina MiSeq platform at OE Biotech Co., Ltd., Shanghai, China. The raw sequences obtained were treated using a series of software (Trimmonatic, FLASH and Userach) to obtain the clean reads [47], which were subjected to primer sequence removal and clustering to generate operational taxonomic units (OTUs) using Vsearch software with a 97% similarity cutoff [48]. Similarities between fingerprints were analyzed statistically (cluster analysis) using the unweighted pair group method with arithmetic averages (UPGMA). The soil enzyme activities including soil catalase, soil uresase and soil phosphatase were quantified using enzyme-linked immunosorbent assay kits (Jinyibai Biological Technology Co. Ltd, Nanjing, China), following the manufacturer’s instructions.

3. Results and discussion 3.1. Activation of PS by ZVI 3.1.1. Characteristics of different ZVI types used in remediation As shown in Fig. 2a (XRD patterns of ZVI types), three Fe0 characteristic peaks

13

could be found in each ZVI XRD patterns at 2θ values of 44.67°, 65.0° and 82.33°, respectively, indicating that the Fe0 was the main structural composition of the ZVI samples used in this study. The specific surface areas for nZVI, C-nZVI and mZVI were 4.04, 6.85 and 0.68 m2/g, respectively (Fig. 2b). According to the SEM micrographs of nZVI and C-nZVI (Figs. 2c and 2d), these two materials present a two-dimensional sheet structure with a nano-sized thickness and an area of micron size in diameter. The C-nZVI possessed a relative lower effect of particle agglomeration than the nZVI due to the stearic-coating. This might be the reason why the C-nZVI showed a higher specific surface area than nZVI. The mZVI, however, were iron particles with the diameter varying from 1 to 100 µm, as the reason of its lowest specific surface area of all ZVI materials.

14

a

b

Specific area (m2/g)

mZVI

Intensity

C-nZVI

nZVI Fe0

15

30

45

2θ

60

75

90

7 6 5 4 3 2 1 0

nZVI

C-nZVI

mZVI

Fig. 2. Characteristics of different zero-valent iron (ZVI) samples: (a) XRD patterns and (b) specific surface areas measured by Brunauer–Emmett–Teller (BET) and scanning electron microscope (SEM) micrographs (c) for nZVI, (d) for C-nZVI and (e) for mZVI.

15

3.1.2. Free radical identification The efficiency of different types of ZVI activated PS was evaluated by the formation of sulfate and hydroxyl radicals identified with EPR spectroscopy coupled with DMPO as spin-trapping agent in both aqueous and soil suspension. As shown in Fig. 3, the DMPO-OH (four lines, 1:2:2:1) and DMPO-SO4 (six lines, 1:1:1:1:1:1) signals were identified according to their hyperfine splitting constants (DMPO−OH: aH = aN = 14.7 G; DMPO−SO4: aN = 13.3 G, aH = 9.5 G, aH = 1.46 G, and aH = 0.77 G) simulated with WinEPR Acquisition software, respectively, as previously reported [30,49]. For PS without adding any activator, the signals of DMPO−OH and DMPO-SO4 were also observed, which was ascribed to the formation of SO4•− and •OH from PS decomposition at ambient temperature [50]. After the addition of nZVI and mZVI to PS solution, the peak intensities of DMPO-SO4 obviously increased compared to the EPR spectra of solely PS with DMPO systems. With a larger reactivity than mZVI, nZVI could efficiently activate PS and presented stronger DMPO-SO4 signals, which indicated that more SO4•− radicals were generated. The peak intensities of the DMPOOH signal in PS after adding nZVI and mZVI did not increase proportionally with the DMPO-SO4 signal. This might be due to the quenching of •OH by a high concentration of PS (30 g/L) in this research. The produced •OH would react with PS with the second-order rate constant of 1.4 × 107 M−1 s−1, while the second-order rate constant of SO4•− with PS was much lower (6.3 × 105 M−1 s−1), as shown in equations (9) and (10) [37]. For a low concentration of PS (2.38 g/L), we reported that both the 16

peak intensities of DMPO-OH and DMPO-SO4 could be increased with the addition of activator in our previous research [30].

SO•4‒ + S2O28‒ → S2O•8‒ + SO42 ‒

k = 6.3 × 105 M - 1 s - 1

HO• + S2O28‒ → S2O•8‒ + OH ‒

(9)

k = 1.4 × 107 M - 1 s - 1 (10)

Conversely, neither the DMPO-SO4 nor DMPO-OH signal was obvious in the CnZVI/PS system. The probably reason is that, in the beginning, the coated hydrophobic stearic layer of C-nZVI not only hindered the contact of PS with ZVI, restricting the activation reaction, but also consumed the SO4•− and •OH radicals at the same time and therefore significantly decreased the intensity of the SO4•− and •OH signal. Predictably, after the stearic layer was oxidized by PS, the surface of C-nZVI would be hydrophilic and could efficiently activate PS. Moreover, the delay time of PS activated by C-nZVI might be convenient for the in situ remediation, as it would take time for the contact of contaminants with the oxidant. Another likely reason was that the DMPO-OH and DMPO-SO4 adducts can be easily sorbed on the hydrophobic stearic layer of C-nZVI, and thus cannot be detected by EPR. Fig. 3b shows that both DMPO-SO4 and DMPO-OH signals were significantly increased in the contaminated soil suspension after adding PS and nZVI compared to in the PS solution. This is mainly due to the co-activation results to the PS of the nZVI and other activator substances in the soil, such as Fe/Mn-oxides (DCB-Fe/Mn)

17

and organic matter, shown in Table S1 [19,24,49]. For the systems PS + mZVI + soil and PS + C-nZVI + soil, DMPO-OH signal could also be identified. However, the DMPO-SO4 signal is not as obvious. That might be because the wide DMPO-OH signal in these systems interfered by other free radical signals. Although the DMPOOH signals of PS + mZVI + soil system and PS + C-nZVI + soil system were slight stronger those of the PS solution, however, which were smaller than those in the PS + nZVI + soil system, indicating that the nZVI was the most effective PS short-term activator of the three tested ZVI both in the aqueous solutions and in the soil suspensions.

18

(b)

(a) 

   

 

mZVI + PS

Intensity

C-nZVI+ PS

  

3440



3460





 

 



3480



 









nZVI + PS



3500

Intensity

 

DMPO-SO4 

DMPO-OH

DMPO-SO4

DMPO-OH















 PS + C-nZVI + Soil



 





 





 

 



PS

3440

3520

3460

3480



PS + mZVI + Soil

PS + nZVI + Soil

3500

PS

3520

Magnetic field (G)

Magnetic field (G)

Fig. 3. EPR spectrum of different ZVI/PS systems: (a) in different treatment solutions and (b) in suspensions contaminated soil after adding treatment oxidant solutions. Reaction conditions: [nZVI]0 = [C-nZVI]0 = [mZVI]0 = 3.5 g/L; [PS]0 = 30 g/L; [DMPO]0 = 0.1 M; T = 25 °C; oxidant solution:contaminated dry soil = 10:1.

19

3.2. PAH removal The concentrations of sixteen PAHs including naphthalene, acenaphthylene, acenaphthene,

fluorene,

benzo(a)anthracene benzo(k)fluoranthene,

phenanthrene,

(BaA),

chrysene

anthracene, (CHR),

benzo(a)pyrene,

fluoranthene,

benzo(b)fluoranthène

dibenz(a,h)anthracene

pyrene, (BbF), (DBA),

benzo(g,h,i)perylene and indeno (1,2,3-cd) pyrene (IND), in the soil before and after remediation are shown in Table 1, The obtained results show that the total

PAH

concentrations in soil decreased rapidly from 16.98 mg/kg to 3.02, 6.32 and 5.24 mg/kg after the remediation with PS activated by nZVI, C-nZVI and mZVI, respectively. The corresponding PAHs removal efficiency was 82.21%, 62.78% and 69.14%, respectively.

20

Table 1. The concentrations of PAHs in soil before and after remediation with different PS activation processes. Soil PAHs concentrations (mg/kg)

Number of

GB 15618—

rings

20081

Original values2

nZVI-103d3

Nap

2

5

0.74

0.00

0.21

0.00

Acenaphthylene

AcPy

3

5

0.47

0.08

0.16

0.12

Acenaphthene

Acp

3

5

0.00

0.00

0.00

0.00

Fluorene

Flu

3

5

0.11

0.00

0.00

0.00

Phenanthrene

PA

3

5

0.98

0.00

0.41

0.21

Anthracene

Ant

3

5

0.00

0.00

0.00

0.00

Fluoranthene

FL

4

5

2.55

0.00

0.63

0.40

Pyrene

Pyr

4

5

0.59

0.00

0.29

0.14

Benzo(a)anthracene

BaA

4

1

1.57

0.29

0.51

0.42

Chrysene

CHR

4

0.5

1.60

0.00

0.00

0.34

Benzo(b)fluoranthene

BbF

5

1

1.28

0.41

1.20

0.69

Benzo(k)fluoranthene

BkF

5

1

0.78

0.00

0.00

0.00

Benzo(a)pyrene

BaP

5

0.5

0.43

0.08

0.47

0.24

Indeno(1,2,3-cd)pyrene

IND

5

0.5

2.04

0.90

0.98

1.15

Dibenzo(a,h)anthracene

DBA

6

0.5

1.08

0.59

0.60

0.70

Benzo(g,hi)perylene

BghiP

6

5

2.76

0.66

0.87

0.83

16.98

3.02

6.32

5.24

Total LPAHs4 removals (%)

96.52

66.08

85.65

Total HPAHs5 removals (%)

80.04

62.19

66.55

Total PAHs removals (%)

82.21

62.78

69.14

PAHs

Abbreviation

Naphthalene

Total

1Highest

C-nZVI-103d3 mZVI-103d3

allowable values for residential land in the Soil Environmental Quality Standards GB

15618—2008 (consultation draft). 2Original

3PAH

PAH concentration values measured before remediation.

concentration residue in the soil after each PS 103 day treatment activated by nZVI, C-nZVI and

mZVI, respectively. 4LPAHs:

low molecular weight PAHs (2 ~ 3 rings)

5HPAHs:

High molecular weight PAHs (4 ~ 6 rings)

21

Generally, 16 PAHs could be classified into low molecular weight PAHs (LPAHs) with 2 to 3 rings and high molecular weight PAHs (HPAHs) with 4 to 6 rings [51]. Table 1 shows that the total LPAHs removals efficiencies for three different treatment were 95.74%, 73.93% and 88.26%, respectively, which were higher than the total HPAHs removals efficiencies (80.82%, 55.55% and 61.14% respectively). The obtained results could support the conclusion of Chen et al. [52] previous lab-scale study, which is that nZVI/PS could provide a higher removal efficiencies to LPAHs than to HPAHs. Moreover, degradation of LPAHs easier than HPAHs have been confirmed by the other advanced oxidation treatment such as Fenton’s reagent and ozonation [39,53,54]. Among these PAHs, the concentrations of five HPAHs (BaA, CHR, BbF, IND and DBA) in the original soil were higher than the highest allowable values for residential land (degree 2) of the Soil Environmental Quality Standards of China (GB 15618—2008,consultation draft). Therefore, these 5 PAHs were the contaminants of primary concern in this research and their concentration variations are shown in Fig. 4. For BaA, CHR and IND, no matter what kind of ZVI was used to activate PS, the concentrations of these PAHs were obviously decreased, especially during the shortterm (12 days) remediation period (Figs. 4a, b and c). Although the concentration of these two PAHs showed a decreasing trend, there were some anomalous points even higher than the original value during the short- and medium-term remediation periods, such as the BbF concentration at 33 days in the experiment using mZVI and the DBA concentration at 12 days in the experiment activated by nZVI (Figs. 4d and e). 22

Conversely, in most sampling points of the reported lab-scale ZVI/PS remediation of spiked PAH-contaminated soil, all the 16 PAHs continuously decreased without any PAHs increasing in concentration [33]. Soil organic matter (SOM) such as humin, humin acid (HA) and fulvic acid (FA) are abundant in the soil. PAHs in soil could easily across the HA/FA boundary layer and into the humin core due to their hydrophobicity and the affinity of PAHs for HA increased with the molecular weight of the PAHs [55]. Bogan and Trbovic [54] reported that, comparing to LPAHs, the HPAHs is more difficult to be released from humin core to HA/FA domain and aqueous phase and to contact with oxidant, thus delaying the oxidation process. The delayed release of HPAHs from SOM after HA/FA was oxidized by persulfate was probably reason for the increase of DBA (6 rings) and BbF (5 rings) at the beginning of the treatment and the relatively lower final removal efficiencies of these two PAHs.

23

(a)

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Fig. 4. Changes of five PAH concentrations in soil during the remediation: (a) Benzo(a)anthracene (BaA); (b) Chrysene (CHR); (c) Indeno(1,2,3-cd)pyrene (IND); (d) Benzo(b)fluoranthene (BbF); (e) Dibenzo(a,h)anthracene (DBA).

24

3.3. Changes in soil pH, bacterial diversity and enzyme activities During the remediation process, soil characteristic changes within each treatment were monitored. Fig.5a shows that the PS concentration continuously decreased during the treatment and almost no PS was detected in the top layer of the soil after the treatment of 104 days, due to the activation through ZVI.

The decomposition of

PS usually generates H+ and leads to the decrease of soil pH. In addition, PS is an acidic solution (pH = 2.3 for 30 g/L PS).Thus, the changes in soil pH was examined during the remediation. Fig. 5b shows that the soil pH decreased markedly from 9.0 to 7.6, 6.0 and 2.8 at 12 days for nZVI, C-nZVI and mZVI, respectively. After that, the soil pH changes slightly during the whole remediation period. This is due to the significant accumulation of H+ by PS decomposition when ZVI contact with PS at the beginning of the reaction. For nano-sized ZVI treatment (nZVI and C-nZVI), the degree of pH reduced was about 3.5-5 units lower than that of mZVI treatments. This may be due to the hydrogen evolution reaction of nZVI and C-nZVI in water (Eq. 6), which could produce a certain amount of OH− and neutralized the H+ ions from the PS solution. Conversely, due to the smaller specific area of mZVI, the hydrogen evolution from the direct reaction with mZVI and water was not obvious, which led to a dramatic pH decrease in the soil by PS addition. These combined results indicated that nano-sized ZVI has limited effects on soil pH, and was more feasible than commercial ZVI power for PS activation and soil remediation. Soil bacterial diversity as an important indicator for soil ecological function was analyzed before and after remediation. Fig.5c illustrates the soil bacterial diversity

25

based on the number of OTUs (at a 97% cutoff) recovered from these samples. The initial soil contained ~250 OTUs, which was slightly higher than the number of OTUs recovered from other PAH-contaminated soil or sediment (50–150 OTUs) [56–58], but lower from common agricultural soils (1500–1900 OTUs) [59]. The results were probably caused by the long-term PAH contamination and/or the low nutrition status of the tested soil. After remediation at 10 day, the OTUs decreased to below 150 and remained at this level until the end of the 104-day remediation. These results indicate that the method of PS activated by ZVI has a long-term and negative impact on the original microbial diversity in the PAH-contaminated soil. In addition, more results regarding the species diversity calculated based on OTUs are given in Table S2 of the SI. Soil enzyme activities are common indicators for monitoring various impacts on soils, as they are very sensitive to both natural and anthropogenic disturbances and show a quick response to the induced change [60,61]. Among these enzymes, catalase is often used as an indicator of oxidative stresses that can split hydrogen peroxide into molecular oxygen and water and thus prevent cells from damage by reactive oxygen species, such as hydrogen peroxide, superoxide radicals and hydroxyl radicals [62,63]. Fig. 5d shows that the soil catalase activity slightly increased after the oxidant application as a consequence of an oxidative stress response of soil microorganisms [64]. However, it did not exhibit a significant difference among different sampling times (p-value = 0.333) nor among different ZVI activations (pvalue = 0.523). In addition, the variation of soil urease activity and soil phosphatase

26

activity are shown in Fig. S3.

27

80

9 8 7

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25 20 15 10 5 0

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0 day

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12 days

33 days

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Time

Fig. 5. Soil characteristic changes: a) Residual PS concentration; b) Soil pH; c) OTU counts; d) Soil catalase activity.

28

3.4. Bacterial community changes Soil bacterial communities across all samples were plotted using weighted Fast UniFrac principal coordinates analysis (Fig. 6a). It should be mentioned that the communities of samples m.12b were not shown, because they were dramatically different from the other samples for unknown reasons. The bacterial communities obtained in this study clustered into four different groups: group A (initial samples I.a and I.b), group B (the sample n.12b), group C (samples n.12a and m.12a) and group D (rest of the samples). Among these groups, groups B, C and D (sample after oxidant application) were distinct from group A (the original bacterial community structure). In general, bacterial communities can be predicted based on time, with a trend shifting from A to D. The shift of communities from group B and C to D was possibly due to the unstable change during the shortterm stage of the treatment. The clustering of samples under different treatment at days 33 and 104 (group D) indicated that bacterial communities might have reached a stable stage after a 1-month treatment and PS activation by different iron forms had a similar effect on the bacterial communities.

29

Fig. 6. Bacterial community analyses. a) Weighted Fast UniFrac principal coordinates analysis (PCoA). b) UPGMA clustering tree with UniFrac Jackknife Environmental Cluster analysis. The characters of “I”, “n”, “m” and “c” indicate the initial samples, nZVI, mZVI and C-nZVI activation samples, respectively; the characters of “a” and “b” indicate duplicate samples; the numbers of “12”, “33” and “104” indicate the sampling time (days). Red indicates a reliability of 75%–100%; yellow, 50%–75%; green, 25%–50%; blue, <25%.

30

Fig.7a shows that Firmicutes (85%–87%) was the dominant phylum in the original soil, followed by Proteobacteria (6.3%–6.8%) and Chloroflexi (2.1%–3.7%). The observed dominance of Firmicutes and Proteobacteria in PAH-contaminated soil was in accordance with a previous study by Lee et al. [58], where they reported that these two phyla could reach up to 91.15% in PAH-contaminated soils. Moreover, studies have shown the importance of microorganisms closely affiliated with members of the phylum Firmicutes and class Bacilli in the biodegradation of PAHs [4,65]. In the soil examined in the current study, Bacillus and Lactococcus were the two main orders of the class Bacilli [shown in Fig. 7(b)]. The capability of Bacillus to degrade PAHs has been widely reported [4,65–67], while the PAH degradation capacity of Lactococcus is rarely reported and remains to be verified. After the chemical application, the bacteria of the phylum Chloroflexi could no longer be detected in the soil samples, indicating that this phylum was susceptible to the ZVI/PS treatment. The relative abundances of Proteobacteria also decreased but to a lesser extent, suggesting that a portion of bacteria from this phylum might be sensitive to ZVI/PS. Conversely, after adding chemicals, no decrease in the relative abundances of Firmicutes was observed (Fig. 7a). Indeed, it showed an increase to >90%, which was very likely due to the decrease of other bacterial members, as represented by a sharp decline in the number of OTUs obtained. Fig. 7b illustrates that the two dominant genera (Bacillus and Lactococcus) of the phylum Firmicutes remained stable during the remediation process. 31

The genus Bacillus is well known for its high tolerance to some extreme environmental conditions. High temperature, high oxidant concentration and high pressure are usually required to sterilize this kind of bacteria [68,69]. Under stressful environmental conditions, the bacteria can produce oval endospores and remain in a dormant state for very long periods. When the environmental conditions ameliorated, the bacteria can reduce themselves from this endospores state [70]. Bacillus’ endospores in the dormant state show very low basic biological activities and PAH metabolic activity, whereas Bacillus in normal status may have be potent PAH degraders [58]. Therefore, the observation in this research raises an interest in elucidating whether there exists a coupling mechanism between Bacillus biodegradation and AOPs.

32

Fig. 7. Soil bacterial community profiles (Top 30) at the phylum (a) and genus levels (b).

33

4. Conclusions An in situ pilot-scale study of persulfate activated by different types of ZVI demonstrated that this method could effectively degraded PAHs in contaminated soil. The results showed that the PAH removals were 82.21%, 62.78%, and 69.14% according to nZVI-, C-nZVI- and mZVI-activation, respectively. Moreover, the impacts on the soil biochemical indicators caused by this strategy should be noted: the PS solution present an acidification effect to the soil; remediation chemical showed an oxidative stress on the soil and led the soil catalase activity to be slightly increased; oxidizing agents had a long-term inhibitory effect on the abundance of soil microbial community. The bacteria in the original soil of the phylum Firmicutes, genus Bacillus present a tolerance of environmental conditions changes caused by the remediation process. Taking into account the PAH-biodegradation effect of this genus, we believe that it is necessary to further study the coupling mechanisms of Bacillus and chemical oxidation remediation. While benefiting from the high organic-contaminant removal effectiveness of the technique of persulfate in situ chemical oxidation (PS-ISCO), we have to face the problem of salinization due to residual sulfate in the soil after remediation, which is one of the major drawbacks of PS-ISCO. Unfortunately, to the best of our knowledge, there is a lack of rapid and effective methods of recovering sulfate from the soil. For reducing the secondary environmental risks of PS-ISCO, it is of great significance to develop effective methods (soil flushing, bio-remediation etc.) for resolving the 34

sulfate salinization problem.

Acknowledgements:

We would like to express our deepest gratitude to Beijing Construction Engineering Group (Beijing, China) for providing us with the test site and necessary managements during our research project. This work was supported by grants from the National Key Basic Research Program of China (2017YFA0207001, 2016YFD0800204), the National Natural Science Foundation of China [41671478], the Natural Science Foundation of Jiangsu Province of China (BK20170050), and the Youth Innovation Promotion Association of CAS (2014270).

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Highlights: 

ZVI activated PS was in situ pilot-scale tested to remediate PAH in soil



Nano-sized ZVI/PS removed 82% of PAHs



Soil pH decline in varying degrees due to different ZVI with PS



ZVI/PS long-term negative effects on soil microbial diversity



The phylum Firmicutes shows a tolerance to PS stress

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