Chemosphere 91 (2013) 1273–1280

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Reductive degradation of oxygenated polycyclic aromatic hydrocarbons using an activated magnesium/co-solvent system

0045-6535/$ - see front matter � 2013 Elsevier Ltd. All rights reserved.http://dx.doi.org/10.1016/j.chemosphere.2013.02.031

⇑ Corresponding author. Tel.: +1 407 823 2135; fax: +1 407 823 2252.E-mail address: cherie.yestrebsky@ucf.edu (C.L. Yestrebsky).

Marc R. Elie, Christian A. Clausen, Cherie L. Yestrebsky ⇑Environmental Chemistry Laboratory, University of Central Florida, 4000 Central Florida Blvd., Orlando, FL 32816, United States

h i g h l i g h t s

" The use of Mg–cosolvent system for the simultaneou s degradation of OPAHs at roo m temperature." Reductive degradation of the selected OPAHs occur red, after the activation of Mg with acetic acid." This system degraded the selected OPAHs (>86%) and produced seve ral deriva tives within 24 h." Degradation products may be less toxic than OPAHs." This is the first report on OPAH remediation and may lead to the development of an actual field app lication.

a r t i c l e i n f o

Article history:Received 11 September 2012 Received in revised form 21 February 2013 Accepted 22 February 2013 Available online 26 March 2013

Keywords:MagnesiumDegradationOxygenated polycyclic aromatic hydrocarbonsChemical reduction

a b s t r a c t

This study evaluates the capability of zero-valent magnesium and a protic co-solvent to promote the deg- radation of oxygenated polycyclic aromatic hydrocarb ons compounds, specifically 9-fluorenone, 9,10- anthraquinone, 7,12-benz(a)anthraquionone, and 7H-benz(de)anthracene-7-one. At room temperature conditions, greater than 86% degradation efficiency is observed after 24 h of reaction time for a mixture containing 0.05 g of magnesium and four selected oxygenated aromatic hydrocarbons with 250 mg L�1

concentrations. It is noted that glacial acetic acid is needed as an activator for the degradation reaction to proceed. It is also presumed that the acid removes oxide and hydroxide species from the magnesium surface. With the GC–MS analysis of the reaction products, possible reductive pathways are suggested.Furthermore, this study is the first report on the degradation of these emerging contaminants and it isproposed that the magnesium-powde r/protic-solvent system is a promising low-cost reagent and may allow for the future development of an economic and environmentally-friend ly remediation application.

� 2013 Elsevier Ltd. All rights reserved.

1. Introduction

Polycyclic aromatic hydrocarbo ns (PAHs) are widely occurring in natural media such as soil, sediment, water, air and plants as aresult of both natural and anthropoge nic processes. Volcano erup- tions, forest fires, car exhaust and combustion of fossil fuels create the majority of PAHs in the environment (Gan et al., 2009 ). These widespread organic compounds are considered as priority pollu- tants due to the mutagenic and carcinogenic properties of several PAHs and their metabolites . As a result of their wide occurrence and toxic properties, effective techniqu es are required for their degradation and detoxification.

PAHs have been extensively investigated throughout the litera- ture, but their oxygenated derivatives may have increased human and environmental risks. Reports on the toxicological importance of oxygenated polycyclic aromatic hydrocarbons (OPAHs) have re-

sulted in a growing interest in the environm ental occurrence and fate of these contaminan ts. Moreover, recent studies have shown that the concentrations of OPAHs are significant and even higher than those of the parent PAH concentr ations (Layshock et al.,2010; Musa Bandowe et al., 2010 ). OPAHs are transformation prod- ucts of PAHs containing one or more carbonylic oxygen(s) attached to the aromatic ring structure and consist of ketones and quinones (Lundsted t et al., 2007 ). These oxidized derivatives can be pro- duced during incomplete combustion or pyrolysis of organic mate- rial and emitted alongside PAHs. Subsequent research has shown that OPAHs result from the photochemical oxidation of PAHs inthe atmosphere (Vione et al., 2006 ) and the microbial degradation of PAH-contam inated soils (Andersson et al., 2003 ). These polar compounds are ubiquitous , potentially more soluble than parent PAH compounds and consequentl y may have a higher bioavailabil- ity (Lundstedt et al., 2007 ). This is a cause of concern, due to the fact that there is growing evidence that OPAHs are direct-acting mutagen s and carcinogens (Durant et al., 1996 ). OPAH quinones can act as generators of reactive oxygen species (ROS) in biological

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systems leading to gene mutations. In addition, OPAHs can under- go enzymatic and non-enzym atic redox cycling; forming DNA ad- ducts which can ultimately induce carcinogenesis (Bolton et al.,2000). Bioassay of solid environmental samples, containing OPAHs,have demonstrat ed to be more toxic to human and bacteria cells,compared to PAH-containing fractions (Xia et al., 2004; Lemieux et al., 2008 ). Thus contaminat ed sites with OPAHs require immedi- ate remedial action to protect human and ecosystem health. To the best of our knowledge, no investigations on OPAH-remedia tion methods have been reported in the literature.

In our previous work (Elie et al., 2012 ), it was demonstrated that the developed reducing-syste m was able to degrade and con- vert 94% of toxic benzo[a]pyren e (initial concentratio n of250 mg L�1) into a mixture of less toxic and partially-hy drogenated derivatives. The objective of this study is to investigate the poten- tial of this system for the reduction of selected OPAH compounds under ambient conditions. The experimental results may provide a better understand ing of the efficiency and necessar y optimiza- tion of this system for future research on the simultaneous degra- dation of PAHs and OPAHs in solid environmental matrices.

2. Materials and methods

2.1. Chemicals

9-Fluorenon e (9-FLUO), 9,10-anthraqu inone (9,10-ANTQ), 7H- benz(de)anthracene-7-one (BEZO), 7,12-Benz[a]an thracenequi-

Fig. 1. Chemical structures

none (7,12-BaAQ) and 9-fluorenol (9-FLUOL) were purchased from TCI America (Portland, OR). Anthracene (ANT), fluorene (FLU),Benz[a]anth racene (BaA) were purchased from Accustandar d Co.(New Haven, CT). Nitrobenzene (internal standard), ethyl lactate,toluene and absolute ethanol solvents were obtained from Fish- er-Scient ific (Ottawa, ON.). Spherica l magnesiu m (Mg) powder (with a particle diameter distribut ion of 20–100 lm) was obtained from Hart Metals, Inc. (Tamaqua, PA). Helium gas, for GC/MS anal- ysis, was purchase d from Air Gas (Atlanta, GA). All chemicals were received in high purity (P98%) and ACS reagent, analytical grade.

2.2. Ball milling procedure and Mg powder characterizati on

Red Devil 5400 series paint shaker, fitted with custom plates tohold milling canisters, provided vibratory energy (670 rpm) for ball milling of the metal. The canister and balls are made of stainless steel. The canister has an internal diameter of 5.5 cm and a length of 17 cm correspondi ng to a capacity of about 250 mL. 76 g of Mgpowder and 9 g of graphite (C) were introduced into the canister with 16 steel balls (1.5 cm diameter), correspondi ng to a ball-to- powder mass ratio of 3:1. The canister was sealed under nitrogen atmosph ere. The milling duration was varied from 30 min to 1 h.

Morphol ogy and elementary composition of the ball-mill edpowder was examine d by a scanning electron microscope (SEM)LEO 1455VP (20 kV) equipped with an Oxford Inca energy-disper -sive X-ray spectromete r (EDS). The analysis of the sample was facilitate d by dispersing the powder onto a conductive carbon

of the selected OPAHs.

Fig. 2. Respective degradation of: (a) 9-fluorenone (9-FLUO), (b) 9,10-anthraqui- none (9,10-ANTQ), (c) 7H-benz[de]anthracene-7-one (BEZO) and (d) 7,12- benz[a]anthraquinone (7,12-BaAQ) with the activated Mg/co-solvent system. Each data point represents the mean ± standard deviation calculated from triplicate samples.

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adhesive, attached to the sample stub, prior to insertion into the microscope. A typical SEM image is shown in Fig. S1 (Supplemen-tary material)). From the micrograph, the morphology is nearly spherical and the particle size range is 4–30 lm, with an average particle diameter of 10 lm. EDS analysis indicated the presence of oxygen on the Mg surface, which could be a combinati on ofmagnesium oxide (MgO) and magnesium hydroxide (Mg(OH)2)-formed during the exposure of the powder to the humid atmo- sphere. This is in accordance with the findings of Nordlien et al.(1997), who have demonstrat ed the presence of both MgO and Mg(OH)2 on the surface of Mg powder exposed to humid air. No at- tempt was made to control humidity, as the goal is to investigate the workings of the reagent under ambient condition s.

2.3. Experimenta l procedure

In our previous work (Elie et al., 2012 ), we determined that the use of glacial acetic acid and ball-milled magnesium /graphite (Mg/C) provided the optimum reducing conditions. The same condi- tions are used here. We chose to work with 9-FLUO, 9,10-ANTQ,7,12BaAQ and BEZO (Fig. 1), which have shown to be prevalent in various solid matrices, such as dust particulates, soils and sedi- ments. Solutions of each of the compounds of 250 mg L�1 in a prot- ic solvent (1:1 ethanol/ethyl acetate) were prepared. The reaction vessels were 20-mL vials with PTFE-lined caps. A 2-mL aliquot ofa solution was placed into a vial along with 0.05 g of Mg/C powder.Then 60 lL of glacial acetic acid (3 v/v%) were added before the vial was capped. The vial was kept on a lab bench at room temperature (�27 �C). After a selected time interval, the reaction was stopped and the reaction mixture was extracted by adding 2 mL of toluene and sonicating for 15 min.

This 4-mL miscible solution was drawn into a filtered syringe (nylon filter/0.45 lm pore size) and then transferred into a 20- mL centrifuge tube. Next, 4 mL of deionized water was added tofacilitate the separation of ethanol/toluene mixture and the parti- tion of OPAHs in the toluene (hydrophobic) phase. The sample was then centrifuged for 20 min to allow the complete separation and partition of the analytes in the toluene layer, and removal ofany residual Mg particles. The toluene layer was then removed for analysis. All of the experime nts were conducted in triplicates.The toluene extracts were analyzed , by gas chromatograp hy–massspectrometry (GC–MS), in triplicates for the residual concentra- tions of parent OPAHs and concentrations of degradat ion products.By using the OPAH standard solution as the reference control and comparing its response factor to that of the toluene extract, inthe absence of acid, it was determined that the extraction effi-ciency of the experime nts was above 88% (see Eq. (S1) in Supple- mentary material). Fig. S2 (Supplementary material) depicts the steps followed during the experime ntal procedure.

2.4. Sample characterizati on method

GC–MS analyses were performed on an Agilent 6850 series IIgas chromatograph fitted with an Agilent 5975 MS detector. ADB-5 capillary column (DB-5MS 30 m � 0.25 mm i.d.; 0.25 lm filmthickness) was used with flow rate (helium) of 1.0 mL min �1. The instrument paramete rs were as follows: initial oven temperature at 45 �C, 1 min hold, ramp to 180 �C at 25 �C min �1, 2 min hold,ramp to 270 �C at 3.5 �C min �1, 1 min hold, final ramp to 310 �Cat 25 �C min �1, 1 min hold. The temperat ure of injector and detec- tor were maintained at 280 and 250 �C respectively. Injection vol- umes were 1 lL and were performed in splitless mode using helium as carrier gas (gas velocity 38 cm s�1). Purge gas time and flow rate were set at 0.5 min and 100 mL min �1, respectively .

Calibration plots for selected OPAH were prepared in the con- centration range of interest and were found to be linear with R2

values >0.98. Eluted compounds were identified by comparing sample mass spectra to the reference spectra database cataloged by the National Institute of Standard s and Technology (NIST).

Fig. 3. Chemical structures of products resulting from chemical reduction of: (a) 9-fluorenone (9-FLUO) and (b) 7,12-benz[a]anthracenequinone (7,12-BaAQ) with the activated Mg/co-solvent system.

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Nitrobenzene was used as an internal standard for the analyses.The quantitation of the recovery products was calculated from the GC/MS response to the parent compound, due to the fact that standards of the hydrogenated products are not commerciall yavailable. Selected ion monitoring (SIM) paramete rs can be found in the Supplementary data (Table S5).

3. Results and discussion

3.1. Acetic acid activation

It was determined that no reaction occurred without the addi- tion of acid. This was expected , as other studies have shown that for the initiation and even enhancement of a magnesiu m-induced reaction, an acid should be added to the mixture (Tilstam and Weinmann, 2002; Patel and Suresh, 2006 ). It has been shown that the reduction of organic compounds with zero-vale nt metals (ZVMs) is surface-control led, and involves adsorption of the organ- ic species onto the ZVM surface, followed by direct electron trans- fer on the metal surface (Matheson and Tranyek, 1994; Burris et al.,

1995). Thus we infer that the lack of reaction, in the absence ofacid, is due to the presence of passivating oxide and hydroxide (MgO/Mg(OH)2) that hinders the contact between active Mg sites and the PAH. The addition of an acid removes the superficialoxide/hyd roxide. In our experime nts, glacial acetic acid was used because it is a weak acid and able to remove the oxide/hydro xide layer without corroding the base metal (Chavez and Hess, 2001 ),and the absence of water avoids building up of passivating Mg(OH)2 (Taub et al., 2002; Grosjean et al., 2006 ).

3.2. Degradat ion of selected OPAHs

A kinetic study of this treatment system was carried out in the presence of acetic acid, in a closed vial, at room temperat ure. The time-dep endent concentratio n profiles, of each OPAH reactant,are illustrated in Fig. 2. The percent degradat ion for 9-FLUO (Fig. 2A), 9,10-ANTQ (Fig. 2B), BEZO (Fig. 2C), and 7,12-BaAQ (Fig. 2D), was 97%, 95%, 87% and 89%, respectivel y during a reaction time of 24 h. As seen from Fig. 2A, the concentratio n of 9-FLUO de- clined rapidly and nearly 98% of it (initial concentration of

Table 1Fitted pseudo-first-order rate constants and half-lives of the four OPAHs degraded by Mg/C.

Reactants Initial concentration (mg L�1) Rate constant a k (h�1) Half-life a t1/2 (h) Coefficient of determination R2 Reaction time (h)

9-FLUO 250 0.2568 ± 0.006 2.7 ± 0.06 0.9729 129,10-ANTQ 250 0.0963 ± 0.014 7.2 ± 1.1 0.9829 24BEZO 250 0.0733 ± 0.011 9.4 ± 1.4 0.9899 247,12-BaANTQ 250 0.0722 ± 0.012 9.6 ± 1.7 0.9863 24

R2 is the statistical measure of the linearity of the curve fit.Reaction conditions: at room temperature selected OPAH in 2 mL of co-solvent solution mixed with 0.05 g of ball-milled magnesium/graphite (9:1 Mg/C) powder and acetic acid added (3 vol.%).

a Mean ± standard deviation calculated from triplicate samples.

Fig. 4. Proposed reaction mechanism for the chemical reduction of a ketone-substituted OPAH into a hydroxylated aromatic compound (OH-PAH) (pathway A).

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250 mg L�1) was reduced within the first 12 h, presumably through direct electron transfer resulting from a newly activated Mg sur- face. Mass spectrometry (GC–MS) analysis revealed that the only two products, formed from this reaction, are 9-fluorenol (9-FLUOL)and fluorene (FLU) (Fig. 3A). Similar trends were observed for 9,10- ANTQ, BEZO and 7,12-BaAQ, where reductive deoxygenation of the ketone functional group(s) occurred and lead to the formation ofthe correspond ing aromatic compounds , hydroxylate d intermedi- ates and hydrogenated derivatives. As an example, we show inFig. 3B the products of degradat ion of 7,12-BaAQ, which are benz[a]anth racene, benz[a]anth racen-12(7H)one and 7,12-dihy- drobenz[a]a nthracene. Figs. S4 and S6 (Supplementary material)show the respective chemical structure s of the products, resulting from the chemical reaction of 9,10-ANTQ and BEZO with the acti- vated Mg/C reducing system. Figs. S3, S5, S7, S8 (Supplementary material) summari ze the product distribut ion analysis from the

chemical reductions of 9-FLUO, 9,10-ANTQ, BEZO and 7,12-BaAQ with Mg/C. Selected ion monitoring (SIM) parameters and example chromatogram s, of the parent compounds and by-products, are presente d in the Supplement ary material section (Table S1 and Fig. S10).

3.3. Degradat ion kinetics

Past studies have shown that rates of degradation of target compounds by ZVM follow pseudo-first-order kinetics with respect to the concentration of the target compound s (Johnson et al., 1996;Wang et al., 2008 ). Our results are similar. A pseudo-first-orderrate constant can be obtained by maintaining the concentration of the ZVM well in excess, and such that any change in the concen- tration of the metal is insignificant compared to the change in con- centration of the target compounds (Wanarat na et al., 2006 ). For

Fig. 5. Proposed reaction mechanism for the deoxygenation of an OPAH and formation of a partially hydrogenated counterpart (pathway B).

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our experimental results, 2 mmol of Mg is needed to convert 0.002 mmol of PAH, which corresponds to a molar ratio (Mg:PAH)of 1000. A pseudo-first-order kinetic model was used to describe the reaction rates observed for this particular study.

Table 1 shows the resulting rate constants obtained by fitting aline to a plot of –lnCt/C0 versus time; where C0 is the initial OPAH concentratio n (250 mg L�1) and Ct is the OPAH concentration attime t. A measure of the quality of the fit (the linearity) is given by the coefficient of determination (R2), in all cases quite close tounity (R2 > 0.97); indicating that the pseudo-first-order kinetic model is in good agreement with the experimental data (seeFig. S9 in the Supplementary data). However, there is a noticeable reduction in the rate of degradation past the 1-h mark. Because activated Mg reacts with ethanol to form magnesium ethoxide,Mg(OCH2CH3)2 (Joseph et al., 2009 ), we presume that the same isoccurring in our reaction system. The precipitatio n and buildup of white solid ethoxide species impede the adsorption of the reac- tant OPAH and ethanol on active Mg sites.

3.4. Possible reaction pathways and advantages of this treatment system

The chemical nature of the identified products, in Fig. 3, suggest that the mechanism of the degradation process is mainly reductive ,converting the carbonyl substitue nt(s) into a hydroxyl group(s)which is in turn deoxygenated to form either the aromatic hydro- carbon or the hydrogenated derivative(s). Proposed reaction path- ways are illustrated in Figs. 4–6.

Kim et al. (2006) have demonst rated that with the use of protic solvents such as ethanol, Mg metal is able to reduce, by single elec- tron transfers, ketones into correspondi ng alcohols. Correlating our results with the aforementione d work, it is proposed that a possi- ble pathway for the reduction of ketone or quinone-substi tuted PAHs proceeds by a single electron transfer (SET) from Mg to the vacant molecula r orbital of the substrate to give a radical anion.This radical anion is subsequently protonated by a nearby ethanol molecule to give a carbon-cent ered radical (Fig. 4). Then a second SET is involved to form a carbanion, which will ensuingly abstract a proton from ethanol and lead to the formatio n of a hydroxyl PAH compound (OH-PAH).

From the product distribution stemming from the reactions of9-FLUO or BEZO, it can be deduced that deoxygenation (loss of car- bonylic oxygen) occurs. Mg has also been reported to deoxygenate certain organic compounds , in the presence of a proton donor, atroom temperature (Khurana et al., 2007 ). Similar to the mechanis mproposed by Hall et al. (1971), in Fig. 5 we show how the OH-PAH reacts further by SET from the Mg giving rise to a radical anion,which subsequent ly results in the loss of the hydroxyl group. The resulting radical undergoes a second SET to make a carbanion and is protonated by ethanol to produce the hydroaromati c end product.

The product distribution, of 9,10-ANTQ or 7,12-BaAQ, reveals that deoxygen ation occurs for quinone-substi tuted compound sand leads to the formation of aromatic compounds . A third reaction mechanis m (pathway C) is proposed in Fig. 6, where the reduction of quinone into diol occurs, followed by deoxygenation and pro- duction of an aromatic parent compound.

Fig. 6. Proposed reaction mechanism for the conversion of a quinone-substituted OPAH into a fully aromatic compound (pathway C).

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The products observed, during the degradation process, are less harmful than the parent compounds . OH-PAHs may not pose a significant environmental risk because of their strong sorption on solid matrices, decreasing their bioavailabili ty and thus their toxicity (Fallahtaf ti et al., 2012; Xiao et al., 2012 ). In addition,they can be easily photodegraded , explainin g their lower abun- dance in the atmosphere (Wang et al., 2007 ). The products ofthe mechanis ms in Figs. 5 and 6 are less toxic and less mutagenic than the parent OPAHs (Durant et al., 1996; Lundstedt et al.,2007).

The system, presented in this work, has the benefit in that itdoes not need gaseous or potentially harmful reagents nor extreme temperature s or pressures. Thus, it might be possible to conceive arelatively simple and low-cost solid matrix remediation process inwhich PAH/OPAH contaminat ed material is washed with an etha- nol–ethyl acetate co-solvent and then reduced with magnesium powder.

4. Conclusions and future work

As a result of this work, a method has been demonstrated wherein oxygenated polycyclic aromatic hydrocarbo ns can be re- duced to less-toxic compounds . Conversion rates of at least 86%are obtained after 24 h, using magnesium powder and a protic co-solvent at room temperat ure and pressure. Addition al study isneeded to see how the Mg/ethanol–ethyl lactate system can beused in actual soil remediation projects for the simultaneou s deg- radation of OPAHs and PAHs. Future efforts will be aimed at opti- mizing the ratio of magnesium to OPAH for increased degradation , and investigating the potential effect of soil organic matter on the efficiency of this system.

Acknowled gements

We would like to thank the McKnight Doctoral Fellowship and all colleagues, past and present, in the Environmental Chemistry Laborato ry at the University of Central Florida for their help and support during this research.

Appendi x A. Supplementar y material

Supplement ary data associate d with this article can be found, inthe online version, at http://dx.doi .org/10.1016/j.chemospher e.2013.02.0 31.

References

Andersson, B.E., Lundstedt, S., Tornberg, K., Schnurer, Y., Oberg, L.G., Mattiasson, B.,2003. Incomplete degradation of polycyclic aromatic hydrocarbons in soil inoculated with wood-rotting fungi, and their effect on the indigenous soil bacteria. Environ. Toxicol. Chem. 22, 1238–1243.

Bolton, J.L., Trush, M.A., Penning, T.M., Dryhurst, G., Monks, T.J., 2000. Role ofquinones in toxicology. Chem. Res. Toxicol. 13, 135–160.

Burris, D.R., Campbell, T.J., Manoranjan, V.S., 1995. Sorption of trichloroethylene and tetrachloroethylene in a batch reactive iron–water system. Environ. Sci.Technol. 29, 2850–2855.

Chavez, K.L., Hess, D.W., 2001. A novel method of etching copper oxide using acetic acid. J. Electrochem. Soc. 148, 640–643.

Durant, J.L., Busby Jr., W.F., Lafleur, A.L., Penman, B.W., 1996. Human cell mutagenicity of oxygenated, nitrated, and unsubstituted polycyclic aromatic hydrocarbons associated with urban aerosols. Mutat. Res. 371, 123–157.

Elie, M.R., Geiger, C.L., Clausen, C.A., 2012. Reduction of benzo[a]pyrene with acid- activated magnesium metal in ethanol: a possible application for environmental remediation. J. Hazard. Mater. 203–204, 77–85.

Fallahtafti, S., Rantanen, T., Brown, R.S., Snieckus, V., Hodson, P.V., 2012. Toxicity ofhydroxylated alkyl-phenanthrenes to the early life stages of Japanese medaka (Oryzias latipes ). Aquat. Toxicol. 106, 56–64.

1280 M.R. Elie et al. / Chemosphere 91 (2013) 1273–1280

Gan, S., Lau, E.V., Ng, H.K., 2009. Remediation of soils contaminated with polycyclic aromatic hydrocarbons (PAHs). J. Hazard. Mater. 172, 532–549.

Grosjean, M.-H., Zidoune, M., Huot, J.-Y., Roue, L., 2006. Hydrogen production via hydrolysis reaction using ballmilled Mg-based materials. Int. J. Hydrogen Energy 31, 109–119.

Hall, S., Lipsky, S., McEnroe, F., Batels, A., 1971. Lithium–ammonia reduction ofaromatic ketones to aromatic hydrocarbons. J. Org. Chem. 36, 2588–2591.

Johnson, T.L., Scherer, M.M., Tratnyek, P.G., 1996. Kinetics of halogenated organic compound degradation by iron metal. Environ. Sci. Technol. 30, 2634–2640.

Joseph, J., Singh, S.C., Gupta, V.K., 2009. Morphology controlled magnesium ethoxide: influence of reaction parameters and magnesium characteristics.Part. Sci. Technol. 27, 528–541.

Kim, J.Y., Kim, H.D., Seo, M.J., Kim, H.R., No, Z., Ha, D.C., Lee, G.H., 2006. Reduction ofketones to corresponding alcohols with magnesium metal in absolute alcohols.Tetrahedron Lett. 47, 9–12.

Khurana, J.M., Sharma, V., Chako, S.A., 2007. Deoxygenation of sulfoxides,selenoxides, telluroxides, sulfones, selenones and tellurones with Mg-MeOH.Tetrahedron 63, 966–969.

Layshock, J.A., Wilson, G., Anderson, K.A., 2010. Ketone and quinone-substituted polycyclic aromatic hydrocarbons in mussel tissue, sediment, urban dust, and diesel particulate matrices. Environ. Toxicol. Chem. 29, 2450–2460.

Lemieux, C.L., Lambert, A.B., Lundstedt, S., Tysklind, M., White, P.A., 2008. Mutagenic hazards of complex polycyclic aromatic hydrocarbon mixtures in contaminated soil. Environ. Toxicol. Chem. 27, 978–990.

Lundstedt, S., White, P.A., Lemieux, C.L., Lynes, K.D., Lambert, I.B., Oberg, L., Haglund,P., Tysklind, M., 2007. Sources, fate, and toxic hazards of oxygenated polycyclic aromatic hydrocarbons (PAHs) at PAH contaminated sites. Ambio 36,475–485.

Matheson, L.J., Tranyek, P.G., 1994. Reductive dehalogenation of chlorinated methanes by iron metal. Environ. Sci. Technol. 28, 2045–2053.

Musa Bandowe, B.A., Shukurov, N., Kersten, M., Wilcke, W., 2010. Polycyclic aromatic hydrocarbons (PAHs) and their oxygen-containing derivatives (OPAHs) in soils from the Angren industrial area, Uzbekistan. Environ. Pollut.158, 2888–2899.

Nordlien, J.H., Ono, S., Masuko, N., Nisancioglu, K., 1997. A TEM investigation ofnaturally formed oxide films on pure magnesium. Corros. Sci. 39, 1397–1414.

Patel, U., Suresh, S., 2006. Dechlorination of chlorophenols by magnesium–silverbimetallic systems. J. Colloid Interface Sci. 299, 249–259.

Taub, I.A., Roberts, W., LaGambina, S., Kustin, K., 2002. Mechanism of dihydrogen formation in the magnesium water reaction. J. Phys. Chem. A 106, 8070–8078.

Tilstam, U., Weinmann, H., 2002. Activation of Mg metal for safe formation ofGrignard reagents on plant scale. Org. Process Res. Dev. 6, 906–910.

Vione, D., Maurino, V., Minero, C., Pelizzetti, E., Harrison, M.A.J., Olariu, R.I., Arsene,C., 2006. Photochemical reactions in the tropospheric aqueous phase and onparticulate matter. Chem. Soc. Rev. 35, 441–453.

Wanaratna, P., Christodoulatos, C., Sidhoum, M., 2006. Kinetics of RDX degradation by zero-valent iron (ZVI). J. Hazard. Mater. 136, 68–74.

Wang, Z., Huang, W., Fennell, D., Peng, P., 2008. Kinetics of reductive dechlorination of 1,2,3,4,-TCCD in the presence of zero-valent zinc. Chemosphere 71, 360–368.

Wang, G.H., Kawamura, K., Zhao, X., Li, Q.G., Dai, Z.X., Niu, H.Y., 2007. Identification,abundance and seasonal variation of anthropogenic organic aerosols from amega-city in China. Atmos. Environ. 41, 407–416.

Xia, T., Korge, P., Weiss, J.N., Li, N., Venkatesen, M.I., Sioutas, C., Nel, A., 2004.Quinones and aromatic chemical compounds in particulate matter induce mitochondrial dysfunction: implications for ultrafine particle toxicity. Environ.Health Perspect. 112, 1347–1358.

Xiao, D., Pan, B., Yu, M., Liu, Y., Zhang, D., Peng, H., 2012. Sorption comparison between phenanthrene and its degradation intermediates, 9,10- phenanthrenequinone and 9-phenanthrol is soils/sediments. Chemosphere 86,183–189.