COMPOSITIONS AND METHODS FOR

PERFLUORO ALKYL ACID REMEDIATION

CROSS-REFERENCE TO RELATED APPLICATION

This application claims benefit of U.S. Provisional Application No. 62/637,929, filed March 2, 2018, which, is hereby incorporated herein by reference in its entirety.

STATEMENT REGARDING FEDERALLY SPONSORED

RESEARCH

This invention was made with Government support under W912CG- ll-C-0022 awarded by the U.S. Army. The government has certain rights in the invention.

FIELD OF THE INVENTION

The field of the invention generally relates to compositions, systems, and methods of enzyme-mediated degradation of target compounds.

BACKGROUND OF THE INVENTION

Perfluoroalkyl acids (PFAAs) are a group of man-made organic chemicals that are extremely stable and have been widely used in nearly every aspect of our lives (Key, et al., Environmental Science & Technology, 31 (9): 2445-2454 (1997)). A great deal of attention has been heightened toward these chemicals because of their global distribution, recalcitrant to natural degradation, and potential toxicity (Pistocchi, et al., Environmental Science & Technology, 43 (24): 9237-9244 (2009); Houtz, et al.,

Environmental Science & Technology, 47: 8187-8195 (2013); Merino, et al., Environmental Engineering Science, 33: 615-649 (2016)). The extreme thermal and chemical stability of PFAAs arises from the high energy carbon- fluorine bond (531.5 kJ/mol) (Hudlicky, et al., Chemistry of Organic Fluorine Compounds II: A Critical Review, American Chemical Society (1995)) and the strong shielding effect of the helical conformation of their molecular structures (Torres, et al., Chemosphere, 76 (8): 1143-1149 (2009)). PFAAs have been extensively applied in industry and consumer products, such as food packaging, firefighting foam formulation, and semiconductor production (Paul, et al., Environmental Science &

Technology, 43 (2): 386-392 (2008)).

Perfluorooctanoic acid (PFOA) and perfluorooctanesulfonate (PFOS) are the two PFAAs that are most frequently detected in the environment and their co-occurrence has been found in groundwater plumes of substantial sizes (Gonzalez-Gaya, et al., Environmental Science & Technology, 48 (22): 13076-13084 (2014); Xiao, et al., Water Research, 72:64-74 (2015)). The major concerns over PFOA and PFOS are their ability to bioaccumulate and further induce a variety of undesirable effects in animals including carcinogenesis, infertility, endocrine disruption, and immunotoxicity (Final Report. 3M Company, US EPA Administrative Record (2002); Grandjean, et al., Environmental Health, 12(1) 35 (2013)). In particular, PFOS was found approximately 10 times more toxic than PFOA to freshwater organisms (Ji, et al., Environmental Toxicology and Chemistry, 27(10): 2159-2168 (2008)) and having a longer elimination half-life of 5.4 years than PFOA (3.8 years) in human serum (Olsen, et al., Environmental Health Perspectives, 115(9): 1298-1305 (2007)). PFOS has been included in the Annex B of the

Stockholm Convention on Persistent Organic Pollutants (POPs) in 2009. The U.S. Environmental Protection Agency (EPA) has classified PFOA and PFOS as emerging contaminants (Post, et al., Environmental Research, 116(0): 93-117 (2012)). Continuous increase in PFOS concentrations at remote regions and in marine organisms was predicted by a global-scale fate and transport model (Armitage, et al., Environmental Science & Technology, 43(24): 9274-9280 (2009)).

The unique structure of PFOS impedes it from being degraded by traditional water treatment processes and utilized by microbes. Because PFOS is more stable than PFOA (Zhang, et al., Environmental Science & Technology, 47(12): 6471-6477 (2013)), and several technologies capable of degrading PFOA, i.e. persulfate oxidation (Park, et al., Chemosphere, 145: 376-383 (2016)) and electrolysis on Ti/Sn02-Sb electrode (Zhuo, et al., Environmental Science & Technology, 45(7): 2973-2979 (2011)), were proven ineffective toward PFOS. In the aquatic environment, PFOS has a half-life of more than 41 years (Hekster, et al., Report RIKZ/2002.043 Prepared at the University of Amsterdam and RIKZ (The State Institute for Coast and Sea) 99 (2002)). There has not been a report to date indicating biotic or abiotic degradation of PFOS under environmentally relevant conditions (Ochoa-Herrera, et al., Environmental Science: Processes & Impacts, 18(9): 1236-1246 (2016)). While numerous studies have aimed to develop means to degrade PFOS for treatment and remediation purposes. Sonochemical oxidation (Cheng, et al, Environmental Matrix Effects.

Environmental Science & Technology, 42 (21): 8057-8063 (2008)), UV photolysis (Vecitis, et al., Environ. Sci. Eng, 3(2): 129-151 (2009)), and electrolysis (Carter, et al., Environmental Science & Technology, 42(16): 6111-6115 (2008)) are able to decompose PFOS. Reductants such as zero- valent iron (Hori, et al., Environmental Science & Technology, 40(3), 1049- 1054 (2006)) and vitamin B12 (Ochoa-Herrera, et al., Environmental Science & Technology, 42(9); 3260-3264 (2008)) have been used to reduce PFOS. However, none of these reaction mechanisms is expected to degrade PFOS in natural conditions. Their applications for treatment or remediation are restricted by the requirement of special equipment and high-energy input.

It is an objection of the invention to provide compositions and methods of use thereof for treatment and remediation of contaminants such as perfluoroalkyl acids.

SUMMARY OF THE INVENTION

Compositions and methods for degrading perfluorinated substances (PFASs) compounds such as perfluoroalkyl acids (PFAAs) are provided. For example, a method of degrading perfluorooctanesulfonate (PFOS) can include contacting material including PFOS with an effective amount of one or more enzymes to increase degradation of the PFOS by an enzyme catalyzed oxidative humification reaction (ECOHR). The material can be in an aqueous environment or a non-aqueous environment. Aqueous environments include, for example, groundwater, aquifers, surface water courses, and subsurface water courses. Non-aqueous environments can include, for example, soil, earth, a building or part thereof, a refuse dump, and transportation infrastructure.

Methods of degrading perfluorooctanoic acid (PFOA) in a non- aquatic environment are also provided. Such methods can include, for example, contacting the non-aquatic environment having a PFOA with an effective amount of one or more enzymes to increase degradation of the PFOA by an ECOHR.

Also provided are compositions and devices including granular activated carbon (GAC) in combination with and one or more ECOHR enzymes, optionally one more or more mediators, and optionally one or more enhancers. The GAC can be functionalized with the ECOHR enzyme. In some embodiments, the GAC is housed in a device such as a column.

Methods of using GAC and devices comprising GAC to degrade target molecules are also provided. Target molecule can be, for example, per- and polyfluoroalkyl substances (PFAS) such as

perfluorooctanesulfonate (PFOS), perfluorooctanoate (PFOA), perfluoro-n- butanoic acid (PFBA), perfluoro-n-pentanoic acid (PFPeA), perfluoro-n- hexanoic acid (PFHxA), perfluoro-n-butane sulfonate (PFBS), and perfluoro- n-hexane sulfonate (PFHxS), or a sulfonamide such as sulfadimethoxine.

Methods can include, contacting the target molecule (e.g., material contaminated with the target molecule), with the GAC. ECOHR enzyme can be contacted with the GAC before, during, or after the target molecule. Thus, in some embodiments, the GAC entraps, absorbs, or is functionalized with ECOHR enzyme and optionally mediator and/or enhancer, and subsequently target molecule is contacted with the GAC. In some embodiments, the GAC entraps, absorbs, or is functionalized with target molecule and subsequently ECOHR enzyme and optionally mediator and/or enhancer is contacted with the GAC. In some embodiments, target molecule and ECOHR enzyme and optionally mediator and/or enhancer are contacted with the GAC at the same time. In some embodiments, target molecule and ECOHR enzyme and optionally mediator and/or enhancer are in the same material. An exemplary material is, but is not limited to, groundwater. In some embodiments, the target molecule and ECOHR enzyme and optionally mediator and/or enhancer are in separate materials.

ECOHR systems for use with the disclosed methods and devices are also provided. The ECOHR systems include one or more ECOHR enzymes and optionally one or more mediators, enhancers, or a combination thereof. Exemplary enzymes for use in the disclosed methods include laccase enzymes, peroxidase enzymes, and combinations thereof. The mediator is typically contacted with the material or area of treatment in an effective amount to increase the rate of the ECOHR reaction. The mediator can be a phenolic or anilinic compound. Exemplary mediators include granular activated carbon, catechol, guaiacol, violuric acid, l-Hydroxy-benzotriazole (HBT), N-hydroxyphthalimide (HPI), 2, 2'-azine-bis (3- ethylbenzothiazoline- 6- sulfonic acid) (ABTS), 2,2,6,6-tetramethyl-l- piperidinyloxyl (TMPO), daidzein, gallic acid, vanillin, apocynin. p- coumaric acid, sinapic acid, acetovanillone, ferulic acid, salicylic acid, and syringaldehyde. In some embodiments, at least one mediator is present in, or a component of, soybean meal extract (SBE) or humate (HM), and the mediator is contacted with the material by contacting soybean meal extract (SBE) or humate (HM) with the material. The enhancer can be contacted with the material or area of treatment in an effective amount to form a complex with the target compound (e.g., a PFAA) and enhance degradation thereof. Preferred enhancers include multivalent metal ions, or a source thereof for example in the form of a salt. Exemplary multivalent metal ions include, but are not limited to, Cu ²⁺ , Mg ²⁺ , and Fe ³⁺ .

The components of the ECOHR system including an enzyme, and optionally a mediator, an enhancer, and other additives can be contacted with the material or area of treatment separately or together in any combination or sub-combination. The components can be part of the same or different formulas and be contacted with the material or area of treatment at the same or different times. Any of the components can be contacted with the material or area of treatment as a dried formulation or an aqueous formulation. In some embodiments, one or more components of the system are contacted with the material or area of treatment 2, 3, 4, 5, or more times, over a period of day, weeks, months, or years.

BRIEF DESCRIPTION OF THE DRAWINGS

Figures 1A-1B are line graphs showing the change of PFOS concentration in ECOHRs over time with the presence 10 mM Cu ²⁺ in the solution (1 A) or 10 mM Mg ²⁺ in the solution (1B). Control: the positive control sample to which no laccase or HBT was added; 1-0: 1 U/mL laccase added every 6 d but no HBT; 1-20: 1 U/mL laccase and 20 mM HBT added every 6 d. Error bars represent standard deviations (n = 3). Figure 1C is a line graph showing a pseudo-first-order rate model fit for the PFOS degradation in Cu ²⁺ and Mg ²⁺ solution with initial PFOS concentration of 1.0 mM and addition of 1 U/mL laccase and 20 pM HBT every 6 days.

Figure 2A is a line graph showing the change of PFOS concentration in ECOHRs over time with the addition of DIPPMPO as a HBT radical scavenger. Control: the positive control sample to which no laccase or HBT was added; 1-20: 1 U/mL laccase and 20 pM HBT added every 6 d; 1-20 RS: 1 U/mL laccase, 20 pM HBT, and 200 pM DIPPMPO added every 6 d. Error bars represent standard deviations (n = 3). Figure 2B is a line graph showing the HBT conversion over time during the ECOHRs. Cu 1-20: 1 U/mL laccase and 20 pM HBT added every 6 d in the 10-mM Cu ²⁺ solution; Mg 1-20: 1 U/mL laccase and 20 pM HBT added every 6 d in the 10-mM Mg ²⁺ solution.

Figures 3A-3B are UV-Vis differential absorbance spectra (DAS) calculated according to the data recorded at varying cation concentrations while maintaining PFOS concentration at 300 pM. Cu-PFOS DAS determined at pH 4.9 (3A) and Mg-PFOS DAS determined at pH 6.5 (3B).

Figures 4A-4C are molecular structures in water optimized by B3LYP/6-31 \+G(d,p) method on PFOS anion (4A), PFOS-Cu (4B), and PFOS-Mg (4C).

Figure 5 is a line graph showing fluoride release during 1-20 treatment in 10 mM Cu ²⁺ solution. Initial PFOS concentration 1.0 pM, 1 U/mL laccase and 20 pM HBT added every 6 d. The result of PFOS degradation is shown in Figure 1A. Error bars represent standard deviations (n = 3).

Figures 6A-6C are line graphs showing change of the normalized PFOA concentration during ECOHRs over time in different natural organic material extraction solutions (6A) soybean meal, (6B) humate, and (6C) mushroom compost. Control: the positive control sample without laccase or HBT; PO HBT: 1 U/mL of PO laccase and 20 mM HBT added every 6 d; PS HBT: 1 U/mL of PS laccase and 20 pM HBT added every 6 d; PO: 1 U/mL of PO laccase added every 6 d but no HBT; PS: 1 U/mL of PS laccase added every 6 d but no HBT. The normalized PFOA concentration was calculated by dividing the PFOA concentration in the sample at any sampling time by that at time zero. Figures 6D-6F are bar graphs showing statistic analysis results of the normalized PFOA concentration during ECOHRs over time (from left to right for each group, bars represent: 0 days, 12 days, 24 days, 36 days) in different natural organic material extraction solutions soybean meal (6D), humate (6E), and mushroom compost (6F). The same letter indicates no statistical difference at a=0.05, for the comparison of the same treatment over different sampling times. Control: the positive control sample without laccase or HBT; PO HBT: 1 U/mL of PO laccase and 20 pM HBT added every 6 d; PS HBT: 1 U/mL of PS laccase and 20 pM HBT added every 6 d; PO: 1 U/mL of PO laccase added every 6 d but no HBT; PS: 1 U/mL of PS laccase added every 6 d but no HBT. The normalized PFOA concentration was calculated by dividing the PFOA concentration in the sample at any sampling time by that at time zero.

Figure 7A is a line graph showing change of PFOA concentration in the soil slurry containing 1.0 g of 0.5 pg/g PFOA, 50 mg soybean meal, and 1.5 mL HPLC water. PO 20/4 wk: 20 U of PO was added every 4 weeks; PS 20/4 wk: 20 U of PS was added every 4 weeks; PO 60: 60 U of PO was added to the reactor at the beginning of the experiment only; PS 60: 60 U of PS was added to the reactor at the beginning of the experiment only. Figure 7B is a bar graph showing statistical analysis results of PFOA concentration in the soil slurry containing 1.0 g of 0.5 pg/g PFOA, 50 mg soybean meal, and 1.5 mL HPLC water over time (from left to right for each group, bars represent: 0 days, 28 days, 84 days, 140 days). The same letter indicates no statistical difference at a=0.05, for the comparison of the same treatment over different sampling times. PO 20/4wk: 20 U of PO was added every 4 weeks; PS 20/4 wk: 20 U of PS was added every 4 weeks; PO 60: 60 U of PO was added to the reactor at the beginning of the experiment only; PS 60: 60 U of PS was added to the reactor at the beginning of the experiment only.

Figures 8A-8B are line graphs showing change of activity of different laccase species in the soil slurry over time. The initial addition of PO and PS are 1.71 U and 1.78 U, respectively. The change of PO and PS enzyme activity in liquid and solid phase respectively (8A), and total laccase activity in soil slurry (8B). The total enzyme activity is calculated by addition of the enzyme activities in liquid and solid phases. The laccase enzyme activity was measured separately in liquid and solid phases and both were expressed in 2,6-dimethoxyphenol (DMP) units.

Figure 9A is sketch is of a micro-column packed with GAC. Figure 9B is a picture of a micro-column packed with GAC. Figure 9C is a picture of a setup of an exemplary bench scale“trap and treat” treatment system.

Figures 10A-10D are graphs showing the sorption isotherm of PFOA on Sigma Aldrich GAC (10A) and CalgonCarbon GAC (10B) or PS laccase on SigmaAldrich GAC (10C) and CalgonCarbon GAC (10D).

Figure 11 is a bar graph showing the removal of PFOA in spiked samples by a laccase-mediator system immobilized on SigmaAldrich GAC in batch reactors over time (from left to right for each group, bars represent: 15 days, 30 days, 60 days)). The total reaction volume is 15 mL with 20 mg GAC added. Concentrations of all the reagents were listed in Table 12.

Figure 12 is a line graph showing breakthrough of PFOA on the GAC column packed with ECOHR induced and a blank control GAC column. The influents contained 1 pmol/L PFOA w/wo PS laccase added at two dosages (PS2: 2 Unit/mL; PS4: 4 Unit/mL) and w/wo HBT added at 20 pmol/L. Figures 13A-13C are line graphs showing the breakthrough curves of PFOA (0.87 pg/L) and PFOS (3.92 pg/L) (13A), HBT (13B) (2 pmol/L), and PS laccase (2 Unit/mL) (13C) on CalgonCarbon GAC (0.1 g) columns.

Figures 14A-14B are line graphs showing breakthrough of PFOA (14A) and PFOS (14B) in groundwater on the GAC column packed with ECOHR induced and a blank control. The columns were packed with 0.5 g CalgonCarbon GAC, and the flow rate was 0.14 mL/min. The influent solutions contain groundwater, w/wo HBT at 20 pmol/L, and w/wo PS laccase at 2 unit/mL.

Figure 15 is a bar graph showing the removal of PFOA in groundwater by laccase-mediator system immobilized on CalgonCarbon GAC in batch reactors. The reaction volume is 45 mL containing 10 mg GAC (from CalgonCarbon), 2 Unit/mL PS laccase, and with 20 pmol/L HBT (GW+GAC+HBT+PS) or without HBT (GW+GAC+PS). Control reactors were prepared and tested at the same time that contained 10 mg GAC but not HBT and laccase (GW+GAC) or no laccase (GW+GAC+HBT) in 45 mL groundwater.

Figure 16 is the FTIR spectra of CalgonCarbon GAC adsorbed with PFOA only (1 pmol/L initial concentration) or PFOA and PS laccase with initial concentration of 1 pmol/L PFOA and 2 Unit/mL for laccase.

DETAILED DESCRIPTION OF THE INVENTION

I. Definitions

As used herein the terms“isolated,”“isolating,”“purified,” “purifying,”“enriched,” and“enriching,” when used with respect to ECOHR components of interest (e.g., ECOHR enzyme), indicate that the component at some point in time were separated, enriched, sorted, etc., from or with respect to other cellular material to yield a higher proportion of component compared to the other cellular material, contaminates, or active agents. “Highly purified,”“highly enriched,” and“highly isolated,” when used with respect to component of interest, indicates that the component is at least about 70%, about 75%, about 80%, about 85%, about 90% or more, about 95%, about 99% or 99.9% or more purified or isolated from other cellular materials, contaminates, or active agents.“Substantially isolated,” “substantially purified,” and“substantially enriched,” when used with respect to nucleic acids of interest, indicate that the nucleic acids of interest are at least about 70%, about 75%, or about 80%, more usually at least 85% or 90%, and sometimes at least 95% or more, for example, 95%, 96%, and up to 100% purified or isolated from other cellular materials, contaminates, or active agents.

As used herein,“degrade” or“degrading” with respect to target contaminant indicates that the ECOHR system is able to break-down portions of the chemical structure of the target contaminant or otherwise act to reduce the amount (measured by weight, thickness, or other measureable variable) in a material (e.g., water, soil, etc.) as compared to material not treated with the ECOHR system or the same material prior to treatment with the ECOHR system.

As used herein, the terms“application,”“administer,” and/or “treatment” with respect to the ECOHR systems refers to the act of contacting a specimen/sample/material/site (e.g., water, soil, etc.) with an ECOHR system.

II. ECOHR Systems

Enzyme Catalyzed Oxidative Humification Reactions (ECOHRs) are important in natural humification processes (Held, et ak, Environmental Technology, 18(5): 479-487 (1997)), catalyzed by extracellular

phenoloxidases and peroxidases which are ubiquitously present in the environment (Bollag, Metal Ions in Biological Systems, 28: 205-217 (1992)), mediating degradation of lignocellulosic materials and formation of humic substances. It has been found that many POPs, such as polycyclic aromatic hydrocarbons (PAHs) (Pozdnyakova, et a , Enzyme and Microbial

Technology, 39(6): 1242-1249 (2006)) and polychlorinated biphenyls (PCBs) (Colosi, et ak, Environ. Sci. Technol, 41(3): 891-896 (2006)) can be degraded by ECOHRs, and covalently bound to natural organic matter (NOM) to result in their detoxification (Bollag, Environmental Science & Technology, 26(10): 1876-1881 (1992)). Compositions, systems, and method of use thereof for ECOHR-based reduction or degradation of perfluoroalkyl acids (PFAAs) such as perfluorooctanesulfonate (PFOS) are provided. The system includes an ECOHR enzyme that can reduce or degrade a target PFAA such as PFOS. The system can also include one or more mediators, enhancers, or a combination thereof that can increase the rate of PFAA degradation by the system. As discussed in more detail below, the methods typically include contacting material contaminated with, or otherwise containing a target molecule, such as a PFAA, with an effective amount of the components of an ECOHR system to increase the reduction or degradation the target molecule in the material, relative the material if left untreated. In some embodiments, the contaminated material is contacted with the components separately.

Thus, in some embodiments, the system is formed or complete when the last component is contacted with the contaminated material. In some

embodiments, two or more components are contacted with the contaminated material together. In some embodiments, a complete ECOHR system is contacted with the contaminated material at once.

A. Enzymes

Extracellular enzymes such as laccase, horseradish peroxidase (HRP), and lignin peroxidase (LiP) present in soils can catalyze such reactions between a wide range of chemicals via oxidative coupling reactions (Gramss et al., Chemosphere, 38:1481-1494 (1999). It is believed that phenolic moieties, which are present in natural organic matter in soils and aquatic systems, serve as substrates for these enzymes. Enzymatic oxidation of such phenolic compounds produces phenoxy radicals that may cross couple with aromatic amines during the process of humus formation (Bollag and Dec, J Environ Qual, 29:665-676 (2000)). Oxidation of phenolic compounds may also generate electron poor sites that can be attacked by nucleophilic aromatic amines (Thom and Kennedy, Environ Sci Technol, 36:3787-3796 (2002)). Peroxidase-mediated oxidative polymerization of phenolic and anilinic compounds has been identified in previous studies and indicated as a potential means to groundwater and soil remediation (Palomo and Bhandari, J Environ Qual, 40:126-132 (2011).

The disclosed ECOHR systems include an ECOHR enzyme. In the most preferred embodiments, the enzyme of the ECOHR enzyme is a laccase enzyme or a peroxidase. The enzyme is contacted with material containing a target molecule in an effective amount to increase the reduction or degradation the target molecule in the material, when alone or when in combination with other components of an ECOHR system, relative the reduction or degradation of target molecule in the material if left untreated.

For example, the enzyme can be administered in concentrated or non concentrated form that is about 0.1 unit/mL to about 10 unit/mL in a liquid or aqueous material such as water, or about 1 unit/g to about 200 unit/g in a non-liquid or non-aqueous material such as soil, after being combined, mixed, or contacted with the contaminated material.

Laccase activity can be assessed by measuring the oxidation of 1 mM of 2,6-dimethoxyphenol (DMP) in a citrate phosphate buffer (pH 3.8) with the absorbance change measured at 468 nm. One unit can be defined as the amount of enzyme that causes a unit absorbance change per minute in 3.4 ml of this solution in cuvette with 1 cm light path.

1. Laccases

Laccases are multicopper-containing monophenol monooxygenase that catalyze the oxidation of phenolic substances with concomitant reduction of oxygen to water. The substrates of laccases are mainly chemicals with phenolic or anilinic moieties Galli, et al., Journal of Physical Organic Chemistry, 17, 973-977 (2004), Canas, et al., Biotechnology Advances, 28, 694-705 (2010)), which form free radicals or quinones upon laccase oxidation and then couple with each other to polymerize during humification reactions (Du, et al., Biomacromolecules, 14, 3073-3080 (2013), Piccolo, et al., Naturwissenschaften, 87, 391-394 (2000)). The active free radicals formed from some small molecular substrates can also react with inert chemicals that are not themselves direct laccase substrates, and incorporate them into humification process. These small molecular mediators (i.e., mediators) are discussed in more detail below. Any suitable laccase enzyme can be used. In some embodiments, the laccase is from a bacteria or fungi. For example, in some embodiments, the fungal laccase enzyme is from white rot fungi. In particular embodiments, the fungal laccase enzyme is from, pleurotus ostreatus, Pycnoporus sp. SYBC-L3 (PS), or Trametes versicolor.

In some embodiments, the laccase enzyme is the partially purified laccase enzyme produced by fungal fermentation and/or membrane purification. See, for example, Liu, et al., Journal of Cleaner Production, 39 154-160 (2013), which describes how to prepare a crude laccase solution concentrated from the fermentation broth of fungus.

In some embodiments a formulation includes a concentration of about 0.1 unit/ml to 2000 unit/mL laccase enzyme.

2. Peroxidases

In some embodiments the ECOHR enzyme is a perodixase.

Peroxidases are hemoproteins containing an ironporphyrin catalytic center that follows the classical catalysis mechanism of a peroxidase that produces free radicals from phenolic substrates (Banci et al., Biochemistry, 38:3205- 3210 (1999)). Exemplary peroxidases are lignin peroxidase and horseradish peroxidase.

3. Sources of Enzymes

Sources of enzyme include, but are not limited to, natural sources of peroxidase or laccase enzyme as well as another cell or organism, such as, for example, e. coli, that is adapted to produce the enzyme, e.g., genetically engineered by transformation with a construct containing a gene encoding a laccase or peroxidase. In particularly preferred embodiments laccase enzyme is produced by fungal fermentation and/or membrane purification. See, for example, Liu, et al., Journal of Cleaner Production, 39 154-160 (2013). In some embodiments, the enzyme is isolated or purified, or substantially isolated or purified from its source before forming part of an ECOHR system or used in an application thereof.

B. Mediators

The disclosed systems and methods typically employ one or more laccase mediators. Laccase mediators are generally compounds that laccase can convert to its free radicals, that can either be quenched via self-coupling or oxidation to, for example, benzotriazole, or as in the case of remediation, attack other non-substrate chemicals such as PFOS. In some embodiments, mediators can form active intermediates such as free radicals and quinones, which can further react with other recalcitrant organic matters that cannot be directly oxidized by laccase.

Redox mediators can be, for example, low molecular weight phenols that produce active free radicals. The total phenolic content may be positively correlated to oxidation efficiency as most of the natural mediators are phenolic compounds. The mediator or mediators is typically present in the system, composition, or method of use in an effective amount to provide free radicals to attack and reduce PFOS or another non-substrate chemical. The precise amount of mediator used can be determined by the practitioner based on, for example, the mediator or mediators selected, the amount of material subject to radiation, and the amount of laccase used.

The mediator can be natural or synthetic. Exemplary mediators include, but are not limited to, granular activated carbon, catechol, guaiacol, violuric acid, l-Hydroxy-benzotriazole (HBT), N-hydroxyphthalimide (HPI), 2, 2'-azine-bis (3-ethylbenzothiazoline-6-sulfonic acid) (ABTS), 2, 2, 6, 6- tetramethyl-l-piperidinyloxyl (TMPO), daidzein, gallic acid, vanillin, and apocynin. Plant-, especially legume-derived natural phenolic compounds (vanillin, p-coumaric acid, sinapic acid, acetovanillone, ferulic acid, salicylic acid, syringaldehyde) can also act as mediators, but are less efficient than synthetic mediators.

In some embodiments, the mediator is part of a mixture, for example, soybean meal extract (SBE) or humate (HM). In some embodiments, mediator may is present in soil naturally, and may or may not be

supplemented. The extraordinary performance of soybean meal extract may be attributed to its high contents of nitrogen, phenolic compounds and divalent cations. Thus, in some embodiments, two or more mediators are used in combination. Co-existence of multiple mediators may lead to more than additive results in laccase-mediator systems. Enhanced reaction rate was observed in previous laccase oxidation study when ABTS and natural mediators were present in combination (Jeon, et al., Appl. Microbiol.

Biotechnol. 81, 783e790 (2008)). This is because not only the primary radical species but also the secondary species generated through chain reactions among primary radicals can attack the organic pollutants, leading to more efficient oxidation.

The mediator can be contacted with material containing a target molecule in an effective amount to increase the rate of the ECOHR reaction, when alone or when in combination with other components of an ECOHR system, relative the reduction or degradation of target molecule in the material absent the addition of the mediator.

For example, the mediator can be administered in concentrated or non-concentrated form that is about 1 mM to about 100 mM per mL of liquid or aqueous material such as water, or about 1 pM to about 100 pM per gram of non-liquid material such as soil, after being combined, mixed, or contacted with the contaminated material. In some embodiments, the mediator is reduced or omitted, particularly in soil where mediators are abundant in natural organic matter.

C. Enhancers

In some embodiments, the composition, system, or method is supplemented with organic carbon, nitrogen, one or more di- or multivalent cations, or a combination thereof. Preferred multivalent cations that can serve as enhancers are discussed in more detail below. The organic content in mixtures such as soybean meal extract or humate may provide natural chelating agents to complex with metals, thus facilitating its oxidation by enzyme (Schlosser, et al., Appl. Environ. Microbiol. 68, 35l4e352l (2002)). Thus, in some embodiments, one or more mediators and one or more enhancers can be provided simultaneously by utilizing a mixture such as soybean meal extract or humate.

1. Ions

Transition metal ions (e.g. Mn ²⁺ ) and metallic complexes can act as mediators when they were oxidized by enzyme (Heofer, et al., FEBS Lett. 451, l86el90 (1999); Schlosser, et al., Appl. Environ. Microbiol. 68, 35l4e352l (2002)). Studies show that certain multivalent metal ions including Cu ²⁺ and Fe ³⁺ can form a complex with PFOA enhance reduction thereof, while others such as Mg ²⁺ and Mn ²⁺ did not. The experimental results presented below indicate that Cu ²⁺ and Mg ²⁺ could serve as a bridge to bring the negatively charged PFOS and laccase into proximity, thus increasing the chance of radicals that are released from laccase to reach and react with PFOS. DFT modeling also shows that PFOS complexation to the metal ions could unlock its helical configuration, and decrease the C-C bond energy of PFOS, making it more prone to free radical attack.

Thus, the system or method includes one or more multivalent metal ions, and/or a source thereof. Multivalent metal ions and multivalent metal cations are used interchangeably. Multivalent metal ions are metal ions that have a charge of at least +2. Exemplary sources of multivalent metal ions are their salts. A salt can include a multivalent metal cation ionically bonded to an anion. Exemplary multivalent metal ions can be selected from transition metal cations, group IIA metal cations, and group IIIA metal cations. Exemplary anions can be selected from hydrogen carbonate, acetate, chlorate, perchlorate, formate, sulfate, hydrogen sulfate, sulfite, nitrate, nitrite, and halides (chlorides, bromides, iodides). It is to be understood that the selection of the multivalent metal cations and the anions is subject to the proviso that the formed salt is soluble in an aqueous medium. A salt can be considered soluble in an aqueous medium, if at least 10 g of the salt dissolve in 100 mL of water at room temperature and atmospheric pressure.

Exemplary salts include copper (II) sulfate, magnesium sulfate, an iron sulfate, magnesium chloride, copper (II) chloride, manganese oxides, manganese sulfate or natural materials that contain these components, such as mushroom compost, clays, etc.

The enhancer can be contacted with material containing a target molecule in an effective amount to increase the rate of the ECOHR reaction, when alone or when in combination with other components of an ECOHR system, relative the reduction or degradation of target molecule in the material absent the addition of the enhancer. In some embodiments, the enhancer is administered in an effective amount to form a complex with a target, serve as a bridge to bring a negatively charged target and laccase into proximity, thus increasing the chance of radicals that are released from laccase to reach and react with the target, form a complex with target to unlock its structure (e.g., helical configuration of PFOS), or a combination thereof.

For example, the enhancer (i.e. the multivalent cations) can be administered in concentrated or non-concentrated form that is about 1 mM to 50 mM of liquid or aqueous material such as water, or about 1 mmole/kg to 50 mmole/kg of non-liquid material such as soil, after being combined, mixed, or contacted with the contaminated material. In some embodiments, the mediator is reduced or omitted, where enhancers are naturally present.

2. Other Enhancers and Additives

In some embodiments, the system includes a source of carbon, a source of nitrogen, or a combination thereof. Any of the systems and methods can include water, and/or other carriers, stabilizers, diluents and/or other ingredients, such as, but not limited to, those that enhance degradation of organic matter or those that serve other purposes. Systems for soil remediation may also include fertilizers, weed killers, and the like.

In some embodiments, particularly for land-based (e.g., soil) related remediation, the system can include or be supplemented with peat or one or more components thereof such as one or more humic constituents. Generally speaking, peat is an accumulation of partially decayed vegetation or organic matter from natural areas called peatlands, bogs, mires, moors, or muskegs.

Studies show that peat can increase the laccese-based reduction of sulfadimethoxine (SDM) in soil both with and without the presence of a mediator. Increasing the concentration level of peat from 0 to 5 mg g ^-1 soil significantly reduced the extractable SDM when treated with laccase.

Inclusion of peat in the reaction mixture enhanced the extent of soil-bound SDM residue. Peat contains humic constituents that are highly reactive and may enhance the coupling reactions (Bollag et al., Sci Total Environ, 118:357-366 (1992); Bollag and Myers, Sci Total Environ, 123: 205-217 (1992)). Covalent binding of laccase-oxidized SDM with soil organic matter might have occurred and similar reactions have been reported in the past (Thorn and Kennedy, Environ Sci Technol, 36:3787-3796 (2002); Colon et al., Environ Sci Technol, 36: 2443-2450 (2002)).

D. Reaction Conditions

Preferably ECOHR system is deployed under conditions favorable for efficient ECOHR-based degradation of the target molecule. Generally, an environmental pH of about 2 to about 7 is preferred for laccase -based systems. In some embodiments, the pH is environmental pH is naturally within this range. The environmental pH can also be adjusted with additives. For example, pH buffer can be an additive, while soybean extract or natural organic matter in soil has pH buffer effect.

Generally the preferred temperature range is from above zero to about 50 degrees Celsius, while around 25-30 degrees Celsius is preferred.

III. Methods of Remediation

The disclosed systems can be used to reduce the presence of one or more perfluorinated substances (PFAS), preferably one or more

perfluoroalkyl acids (PFAAs), particularly perfluoroctanesulfonate (PFOS), perfluorooctanoic acid (PFOA), or a combination thereof at a site in need thereof.

Perfluoroalkyl acids (PFAAs) are globally distributed organic pollutants with long half-lives in the environment and the potency to induce undesirable adverse effects such as infertility, birth defects, and reduced immune function to animals and human (Betts, Environmental Health Perspectives, 115, A550-A550 (2007); Grandjean, et al., Environmental Health, 12: 35 (2013)). PFAAs can be introduced into soil through the conversion of PFAAs precursors presented in the aqueous film- forming foams (AFFFs) (Houtz, et al., Environmental Science & Technology, 47 8187-8195 (2013); Backe, et al., Environmental Science & Technology (2013)), landfill of PFAAs -containing wastes (Sepulvado, et al., Environmental Science & Technology, 45: 8106-8112 (2011)), and atmospheric wet deposition (Wania, et al., Science of The Total Environment, 160-161: 211-232 (1995)). PFAAs have extreme persistency and relatively high mobility (Shin, et al., Environmental Science & Technology, 45: 1435- 1442 (2011)), and thus their presence in soil may pose the risks of leaching to groundwater (Eggen, et al., Science of The Total Environment, 408: 5147- 5157 (2010)), transport to surface water via runoff (Houtz, et al.,

Environmental Science & Technology, 46:9342-9349 (2012)),

bioaccumulation in food chains (Stahl, et al., J. Agric. Food Chem., 61:1784- 1793 (2013)), and direct or indirect exposure to animals (Houde, et al., Environmental Science & Technology, 45: 7962-7973 (2011)).

The methods typically include contacting material having a PFAA therein or associate therewith, with an effective amount of an ECOHR system to increase the reduction or degradation of the PFAA compared to the material when left untreated. Typically, the PFAA is, or includes, PFOS. In some embodiments, the PFAA is, or includes, PFOA. Other target PFAS and PFAA include, but are not limited to, perfluoro-n-butanoic acid (PFBA), perfluoro-n-pentanoic acid (PFPeA), perfluoro-n-hexanoic acid (PFHxA), perfluoro-n-butane sulfonate (PFBS), perfluoro-n-hexane sulfonate (PFHxS), or the precursors that can lead to formation of these compounds during natural degradation. In some embodiments, one or more of the foregoing PFAAs are present in the contaminated material. In some embodiments, one or more of the foregoing PFAAs are absent in the contaminated material. By way of non-limiting example, in some embodiments, the material does not have PFOA therein or associate therewith.

Other targets of remediation include sulfonamides such as sulfadimethoxine. However, in some embodiments, the material in need of remediation does not include a sulfonamide in general, or sulfadimethoxine in particular.

Environmental remediation typically refers to the removal or neutralization of pollution or contaminants, e.g., PFAAs, from environmental media, e.g. soil, groundwater, sea water or surface water or man-made environments. The disclosed methods can be generally classified as in situ or ex situ. In situ remediation involves treating the contaminated site or location, while ex situ involves the removal of the contaminated material to be treated elsewhere. A typical remediation method includes contacting a site or location or a material in need of remediation with an effective amount of an ECOHR system to increase reduction or degradation of the contaminant·

Sites or locations in need of remediation are not restricted, although typically such sites or locations include, but are not limited to, groundwater, aquifers, surface water courses, subsurface water courses, soil, earth and costal and marine environments. Artificial (i.e., man-made) sites and locations may also in be included, e.g. buildings (domestic and industrial) intact, demolished or otherwise and their foundations, refuse dumps (domestic and industrial), transport infrastructure and so on. A material in need of remediation is a material present at or taken from such sites or locations.

The ECOHR system can be contacted with, or administered to, the site or location or a material in need of remediation as individual ECOHR system components or a combination composition thereof. In embodiments in which more than one ECOHR enzyme, mediator, and/or enhancer is used, each component may be contacted with the target undergoing treatment separately or together as a mixture. Thus, the system can be administered to the site in need of remediation as a single formulation, or a series of separate formulations that when brought together at the site of remediation form the system.

It may be advantageous to effect contact of one or more, or all, of the different components to be used with the target undergoing treatment more than once. In further embodiments it may, at certain times, be advantageous to effect contact by providing a continuous feed of ECOHR system components and/or contaminated material. In some embodiments, a treatment is repeat 2, 3, 4, 5, 6, 7, 8, 9, 10, or more times. In some embodiments, the treatment is repeated two or more time“N” days apart, wherein N is any integer between about 1 and about 365. The treatments can be, for example, 1, 2, 3, 4, 5, 6, 7, 10, 25, 50, 75, 100, 150, or 200 day apart. Generally, the frequency of repetition or supplementation is typically higher in liquid material (e.g., water) than in non-liquid material (e.g., soil) because enzyme can inactivate faster in water. Thus in some treatment regimens for liquid environments (e.g., water) enzyme is administered at a lower dosage and higher frequency than in a non-liquid environment (e.g., soil). In particular embodiments, treatment of a liquid environment (e.g., water) is repeated or supplemented about 1 to about 7 days apart, with 1 day or 2 days being preferred. In other particular embodiments, treatment of a non-liquid environment (e.g., soil) is repeated or supplemental about 1 about 100 days apart, with 25-75 days being preferred as it corresponds with the lifetime of the enzyme. A repeat or supplementation of a first treatment may include all of the previously administered ECOHR components, a subset thereof, or additional components. For example, in some embodiments, only enzyme, or enzyme and mediator, or enzyme, mediator, or enhancer, are contacted with the contaminated material.

Components to facilitate the use of the ECOHR enzyme e.g.

mediators, enhancers, and other additives, and/or components can be used alongside the enzyme in the methods of remediation, may be administered together with the enzyme, separately but contemporaneously with the enzyme or entirely separately to the enzyme.

An ECOHR system, or individual components thereof, can be provided in any convenient form. Liquid forms, e.g. aqueous or organic or a mixture of both, or dried forms, e.g. lyophilized forms, are specifically contemplated. An ECOHR system, or individual components thereof, can be formulated into a composition also including additives, e.g. preservatives, stabilizers, colorings, etc. Lyophilized forms can include one or more lyophilization excipients, e.g. salts (organic and inorganic), chemical elements, amino acids, carbohydrates (mono-, di-, oligo- and

polysaccharides), etc. Other additives include components of use in methods of remediation. Thus, in some embodiments, all components of the system are administered to the remediation site exclusively as liquid or exclusively in dried form. In some embodiments, the components of the system are administered to the remediation site as a combination of liquid and dried forms. In some embodiments a dried form is administered to a non-aqueous environment, and subsequently treated with water or buffer or another liquid.

An ECOHR system can be formulated so that it can be diluted with water or other ingredients before or during application. In embodiments, the composition may be a liquid formulation with the ECOHR enzyme diluted with water or other liquid carrier. Dried forms can be diluted with sand or inert ingredients.

The composition(s) can be administered to the remediation site using any reasonable method. For example, the system or components thereof in liquid form can be administered via a sprayer or other liquid application device. A dried formulation can be administered using, for example, a spreader (e.g., a drop spreader or broadcast spreader), other suitable device. In some embodiments, the contaminated material and the ECOHR system are mixed, agitated, blended, infused or otherwise combined to increase contact between the ECOHR system and the target.

The laccase enzyme and/or other components of the ECOHR system can be included as a component in combination with other lawn care products (example weed and feed products contain herbicide impregnated fertilizer, etc.). In some embodiments, the ECOHR system components are provided in a dry powder formulation or in formulations in which one or more components are encapsulated, immobilized to carriers, or modified with stabilizers and dispersants for application to the site. The compositions can include a particulate topdressing, where the enzyme and/or other components of the ECOHR system is immobilized to particles of the particulate topdressing. The particulate topdressing may include various topdressings used for land application, such as, for instance, sand, synthetic granules, diatomaceous earth, calcined clay, ground corn cobs or other organic materials, silica/quartz sand, zeolite, lassinite, resins, and the like. In some embodiments, the particulate topdressing is sand and one or more of the ECOHR system components, such as the ECOHR enzyme is immobilized to sand particles. In some embodiments, the component or components are immobilized to particles of topdressing (e.g., sand, or other natural or synthetic particulate material) by activating the sand or other particulate topdressing with a linking material, such as but not limited to, chitosan and/or gluteraldehyde to activate the particle surface for enzyme attachment. Next, the enzymes are immobilized to the particles via the chitosan and/or glutaraldehyde. In an example embodiment, the surfaces of the particles are first activated with polyethyleneimine followed by crosslinkage with gultaraldehyde to graft aldehyde groups onto the surface of the particle. Then enzymes can be covalently bonded to the particles by reaction between the aldehyde groups and free amino groups on protein surface. See, e.g., U.S. Published Application No. 2012/0079764.

In other embodiments, a layer-by-layer (LbL) assembly approach can be used to immobilize the enzymes on the particle surface. This embodiment involves alternate sorption of a polycation substrate, a polyanion substrate, and the enzyme onto the particles. For application of each layer pH can be controlled to provide the substrates and/or the enzyme with the appropriate charges. Each sorption step leads to a reversal of the terminal surface charge after adsorption of a new layer.

In some embodiments the particles are coated with alternating layers of a polycation, a polyanion, and the enzyme, such that the enzyme is captured between layers on the particle. One example of a conventional LbL method is described in U.S. Published Application No. 2012/0079764 where poly(allylamine hydrochloride) (PAH) and poly(sodium 4-styrenesulfonate) (PSS) are used as the polyanions and polycations, respectively. The pH of the enzyme solution is carefully adjusted to several units away from their respective isoelectric points to maintain a net negative or positive charge. Sequential polyelectrolyte/enzyme layers are deposited to form repetitive particle-PAH-PSS-enzyme or particle-PAH-PSS-PAH-enzyme sandwich.

For each assembly step, the polyelectrolyte/enzyme is allowed to equilibrate with the sand particles before the next layer is added.

In other embodiments, ECOHR enzymes can be immobilized to the particles of topdressing by methods known to those in the art for immobilizing proteins to solid surfaces. In embodiments where ECOHR enzyme is immobilized to a topdressing formulation, frequency of application may be reduced (e.g., to as little as about once every 12 months or even less, depending on the location, application, etc.).

The objective of the contacting step is to introduce the enzyme and other components of the ECOHR system to the site or location or a material in need of remediation in such a way that the enzyme can be active, and provide an environmental remediation effect. For example, when the ECOHR system includes laccase enzyme and HBT mediator for remediating PFOS, the remediation effect may be by, for example, laccase catalyzing the oxidation of a mediator such as HBT to generate free radicals, which attack the bonds of the PFOS, leading to the formation of a shorter-chain length perfluoroalkyl free radical, with or without the sulfonic functionality, and subsequently mineralized or bound to natural organic matter.

Once an ECOHR system has been introduced and allowed to act on the target undergoing treatment, natural environmental processes, e.g. the water cycle, tides, wind, biodegradative and photodegradative processes, may be relied upon to effect the reduction in contamination at the treatment site, location or material. In other embodiments, especially in the context of ex situ treatments, the method may include a step in which a target undergoing treatment is washed, typically with an aqueous vehicle of low environmental impact, e.g. water or an aqueous salt solution, and/or a step in which treated material is isolated/removed. Multiple cycles of contact, washing and/or isolation/removal may occur.

Delivery to the target site, location or material undergoing treatment may conveniently be achieved by flooding, spraying, or injecting the site or location or the material in need of remediation with a delivery vehicle containing the ECOHR system, typically an aqueous vehicle of low environmental impact e.g. an aqueous salt solution, or water. Treatment of contaminated materials may take place ex situ in more controlled conditions. In these embodiments the contaminated material may be added to the ECOHR system. In the ex situ treatments of the invention the contaminated material may be treated in a bioreactor containing the ECOHR system.

Thus, bioreactors containing an ECOHR system are also provided.

In some embodiments, the ECOHR system is deployed together with or immobilized on or in a particulate solid support.

Some ECHOR systems include granular activated carbon (GAC). GAC can serve a trap for PFAAs, but does not result in their degradation. GAC can be coupled to an ECOHR enzyme to both trap and degrade the target contaminants such as PFAAs. Using a granular activated carbon (GAC) can increase PFAA concentration and effective enzyme/contaminant contact time. In some embodiments, the enzyme and optionally a mediator, an enhancer, or a combination thereof are absorbed, coupled, or otherwise attached or combined with the granular activated carbon.

In some embodiments, a reactor, column, or other such device can be utilized to house the GAC and ECHOR system. For example, in the ex situ treatments, the contaminated liquid (e.g., water) can be passed through a granular activated carbon (GAC) column or other suitable device, to concentrate the target chemicals (PFAAs, and other target molecules) on the GAC column. The ECHOR active components can be loaded to the column in the contaminated water or in a formulation after the contaminants are loaded to induce ECOHOR on the GAC column to degrade PFAAs.

In some embodiments, the PFAA can be treated as a batch (e.g., without flow through the column or other such device). In some

embodiments, PFAA can be treated while the contaminated material is flowed through the column or other such device (e.g., flow-through).

The ECHOR system can be loaded onto the column or other such device before, during or after sorption of PFAA.

Laccase and HBT has less sorption than PFAA. Thus, in some embodiments, laccase and optionally a mediator such as HBT can be loaded after PFAA is loaded to GAC column. The ECHOR system can be incubated with the PFAA with or without passing through water to allow ECHOR treatment.

Such treatment compositions, systems, and methods are amenable to large scale remediation projects. For example, in a particular embodiments, contaminated groundwater is passed through a GAC column having at least 200 lb of GAC, followed by loading HBT at 3.5 mmole/Kg GAC and laccase at 85 unit/g GAC to induce ECHOR on GAC.

GAC is commercially available and sources include, for example, Sigma Aldrich GAC and Calgon Carbon GAC.

As described herein, different ECOHR enzymes or combinations of two or more ECOHR enzymes can be used. The selected enzyme or combination, or subsets thereof, may be administered to the remediation site together or separately, concurrently or in tandem. Studies show that the apparent total enzyme activity at time zero was 0.48 U for pleurotus ostreatus (PO) laccase and 0.83 U for Pycnoporus sp. SYBC-L3 (PS) in soil slurry, although a total of 1.71 U of PO or 1.78 U of PS was added. This is believed to be because a portion of laccase was adsorbed to soil, and the apparent activity of enzyme usually declines in adsorption status (Mateo, et al., Enzyme and Microbial Technology, 40:1451-1463 (2007)). The liquid phase activity accounted for the majority of the total apparent activity in soil slurry for PS laccase, while, for PO laccase, the contribution of liquid and solid phases tended to be equal. PO activity became undetectable after about two months of incubation in the soil slurry, while after 84 days the total activity of PS still remained 0.18 U. The relatively less inactivation of PS than PO in soil slurry may explain observed differences in PFOA

degradation patterns. Thus, because it is one of the key factors in ECOHRs, the ECOHR enzyme(s) can be selected based on its stability in the remediation environment (e.g., soil, water, etc.), alone or in combination with its activity in reducing or decomposing the target. The relative proportions of the ECOHR enzymes, may be same or different. By varying the proportions as well as the identity of ECOHR enzymes greater control over the proprieties of the remediation processes may be achieved. Examples

Luo, et al., Environ Pollut. 2017 May ;224: 649-657. doi:

l0.l0l6/j.envpol.20l7.02.050. Epub 2017 Mar 3; Liang, et al.,

Chemosphere, 2017 Aug; 181:320-327. doi:

l0.l0l6/j.chemosphere.20l7.04.100. Epub 2017 Apr 23; and all

Supplemental materials associate therewith, are specifically incorporated by reference herein in their entireties.

Example 1: PFOS degrades during ECOHRs.

Materials and Methods

Chemicals and reagents.

All chemicals used in the experiments were reagent grade or higher and used as received. Laccase from Pleurotus Ostreatus (EC 420-150-4), 2, 6-dime thoxyphenol (DMP), and l-hydroxybenzotriazole (HBT) were from Sigma-Aldrich (St. Louis, MO). The 5-diisopropoxy-phosphoryl-5-methyl-l- pyrroline-N-oxide (DIPPMPO) was purchased from Enzo Life Sciences (Farmingdale, NY). Perfluoroctanesulfonate (PFOS) was purchased from Inpofine Chemical Company (Hillsborough, NJ). Both branch and linear PFOS were quantified during the chemical analysis. Other perfluoroalkane sulfonates (PFSAs) and PFCAs and surrogate standard sodium perfluoro-l- [ ¹³ C ₈ ]-octanesulfonate (M8PFOS) were obtained from Wellington

Laboratories (Ontario, Canada) (see Table 1). The cupric/magnesium sulfates were obtained from the Fisher Scientific (Pittsburgh, PA). All the HPLC- grade organic solvent including acetonitrile, methanol, and dichloromethane were also purchased from Fisher Scientific (Pittsburgh, PA). Milli-Q water (18.2 MW/cm resistivity) was prepared using the Nanopure Stand purification system (Thermo Scientific, San Jose, U.S.)

Table 1: The molecular formula, retention time, precursor ion, transition ion, and detection limit of the perfluoroalkane sulfonates (PFSAs), perfluorocarboxylic acids (PFCAs), and HBT monitored in UPLC-MS/MS analysis. Precursor Transition Detection

Molecular RT

Chemicals ion ion limit

Formula (min)

PFBA (C4) 2.6 213 169 0.018

(CF ₂ ) ₃ COOH

PFPeA F-

4.5 263 219 0.027

(C5) (CF ₂ ) ₄ COOH

PFHxA F-

5.5 313 269 0.022

(C6) (CF ₂ ) ₅ COOH

PFHpA F-

6.1 363 319 0.008

(C7) (CF ₂ ) ₆ COOH

F-

PFOA (C8) 6.6 413 369 0.011

(CF ₂ ) ₇ C00H

F-

PFNA (C9) 7.0 463 419 0.15

(CF ₂ ) ₈ COOH

PFDA F-

7.3 513 469 0.10

(C10) (CF ₂ ) ₉ COOH

PFUA F-

7.6 563 519 0.012

(Cl l) (CF ₂ ) _IO COOH

PFBS (C4) F-(CF ₂ ) ₄ S0 4.8 299 99 0.021

PFHxS

F-(CF ₂ ) ₆ S0 6.2 399 99 0.064

(C6)

PFOS (C8) F-(CF ₂ )SS0 7.0 499 99 0.082

PFBA: perfluoro-n-butanoic acid; PFPeA: perfluoro-n-pentanoic acid;

PFHxA: perfluoro-n-hexanoic acid; PFHpA: perfluoro-n-heptanoic acid; PFNA: perfluoro-n-nonanoic acid; PFDA: perfluoro-n-decanoic acid; PFUA: perfluoro-n-undecanoic acid; PFBS: perfluorobutaesulfonate; PFHxS:

perfluorohexanesulfonate; PFOS: perfluorooctanesulfonate; HBT: 1- hydroxybenzotriazole. Experimental setup.

The working solution was prepared by dissolving 1.0 mM PFOS in 10 mM CuS0 ₄ or MgSC solutions. The pH values of the working solution were 4.9 and 6.5 for the Cu ²⁺ and Mg ²⁺ solution, respectively. Seventy-two reactors were prepared for each cation group. Each reactor included 10 mL working solution, 60 pL of 167 U/mL laccase stock solution (1 U/mL), and 20 pL of 10 mM HBT dissolved in acetonitrile (20 pM HBT: named as 1-20) or 20 pL acetonitrile (0 pM HBT: named as 1-0). The reactor with the addition of 60 pL of Mill-Q water and 20 pL of acetonitriles instead of laccase and HBT were designated as positive controls. Degradation experiments were conducted using time- sequenced, multiple-addition scheme at 22 °C with continuously shaking at 120 rpm in an incubator (Innova 42, New Brunswick Scientific). The whole experiment lasted for 162 days. Every six days, the reactors were dosed with the same amount of freshly prepared laccase solution (1 U/mL) and HBT in acetonitrile (20 pM HBT) or acetonitrile. At selected times, a set of nine reactors including triplicates from 1-20, 1-0 treatments and positive control were sacrificed and diluted with an appropriate volume of Milli-Q water to the final volume of 12.16 mL for all reactors. A 0.5-mL aliquot of the solution was withdrawn, adjusted the pH to approximately 10.5 with 60 pL of 200 pM NaOH, and then spiked with 0.5 mL of 0.5 pM M8PFOS before subjected to solid phase extraction (Benskin, et ak, Environmental Science & Technology, 46(11): 5815-5823 (2012)) cleanup as reported in previous studies (Luo, et ak, Environmental Science & Technology Letters, 2(7): 198-203 (2015);

Yamashita, et ak, Environmental Science & Technology, 38(21): 5522-5528 (2004); Yu, et ak, Water Research, 43(9): 2399-2408 (2009)). The detail process was described in the SI.

Laccase activity assay was reported by Park et ak (Park, et ak, Environ. Sci. Technol, 33(12): 2028-2034 (1999)). One unit of laccase activity is defined as the amount of enzyme that causes one unit change in absorbance at 468 nm per minute of a DMP solution at pH 3.8 in a l-cm light path cuvette (Johannes, et al., Applied and Environmental

Microbiology, 66(2):524-528 (2000)).

Laccase Activity Assay.

Laccase activity was assayed using 2,6-dimethoxyphenol (DMP) as substrate. One unit of activity of laccase equals the amount of enzyme that causes an absorbance change in 468 nm at a rate of 1.0 unit/min in 3.4 mL of 1 mM 2,6-dimethoxyphenol in citrate -phosphate buffer (pH 3.8) in a l-cm light path cuvette (Park, et al., Environmental Science & Technology 33(l2):2028-2034). All activities were reported as measured by this method in this study. Laccase activity was also assayed using a method with 2, 2’- Azino-bis(3-ethylbenzothiazoline-6-sulfonic acid) (ABTS) as the substrate to enable comparison with some earlier studies. Laccase activity was measured by monitoring the oxidation of 3 mL 100 mM ABTS substrate solution at 420 nm in 0.1 M Phosphate buffer (pH 6). One unit of laccase activity was defined as the amount of enzyme catalyzing the oxidation of 1 pmol ABTS per min (Johannes, et al., Applied and Environmental Microbiology

66(2):524-528 (2000)). By the measurements, the laccase activity measured by the two methods can be converted using the following equation:

Activity (measured by DMP ) = 2.65 x Activity (measured by ABTS) The laccase activity used in this study (1 U/mL by DMP method) is equivalent to 0.377 U/mL (by ABTS method).

Solid Phase Extraction.

The samples were subjected to solid phase extraction (SPE) (Oasis HLB SPE cartridges, 3 cc, 60 mg, Waters, Milford, MA) as described in earlier studies (Yamashita, et al., Environmental Science & Technology 38(2l):5522-5528 (2004); Yu, et al., Water Research 43(9):2399-2408 (2009)) with minor changes. The SPE cartridge was conditioned with 3 mL methanol, two 3 mL aliquots of Milli-Q water sequentially, followed by loading the sample, then rinsed with 3 mL Milli-Q water and blown to dry under vacuum. The cartridge was then eluted with 1 mL methanol for three times and then 1 mL acetonitrile for two times. All eluents were combined, and the mixture was blown to 1 mL with nitrogen gas for PFOS and HBT quantification. Recovery of the method has been evaluated using a standard addition method by adding 1 mM PFOA to selected sample aliquots before being processed as described above. The standard addition recovery was 83.69 ± 2.98% and 85.01 ± 2.63% for Cu ²⁺ and Mg ²⁺ solution respectively. Results

Laccases are a major group of phenoloxidases that is able to catalyze ECOHRs, and has wide industrial applications (Jeon, et al., Trends in Biotechnology, 31(6): 335-341 (2013)). Laccases catalyze the single-electron oxidation of phenolic or anilinic compounds (i.e. mediators) to form active intermediates such as free radicals and quinones, which can further react with other recalcitrant organic matters that cannot be directly oxidized by laccase (Baiocco, et al., Organic & Biomolecular Chemistry, 1 (1): 191-197 (2003)). Chemicals like l-hydorxybenzotriazole (HBT), ferulic acid, and vanillin have been commonly used as mediators in the laccase-mediator system since they have high reaction efficiency, low environmental impact, and representation of functionalities commonly present in NOM (Canas, et al., Biotechnology Advances, 28(6): 694-705 (2010)). Such laccase-mediator system has been applied in de-lignifying pulp, decoloring denim, and detoxifying textile dyes (Canas, et al., Biotechnology Advances, 28(6): 694- 705 (2010)) Laccase is able to mediate PFOA degradation using HBT as the mediator (Luo, et al., Environmental Science & Technology Letters, 2(7): 198-203 (2015)).

The experimental evidence herein shows PFOS degradation in the laccase-HBT system, and is believed to be the very first report of PFOS degradation under environmentally relevant conditions. The reaction kinetic was evaluated, the factors controlling PFOS degradation rate were examined, and the underlying mechanisms were proposed based on the degradation products identified by high-resolution mass spectrometry (HRMS) and molecular modeling.

A first experiment was carried out in a solution containing 1.0 mM PFOS and 10 mM Cu ²⁺ or Mg ²⁺ . Appropriate amounts of laccase and HBT were added to different reactors every six days to induce ECOHRs. As shown in Figures 1A-1B, PFOS was fairly stable in the control reactor in which neither laccase nor HBT is present. The treatment with the repetitive addition of 1 U/mL laccase alone (named as 1-0) followed the similar trend as the control, while continuous degradation of PFOS was observed in the treatment with the repetitive addition of 1 U/mL laccase and 20 mM HBT (named as 1-20) regardless of the types of cations present in the solution.

Certain cations play an important role in PFOA degradation by ECOHRs (Luo, et al., Environmental Science & Technology Letters, 2(7): 198-203 (2015)). The extent of PFOS degradation was also found to be cation dependent: 59.0 ± 7.42% of PFOS was degraded in Cu ²⁺ solution and 34.5 ± 7.27% in Mg ²⁺ solution after 162 days of incubation. The decomposition of PFOS in both 1-20 treatments followed the pseudo-first- order kinetics. The reaction rate constants (k) were 0.0043/d (r ² = 0.863) and 0.0027/d (r ² = 0.925) for the Cu ²⁺ and Mg ²⁺ solutions, respectively (Figure 1C).

Example 2: Benzotriazole nitroxyl (BTNO) radicals drive PFOS degradation.

Materials and Methods

Three groups of reactors were set up and processed using the same procedure as described above with 1 U/mL laccase and 20 mM HBT added every six days, except that 200 pM DIPPMPO was also added to one set of the reactors each time by adding 40 pL of a 500 mM DIPPMPO solution in Milli-Q water, and the same amount of water was added to the other reactors without DIPPMPO.

Results

An additional experiment was conducted to investigate the PFOS degradation in ECOHRs with the present of DIPPMPO, a spin trap that can effectively scavenge BTNO (Zoia, Journal of Physical Organic Chemistry, 23(6): 505-512 (2010)).

Laccase catalyzes the oxidation of HBT to generate BTNO radical. Experiments were designed to test the effect of BTNO radicals generated by the laccase-HBT reaction on PFOS. The 1-20 treatment was carried out in the 10-mM Cu ²⁺ solution in the absence and presence of an efficient scavenger of BTNO radical, 5-diisopropoxy-phosphoryl-5-methyl-l- pyrroline-N-oxide (DIPPMPO) (Zoia, Journal of Physical Organic

Chemistry, 23(6): 505-512 (2010)). It was evident that PFOS degradation was considerably suppressed in the 1-20 treatment with DIPPMPO present in the solution, while significant reduction (43.7 ± 7.34%) of PFOS concentration was found in the corresponding 1-20 treatment without DIPPMPO over 108 days of incubation (Figure 2A). This phenomenon strongly supports the conclusion that BTNO radicals induce PFOS degradation.

The consumption of HBT during both 1-20 treatments shown in Figures 1A-1B was documented (Figure 2B). It was found that the total quantities of HBT converted in Mg ²⁺ and Cu ²⁺ systems were not significantly different (5.79 ± 0.90 pmole for Cu ²⁺ and 6.38 ± 0.72 pmole for Mg ²⁺ ), however the PFOS transformation efficiency was significantly higher in the Cu ²⁺ system (k = 0.0043/d) than in the Mg ²⁺ system (k = 0.0027/d). A previous study reported that the formation of a complex between Fe ³⁺ and PFOS enhanced the photodegradation of PFOS (Jin, et al., Journal of Hazardous Materials, 271: 9-15 (2014)).

Example 3: Cations can enhance PFOS degradation by ECOHRs.

Materials and Methods

Theoretical Computation Details.

The calculation of optimized geometry were performed by Density Functional Theory (DFT) (5-7) at B3LYP/6-3 1 \ +G(cl,p) level of theory (Becke, J. Chem. Phys, 98: 5648-5652 (1993); Lee, et al., Phys. Rev. B, 37:785-789 (1988)). Vibrational analysis indicates all structures are stable and at the minimum point on the potential energy surface. In order to simulate the behavior in water solution, Polarizable Continuum Model (PCM) optimizations were performed for all studied systems (Cances, et al., J. Chem. Phys, 107: 3032-3041 (1997)). All calculations were performed using Gaussian 09 software (Frisch, Gaussian 09, Revision B.01. (2009)). Results

A differential absorbance spectra (DAS) approach (Yan, et al., Environmental Science & Technology, 49(14): 8323-8329 (2015)), a tool that is commonly used to probe the complexation between organic constituents and metal ions (Wania, et al., Science of The Total Environment, 160-161, 211-232 (1995)), was adopted to investigate possible interactions between PFOS and Cu ²⁺ or Mg ²⁺ in the ECOHR systems. The differential UV-Vis absorbance spectra ranged from 220 to 620 nm (Figures 3A-3B) exhibit subtle but consistent changes when PFOS was mixed with Cu ²⁺ or Mg ²⁺ at increasing concentrations. As illustrated in Figures 3A-3B, a distinct positive peak is shown in the DAS at the wavelength of 313 nm for both Cu-PFOS and Mg-PFOS systems, and the peak intensity increases with the increase of Cu ²⁺ and Mg ²⁺ concentration. The peak may reflect the shift of electronic density in PFOS molecule by complexation with Cu ²⁺ and Mg ²⁺ (Yan, et al., Environmental Science & Technology, 49(14): 8323-8329 (2015); Yan, et al., Environmental Science & Technology, 48(6): 3177-3185 (2014)). Such a result confirms that Mg ²⁺ and Cu ²⁺ can interact with PFOS, as Fe ³⁺ does (Jin, et al., Journal of Hazardous Materials, 271: 9-15 (2014)). A peak at 277 nm with gradually decreasing intensity shows up only in the Cu-PFOS system, but not in the Mg-PFOS system, indicating that PFOS may interact differently with Mg ²⁺ and Cu ²⁺ , therefore leading to the difference in PFOS degradation efficiency.

An attempt was made to explore the complexation between the cations and PFOS using density functional theory (DFT) method. The molecular structures of PFOS anion and the PFOS -cation complexes were optimized by B3LYP method at 6-3 1 1 +G(d,p) level. A comparison of the molecular geometries is provided in Figures 4A-4C and the calculated structural parameters are listed in Tables 3-5. The optimized structure of PFOS anion has a helical conformation (Figure 4A). After complexation to Cu ²⁺ or Mg ²⁺ , the bond lengths in PFOS anion, including C-S and C-C, vary only slightly (Table 5), but its conformation changes considerably.

Comparing to the dihedral angels S(26)C(8)C(l)C(5) of PFOS (-109.45°), those of PFOS-Cu and PFOS-Mg decrease to -104.81° and -106.78°, respectively, while the other dihedral angles decrease as well (Table 2). This shows that complexation to Cu ²⁺ and Mg ²⁺ tends to unlock the helical structure of PFOS, thus probably making the access by BTNO easier.

Moreover, the DFT calculation of bond energy indicates that such

complexation also reduce the C-C bond energy (Table 6)

Table 2. Partial dihedral angles of PFOS-Cu and PFOS-Mg complexes in water optimized by B3LYP/6-311+G** method

PFOS PFOS-Cu PFOS-Mg

Dihedral angles

angles/ ⁰ angles/ ⁰ D angles/ ⁰ D

S(26)C(8)C(l)C(5) -109.45 -104.81 4.64 -106.78 2.67

C(5)C(l)C(8)C(7) -85.15 -83.11 2.04 -81.08 4.07

S(26)C(8)C(6)C(4) -57.41 -52.84 4.57 -54.44 2.97

S(26)C(8)C(3)C(2) -84.96 -80.51 4.45 -81.85 3.11

Table 3. The bond lengths of PFOS anions, PFOS-Cu, and PFOS-Mg complexes in water obtained by DFT-B3LYP/6-3 1 \ +G(d,p).

PFOS PFOS-Cu PFOS-Mg

Bonds Bond lengths Bond lengths D Bond lengths D

C(l)-C(2) 1.57 1.57 0.00 1.57 ooo

C(l)-C(5) 1.57 1.57 0.00 1.57 0.00

C(l)-F(21) 1.35 1.35 0.00 1.35 0.00

C(l)-F(22) 1.35 1.35 0.00 1.35 0.00

C(2)-C(3) 1.57 1.57 0.00 1.57 0.00

C(2)-F(19) 1.35 1.35 0.00 1.35 0.00

C(2)-F(20) 1.35 1.35 0.00 1.35 0.00

C(3)-C(4) 1.57 1.57 0.00 1.57 0.00

C(3)-F(17) 1.35 1.35 0.00 1.35 0.00

C(3)-F(18) 1.35 1.35 0.00 1.35 0.00

C(4)-C(6) 1.5? 1.58 0 00 1.57 0.00

C(4)-F(15) 1.35 1.35 0.00 1.35 0.00 C(4)-F(16) 1.35 1.35 0.00 1.35 0.00

C(5)-F(23) 1.34 1.34 0.00 1.34 0.00

C(5)-F(24) 1.33 1.33 0.00 1.33 0.00

C(5)-F(25) 1.34 1.34 0.00 1.33 -0.01

C(6)-C(7) 1 .58 1.57 -0.01

C(6)-F(13) 1.35 1.35 0.00 1.35 0.00

C(6)-F(14) 1.35 1.35 0.00 1.35 0.00

C(7)-C(8) 1.56 1.57 0.01 1.57 0.01

C(7)-F(l l) 1.35 1.35 0.00 1.35 0.00

C ^' ( 7 )-!· ^' ( 1 2 ) 1 .35 1.34 -0.01

C(8)-F(9) 1.36 1.33 -0.03 1.34 -0.02

C(8)-F(10) 1.36 1.33 -0.03 1.32 -0.04

C(8)-S(26) 1.91 1.93 0.02 1.93 0.02

S(26)-0(27) 1.48 1.54 0.06 1.56 0.08

S(26)-0(28) 1.48 1.54 0.06 1.56 0.08

S(26)-0(29) 1.48 1.44 -0.04 1.43 -0.05

0(27)-Cu/Mg 1.97 1.95

0(28)-Cu/Mg 1.96 1.95

Table 4. The bond angles of PFOS anions, PFOS-Cu, and PFOS-Mg complexesin water obtained by DFT-B3LYP/6-3 1 \ +G(d,p).

PFOS Bond PFOS PFOS-

Bond angles PFOS PFOS

-Cu angles -Cu Mg

F(24)-

114.4 114.5 108.5 108.6

C(2)-C(l)-C(5) C(5)- 108.86

4 4 8 2

F(25)

C(4)-

108.8 108.7 113.7 113.1

C(2)-C(l)-F(21) C(6)- 113.20

0 5 0 0

C(7)

C(4)-

109.4 109.3 108.6 109.3

C(2)-C(l)-F(22) C(6)- 109.95

3 0 3 1

F( 13)

C(4)-

107.6 107.6 108.0 108.7

C(5)-C(l)-F(21) C(6)- 109.51

6 2 2 4

F(14)

C(7)-

107.5 107.5 108.4 108.2

C(5)-C(l)-F(22) 107.83 C(6)- 107.53

1 8 9 8

F( 13)

C(7)-

F(21)-C(l)- 108.8 108.9 109.2 108.3

109.00 C(6)- 107.54

F(22) 8 2 2 2

F(14)

F( 13)-

113.5 113.4 108.6 109.0

C(l)-C(2)-C(3) 113.41 C(6)- 109.01

6 7 7 2

F(14)

C(6)-

108.0 108.1 114.0 113.1

C(l)-C(2)-F(19) 108.55 C(7)- 112.62

9 7 5 0

C(8)

C(6)-

108.7 108.9 107.6 109.0

C(l)-C(2)-F(20) 109.28 C(7)- 109.73

8 3 1 3

F( 11)

C(6)-

109.1 109.0 108.2 109.4

C(3)-C(2)-F(19) 108.54 C(7)- 110.45

6 7 7 3

F(12)

C(8)-

108.4 108.3 109.5 107.6

C(3)-C(2)-F(20) 108.07 C(7)- 107.10

3 6 7 8

F( 11)

C(8)-

F(19)-C(2)- 108.7 108.7 108.6 108.1

108.91 C(7)- 107.28

F(20) 2 5 3 7

F(12)

F( 11)-

113.4 113.2 108.5 109.3

C(2)-C(3)-C(4) 113.24 C(7)- 109.57

4 5 8 6

F(12)

C(7)-

108.8 109.0 108.8 111.0

C(2)-C(3)-F(17) 109.49 C(8)- 112.06

2 6 5 7

F(9)

C(7)-

108.1 108.3 108.0 111.2

C(2)-C(3)-F(18) 108.85 C(8)- 112.30

6 7 9 6

F(10)

C(7)-

108.4 108.3 115.7 111.0

C(4)-C(3)-F(17) 107.97 C(8)- 111.46

6 6 5 9

S(26) F(9)-

109.0 108.8 107.8 110.2

C(4)-C(3)-F(18) 108.23 C(8)- 110.07

1 4 9 3

F(10)

F(9)-

F(17)-C(3)- 108.8 108.8 107.9 107.2

108.99 C(8)- 104.68

F(18) 6 9 2 1

S(26)

F(10)-

113.4 113.2 108.0 105.7

C(3)-C(4)-C(6) 113.09 C(8)- 105.84

1 6 9 8

S(26)

C(8)-

108.1 108.5 104.2 107.3

C(3)-C(4)-F(15) 109.07 S(26)- 106.21

8 1 1 1

0(27)

C(8)-

108.7 109.0 101.7 104.4

C(3)-C(4)-F(16) 109.66 S(26)- 102.19

1 4 4 3

0(28)

C(8)-

109.1 108.6 103.7 108.7

C(6)-C(4)-F(15) 107.91 S(26)- 111.64

6 7 5 4

0(29)

0(27)-

108.4 108.2 114.6

C(6)-C(4)-F(16) 107.79 S(26)- 99.02 99.67

6 8 3

0(28)

0(27)-

F(15)-C(4)- 108.8 109.0 115.2 118.2

109.25 S(26)- 117.79

F(16) 4 2 2 3

0(29)

0(28)-

110.6 110.6 114.9 117.8

C(l)-C(5)-F(23) 110.28 S(26)- 117.46

7 2 1 0

0(29)

S(26)-

110.9 110.9

C(l)-C(5)-F(24) 110.78 0(27)- 93.95 92.57

7 8

Cu/Mg

S(26)-

108.6 108.6

C(l)-C(5)-F(25) 108.63 0(28)- 94.07 92.44

9 6

Cu/Mg

0(27)-

F(23)-C(5)- 109.1 109.1

109.23 Cu/Mg 72.96 75.25

F(24) 0 3

-0(28) F(23)-C(5)- 108.7 108.7

109.03

F(25) 7 8

Table 5. The dihedral angles of PFOS anions, PFOS-Cu, and PFOS-Mg complexes in water obtained by DFT-B3LYP/6-3 1 \ +G(d,p).

Dihedral angles PFOS PFOS-Cu PFOS-Mg C(5)-C(l)-C(2)-C(3) 163.86 164.45 164.03 C(5)-C( 1 )-C(2)-F( 19) -74.87 -74.40 -75.26 C(5)-C( 1 )-C(2)-F(20) 43.02 43.66 43.40 F(21)-C(l)-C(2)-C(3) -75.71 -75.13 -75.35 F(21)-C(l)-C(2)-F(19) 45.57 46.03 45.36 F(21)-C( 1 )-C(2)-F(20) 163.46 164.08 164.02 F(22)-C(l)-C(2)-C(3) 43.15 43.67 43.23 F(22)-C( 1 )-C(2)-F( 19) 164.42 164.83 163.94 F(22)-C( 1 )-C(2)-F(20) -77.69 -77.12 -77.40 C(2)-C(l)-C(5)-F(23) 50.89 51.53 51.28 C(2)-C( 1 )-C(5)-F(24) -70.39 -69.76 -69.77 C(2)-C( 1 )-C(5)-F(25) 170.27 170.88 170.71 F(21)-C(l)-C(5)-F(23) -70.19 -69.52 -69.67 F(21)-C(l)-C(5)-F(24) 168.54 169.19 169.29 F(21)-C(l)-C(5)-F(25) 49.20 49.82 49.77 F(22)-C( 1 )-C(5)-F(23) 172.66 173.25 172.72 F(22)-C( 1 )-C(5)-F(24) 51.38 51.96 51.67 F(22)-C( 1 )-C(5)-F(25) -67.95 -67.41 -67.85 C(l)-C(2)-C(3)-C(4) 162.35 161.57 162.13 C( 1)-C(2)-C(3)-F( 17) 41.52 40.84 41.60 C(l)-C(2)-C(3)-F( 18) -76.59 -77.56 -77.45 F( 19)-C(2)-C(3)-C(4) 41.67 40.93 41.41 F( 19)-C(2)-C(3)-F( 17) -79.15 -79.81 -79.12 F( 19)-C(2)-C(3)-F( 18) 162.73 161.79 161.83 F(20)-C(2)-C(3)-C(4) -76.62 -77.31 -76.55 F(20)-C(2)-C(3)-F( 17) 162.56 161.95 162.92 F(20)-C(2)-C(3)-F( 18) 44.44 43.55 43.86 C(2)-C(3)-C(4)-C(6) 162.12 162.34 162.68 C(2)-C(3)-C(4)-F( 15) -76.65 -76.91 -77.25 C(2)-C(3)-C(4)-F( 16) 41.39 41.71 42.34 F( 17)-C(3)-C(4)-C(6) -76.86 -76.53 -75.94 F(17)-C(3)-C(4)-F(15) 44.38 44.22 44.14 F( 17)-C(3)-C(4)-F( 16) 162.42 162.84 163.73 F( 18)-C(3)-C(4)-C(6) 41.54 41.74 41.91 F( 18)-C(3)-C(4)-F( 15) 162.77 162.49 161.98 F( 18)-C(3)-C(4)-F( 16) -79.19 -78.89 -78.43 C(3)-C(4)-C(6)-C(7) 161.84 162.04 162.76 C(3)-C(4)-C(6)-F( 13) 40.94 41.31 42.48 C(3)-C(4)-C(6)-F( 14) -76.77 -77.60 -77.27 F( 15)-C(4)-C(6)-C(7) 41.15 41.38 42.02 F( 15)-C(4)-C(6)-F( 13) -79.75 -79.35 -78.26 F( 15)-C(4)-C(6)-F( 14) 162.55 161.74 162.00 F( 16)-C(4)-C(6)-C(7) -77.29 -76.91 -75.85 F( 16)-C(4)-C(6)-F( 13) 161.80 162.37 163.87 F( 16)-C(4)-C(6)-F( 14) 44.10 43.46 44.13 C(4)-C(6)-C(7)-C(8) 162.10 162.74 163.46 C(4)-C(6)-C(7)-F(l 1) -76.13 -77.50 -77.34 C(4)-C(6)-C(7)-F( 12) 41.06 42.08 43.55 F( 13)-C(6)-C(7)-C(8) -76.92 -75.95 -74.89 F( 13)-C(6)-C(7)-F( 11) 44.85 43.81 44.30 F( 13)-C(6)-C(7)-F( 12) 162.04 163.39 165.20 F( 14)-C(6)-C(7)-C(8) 41.38 42.14 42.37 F(14)-C(6)-C(7)-F(l 1) 163.15 161.90 161.56 F( 14)-C(6)-C(7)-F( 12) -79.66 -78.52 -77.55 C(6)-C(7)-C(8)-F(9) 41.18 44.80 45.62 C(6)-C(7)-C(8)-F( 10) -75.75 -78.40 -78.88 C(6)-C(7)-C(8)-S(26) 162.89 164.03 162.56 F(11)-C(7)-C(8)-F(9) -79.50 -75.73 -75.09 F(11)-C(7)-C(8)-F(10) 163.57 161.07 160.41 F(11)-C(7)-C(8)-S(26) 42.21 43.50 41.85 F(12)-C(7)-C(8)-F(9) 162.03 166.18 167.36

F( 12)-C(7)-C(8)-F( 10) 45.10 42.97 42.85 F(12)-C(7)-C(8)-S(26) -76.26 -74.59 -75.71 C(7)-C(8)-S(26)-0(27) -73.13 -84.90 -78.89 C(7)-C(8)-S(26)-0(28) 167.42 170.64 177.11 C(7)-C(8)-S(26)-0(29) 47.83 44.07 50.72 F(9)-C(8)-S(26)-0(27) 49.08 36.62 42.45 F(9)-C(8)-S(26)-0(28) -70.37 -67.84 -61.55 F(9)-C(8)-S(26)-0(29) 170.04 165.59 172.06 F(10)-C(8)-S(26)-O(27) 165.52 154.25 158.75 F(10)-C(8)-S(26)-O(28) 46.07 49.79 54.76 F(10)-C(8)-S(26)-O(29) -73.53 -76.79 -71.64 C(8)-S(26)-0(27)-Cu/Mg -108.90 -107.99 0(28)-S(26)-0(27)-Cu/Mg -0.61 -2.17 0(29)-S(26)-0(27)-Cu/Mg 127.79 126.04 C(8)-S(26)-0(28)-Cu/Mg 111.22 111.23 0(27)-S(26)-0(28)-Cu/Mg 0.612 2.17 0(29)-S(26)-0(28)-Cu/Mg -128.07 -126.26 S(26)-0(27)-Cu/Mg-0(28) 0.49 1.77 S(26)-0(28)-Cu/Mg-0(27) -0.49 -1.76 Table 6. Density Function Theory calculated C-C bond energies of PFOS before and after complex with Cu ²⁺ in aqueous phase.

Bond Energies of Bond Energies of A

Bonds

PFOS/a.u. PFGS-Cu/a.u.

1 0.1306 0.1286 0.002

2 0.1211 0.1188 0.0023

3 0.1192 0.1162 0.003

4 0.1194 0.1153 0.0041

5 0.1209 0.1155 0.0054

6 0.1215 0.1033 0.0182 7 0.1299 0.1200 0.0099

Example 4: Major reaction products of PFOS by ECOHRs.

Materials and Methods

Identification of reaction products.

The concentration of fluoride was determined by ion chromatography (Zhang, et al., Environmental Science & Technology, 47(12): 6471-6477 (2013)) while other potential organic products were identified by HRMS

(Luo, et al., Environmental Science & Technology Letters, 2(7): 198-203 (2015)). To analyze fluoride concentration, an additional 3-mL solution was taken from each reactor and filtered through 0.2 pm acetate cellulose membrane prior to ion chromatography quantification. For other organic reaction products identification, a l0-mL solution sample was taken from each reactor and repetitively extracted with 1 mL dichloromethane for four times at the end of incubation. The extractants were then combined and reconstituted in 40 pL methanol and analyzed using Orbitrap Elite ESI- HRMS (Thermo Scientific, San Jose, U.S.) by the full scan acquisition mode.

The fluoride concentration was quantified using an ion- chromatographer (Dionex, DX500, USA). The system was equipped with an autosampler (sample injection volume: 100 pL), a pump, a degasser, a guard column, and a separation column (Dionex IonPac AS 12 A, 4 mm i.d x 200 mm, USA) operating at 30 °C. The mobile phase was a solution containing 15 mM KOH at a flow rate of 1.0 mL/min. The retention time for fluoride is 4.0 min. An external six-point calibration curve was generated by measuring standard fluoride samples with concentration ranged from 50 to 5000 pg/L.

The high resolution mass spectrometer with full scan (resolution R= 250,000 at m/z 400, for m/z = 10 to 1000) and tandem mass fractionation was performed using an Orbitrap Elite (resolution R = 60 000 at m/z 400, for m/z = 100 to 1000) from Thermo Scientific (San Jose, U.S.) with ESI negative mode.

The full scan spectra of three samples, ECOHRs treatment, negative and positive control, were compared. A peak with m/z value only detected in the ECOHRs treatment sample but not in both of the controls was considered as being possibly associated with a product. The molecular formula of possible products comprising C, H, O, and N was determined using Formula Finder in the Thermo Xcalibur program within a 5-ppm mass error tolerance, and the common rules including carbon-hydrogen ratio, nitrogen rule, and isotopic ratios were employed to exclude unreasonable formula. Targeted MS/MS analysis was then performed on the possible products using the Thermo Orbitrap Elite HRMS, and the molecular structure for each potential reaction product was deduced according to its ion fragmentation pattern.

Differential UV-Vis spectrometry.

The interaction between cations and PFOS was investigated using the differential UV-Vis spectrometry method (Yan, et al., Environmental Science & Technology, 49(14): 8323-8329 (2015)). The same volume (1 mL) of CuS0 ₄ or MgS0 ₄ stock solutions with different concentrations were added to 4 mL of 300 mM PFOS solutions to obtain sequentially increasing Cu ²⁺ or Mg ² concentrations from 2 to 300 pM. The resulting mixtures having cation/PFOS ratio increased from 1:150 to 1: 1. Then, 15 mL of citric buffer was added to each mixture to maintain the pH at 4.9 for Cu ²⁺ and 6.5 for Mg ²⁺ system that was equal to the unbuffered solutions containing 10 mM cations. Additional reference solutions without cation or PFOS were prepared in the same manner. After sample preparation, all mixtures and reference solutions were allowed to equilibrate for 24 hours. Absorbance spectra were recorded using Beckman DU800 spectrophotometer at a wavelength from 200 to 700 nm. A differential absorbance spectrum (DAS) was calculated using the equation:

DAl, DAS = Al, mixture A , cation

where Al, mixture and Al, cation are the absorbance at l wavelength of the mixture solution, the corresponding reference cation solution without PFOS, respectively.

Chemical Analysis.

The PFOS quantitative analysis was done with a Waters AQCUITY I class UPLC system coupled with a XEVO TQD mass spectrometer (Waters, Milford, MA). The separation was carried out by a Waters UPLC BEH C18 column (2.1 x 100 mm, l.7pm, Waters, Milford, MA) at 40 °C using a gradient composition of solvent A (water of 5 mmol/L ammonium acetate) and solvent B (methanol of 5 mmol/L ammonium acetate). The flow rate was 0.3 mL/min and the gradient program lasted for 10 min: 0-0.5 min, hold at 10% B; 0.5-8 min, linearly increase B from 10% to 95%; 8-8.1 min, a linear increase from 95% to 100%; 8.1-9 min, linearly reduced B to 10%, and then equilibrium at 10% B for 1 min. Electrospray ionization was operated in a negative mode with the parameters set as capillary voltage at -1.0 kV, desolvation temperature at 400 °C, source block temperature at 150 °C. Nitrogen (> 99.999% purity, Airgas) was used as desolvation gas with the flow rate of 550 L/hour. In addition to PFOS, M8PFOS and HBT, additional perfluoroalkane sulfonates (PFSAs) and perfluoroalkylic acids (PFCAs) with total carbon-chain length ranging from C4 to Cl 1 were monitored simultaneously using multiple reaction monitoring. The precursors and transitions m/z values of all the monitored PFSAs, PFCAs and HBT as well as their detection limits were listed in Table 1.

Results

Formation of fluoride is a key indicator of PFOS degradation. The concentrations of fluoride in the 1-20 treatment samples shown in Figures 1A-1B were monitored using ion chromatography. As shown in Figure 5, the F concentration increased with time at a rate of 0.617 mhioΐ F /d in the Cu ²⁺ solution with the final concentration at 5.41 pmol/L at the end of incubation; while in the Mg ²⁺ solution in which PFOS degradation was slower, with the final concentration of fluoride determined at 3.18 pmol/L. The defluorination ratio (R) was calculated from the equation X 100% where the

CF- is the concentration of released fluoride, the Co and C _t are concentrations of PFOS at time zero and t, respectively, and the 17 is a factor of the number of fluorine atoms contained in PFOS molecule (Lin, et al., Water Research, 46(7), 2281-2289 (2012)). According to this equation, the defluorination ratios were 47.4 and 47.1% for the 1-20 treatments in the Cu ²⁺ and Mg ²⁺ solution, respectively, at the end of 162-day treatment.

To identify possible products, the 1-20 treatment samples were extracted and analyzed at the end of incubation. By comparing the high resolution mass spectra of the 1-20 treatment samples with corresponding positive (PFOS only) and negative (no PFOS but with laccase and HBT) controls, the m/z peaks corresponding to the molecular ions only present in the 1-20 treatment samples can be identified. The element compositions of the possible PFOS degradation products were determined according to their accurate molecular weights given by HRMS (mass accuracy < 5 ppm). Possible structures of the PFOS degradation products were then deduced from their fragment ion patterns (Table 7). For those products containing the ³² S element such as product No. 1, 2, 13, 14 and 17, their corresponding ³⁴ S isotope intensities were also used to verify the product structures.

The formation of shorter-chain perfluorocarboxylic acids (PFCAs) from PFOS degradation by sonochemical (Moriwaki, et al., Environmental Science & Technology, 39(9): 3388-3392 (2005)), photodecomposition (Jin, et al., Journal of Hazardous Materials, 271: 9-15 (2014)), electrochemistry (Carter, et al., Environmental Science & Technology, 42(16): 6111-6115 (2008)) technology was reported in earlier studies, while none of them was identified as a reaction product in this study. It was found that all of the tentative products were partially fluorinated compounds (Table 7), which are similar to those found in studies of PFOA degradation during ECOHRs. Product No. 5 and 16 were detected in both Cu ²⁺ and Mg ²⁺ systems. As shown in Table 7, the degradation products could be categorized into three different groups: 1) the ones with shorter-carbon chain perfluoroalkyl ends 5 like product 3, 7, 11, 12, and 16; 2) the ones with carboxylic/sulfonic acid end such as product 2, 5, 8, 9, 13, 14, and 17; and 3) the ones containing multiple non-fluoro moieties with perfluoro-moieties (product 1, 4, 6, 10, and 15). It is worth noting that the molecular structures of products 10, 13, and 15 contained HBT moieties. This is the direct evidence confirming the 10 reaction between BTNO and shorter-chain fluoroalkyl radicals.

Table 7. Molecular formulas, theoretical and measured deprotonated molecule weight, mass error (ppm) and possible structures of PFOS degradation products from ECOHRs.

[M-H]- Mass

A previously proposed PFOS decomposition pathway usually starts with the dissociation of the sulfonic functional group, followed by stepwise unzipping of CF2 units from the perfluoroalkyl carbon backbone (Vecitis, et ak, Environ. Sci. Eng, 3(2): 129-151 (2009); Moriwaki, et al., Environmental Science & Technology, 39(9): 3388-3392 (2005)). Such a pathway would generate shorter-carbon chain PFCAs as degradation products (Niu, et ak, Environmental Science & Technology, 47(24): 14341-14349 (2013)).

However, the reaction products identified in this study indicate that other reaction mechanism may be the major route for PFOS degradation during

ECOHRs. The DAS results and DFT modeling showed that metal ion-PFOS complexation could unlock the helical configuration, and decrease the C-C bond energy of PFOS. These changes allow the attack of PFOS C-C backbone by BTNO becomes easier. In addition, complexation to metals neutralized the negatively charged PFOS, brought it and negatively charged laccase to proximity. Thus, BTNO had a greater chance to attack PFOS once it was generated and released from the enzyme catalytic center (Lu, et a , Environmental Pollution, 220: 1418-1423 (2017)).

Based on the products identification and DFT computation information, a free radical chain reaction and cross-coupling mechanism could be a potential reaction mechanism for PFOS degradation during ECOHR. First, HBT is activated by laccase to form free radicals (BTNO). The free radicals could directly attack the C-C bonds in PFOS, leading to the formation of a shorter-chain length perfluoroalkyl free radical, with or without the sulfonic functionality. In the meantime, other non-fluorinated

organic compounds present in the reactor can also be converted to radicals

by BTNO during free radical propagation process (Equation 2-3, below).

The perfluoroalkyl free radicals and the non-fluorinated radicals could

couple to produce the degradation products (Equations 4-7, below). Such

free radical chain reaction and coupling process may happen repetitively,

leading to products with multiple non-fluorinated moieties and fluorinated

moieties embedded with each other (Equation 8, below).

The reaction mechanism of PFOS degradation during ECOHRs via

direct dissociation at the C-C bond.

Initiation

Laccase

Mediator (M) - _► Mediator radcial (M ^' ) ©

Propagation

M ^' + other organic compounds (R) R ©

Termination

CT2 ₁₁₊ 1 + R CF ₃ R (Product 3, 7, 11, 12, 16) Q c _m F ₂ mSO: ⁺ R ^RC m ^F 2m ^S °3 (Product 13, 14) © C _m F _2m C0 ₂ ^_ + R - ^► RC _m F _2m c5 ₂ (Product 2, 5, 8, 9) ( ό

C ₈ F _{1 7} S0 ₃ ^" + R ^{' RearrangCme} '» C ₂ F ₅ CFCFC ₃ F ₆ CHFS0 ₃ ~ (Product 17) ©

Multiple Reaction

Rearrangement

C ₂ F ₅ R + M ^' - C ₂ F ₅ R + R’· - ► RC ₂ F ₄ R’ (Product 1 , 4, 6, 10, 15)

This study demonstrates PFOS degradation in the laccase-HBT

reaction system, showing the effectiveness of ECOHRs on the degradation of perfluroalkylsulfonates at environmentally relevant conditions. ECOHRs

induce PFOS degradation via a radical chain reaction by directly attacking

the C-C bond of PFOS and generating the perfluoroalkyl or acid radical followed by formation of partially fluorinated products via radical rearrangement and cross-coupling. Products formed during ECOHRs having less fluorine and more hydrogen atoms are expected to be less toxic and more available for microbial degradation (Lau, et al., Toxicological Sciences, 99(2): 366-394 (2007)). The ECOHR mechanism may be effective in natural water and soil systems to transform and incorporate PFOS into the natural organic matter, thus detoxifying and immobilizing PFOS. Moreover, it is possible that ECOHRs may be engineered by amending catalysts and mediators to serve as a remediation technique to degrade PFOS and other perfluorinated chemicals in the environment.

Example 5: Natural compositions can mediate PFOA degradation

Materials and Methods

Chemicals and reagents

Perfluorooctanoic acid (PFOA), laccase (EC 1.10.3.2), 2,2'-azino- bis(3-ethylbenzothiazoline-6-sulphonic acid) (ABTS), 1- hydroxybenzotriazole (HBT), and 2,6-dimethoxyphenol (DMP) were purchased from Sigma- Aldrich (St. Louis, MO). HPLC-grade acetone, acetonitrile (ACN), methanol, and dichloromethane (DCM), were obtained from Fisher Scientific (Pittsburgh, PA). The perfluorocarboxylic acids (PFCAs) with the total carbon number from C4 to Cl 1 and the surrogate standard perfluoro-n-[ ¹³ C8]-octanoic acid (M8PFOA) were acquired from Wellington Laboratories (Ontario, Canada). Natural organic materials including SBM, HM, and MC were obtained from Eureka Springs Organics (Eureka Springs, AR) and used as model natural organic mediators. These organic materials are usually by-products of agricultural industry. They are widely available at low prices and have already been utilized as soil amendment.

Two types of laccase were evaluated. A purified laccase (activity ~ 220 U/mg) from pleurotus ostreatus (PO) was purchased from Sigma- Aldrich (St. Louis, MO). A crude laccase solution (activity of ~ 873 U/mL) concentrated from the fermentation broth of the fungus Pycnoporus sp. SYBC-L3 (PS) was produced using a published procedure (Liu, et ak, Journal of Cleaner Production, 39: 154-160 (2013)). Earlier studies have demonstrated the effectiveness of this enzyme in decontamination and biofuel production (Liu, et al., Bulletin of Environmental Contamination and Toxicology, 89: 269-273 (2012); Liu, et al., Bioresource Technology, 135: 39-45 (2013)). The PS enzyme exhibited great stability in the crude form of concentrated solution, retaining over 90% activity after one year of shelf storage at room temperature (Johannes, et al., Applied and Environmental Microbiology, 66: 524-528 (2000)).

ECOHR in water

The experiment was carried out in the water extract of individual natural organic material spiking with 1.0 mM of PLOA. The extract solutions of natural organic materials were prepared by equilibrating 20 g of the natural organic materials, either SBM, HM, or MC, with 600 mL HPLC water for five days and then filtered with 0.45-pm cellulose acetate membrane. Each treatment reactor contained 10 mL of an extract solution with 1.0-mM PLOA, and was dosed with 1 U/mL laccase, and 0 or 20 mM HBT repeatedly every six days, respectively. The laccase stock solutions (167 U/mL) were prepared in HPLC water and the HBT (0 or 5 mM) was dissolved in acetonitrile. The positive control reactors were prepared and processed in the same time and manner, except that they were dosed with the same amounts of HPLC water (without laccase) and acetonitrile (without HBT). All reactors were gently agitated at 22 °C in an incubator (Innova 42, New Brunswick Scientific). Every 12 days, a set of five replicate treatment reactors and five positive controls were sampled, and each was added with an appropriate volume of HPLC water to bring the final volume to 11 mL. A 0.5-mL portion was then withdrawn from each reactor and subjected to solid phase extraction (Luo, et al., Environmental Science & Technology Letters, 2 198-203 (2015)). Additional negative soil slurry controls were also prepared without spiking PLOA but with laccase and SBM. The details are included in the Methods and Materials section of Example 6. Statistic analysis

A general linear model ANOVA with significant differences (a = 0.05) test was conducted on SAS followed by a least significant difference test to compare the PFOA concentration of the same treatment over different sampling times.

Results

Perfluorooctanoic acid (PFOA) is one of the most predominant PFAAs in the environment, having been frequently detected at levels higher than parts-per- trillion in soil and water (Xiao, et al., Water Research, 72: 64-74 (2015)). A recent global survey reported PFOA concentrations ranging from 0.14 to 2.67 ng/g in background surface soil with limited anthropogenic impact (Rankin, et ak, Chemosphere, 161:333-341 (2016)). It was reported that the soil at sites with historical fire-fighting training activities was polluted by high concentrations of PFOA (0.23 to 12.2 pg/kg) (Karrman, et ak,

Environmental Chemistry, 8: 372-380 (2011)), and those in the soil at a former industrial area reached 47.5 pg/kg (McGuire, et ak, Environmental Science & Technology, 48:6644-6652 (2014)). McGuire and etc. (McGuire, et ak, Environmental Science & Technology, 48:6644-6652 (2014)) reported the PFOA concentration in a former firefighter training area ranging from 10 pg/kg to 10,500 pg/kg.

Various techniques that have been developed to degrade PFAAs in aqueous phase are often not suitable for soil remediation, while the studies focusing on PFAA degradation in soil are rare. Biodegradation of PFOA under anaerobic condition with microorganism was found ineffective (Liou, et ak, Chemosphere, 80:176-183 (2010); Washington, et al., Environmental Science & Technology, 49: 915-923 (2015)). Persulfate oxidation has been shown able to break down PFOA, but not PFOS, and the process requires heating, extreme pH, UV or chemical activation (Park, et ak, Chemosphere, 145:376-383 (2016)). For example, an experiment using persulfate to degrade PFOA in a soil slurry involved repeated injections of persulfate and activation agents to achieve desirable removal (Park, et ak, Chemosphere, 145:376-383 (2016)). A series of experiments were conducted to examine PFOA degradation by ECOHRs in water and in soil slurry in this study. Two laccase species at different purities were tested at various dosages and three natural organic materials, soybean meal (SBM), humate (HM) and mushroom compost (MC), were examined as mediators. High-resolution mass spectrometry was used to identify the products resulting from PFOA degradation to explore the reaction mechanisms. The change of laccase activity in the soil was also monitored to assess the stability of laccase in soil.

Figures 6A-6C depicts the change of PFOA concentration over time in the water extract of three natural organic materials with or without the presence of laccase and HBT. The statistic analysis results were summarized in Figures 6D-6F. PFOA concentration remained stable over the incubation period in the control samples with the absence of laccase or HBT in all three extract solutions. Continuous decrease of PFOA concentration was observed in both SBM and HM solutions with ECOHR treatment, but was not evident in MC solution during the 36-d incubation period (Figure 6A-6C). In the presence of HBT and PO laccase (PO HBT treatment), PFOA concentration was decreased by approximately 29 and 24% in SBM and HM solutions, respectively. It is worth noting that PFOA concentration also decreased in SBM (23%) and HM (14%) system even in the absence of HBT, indicating that these extract solutions might contain chemicals that could serve as natural mediators for ECOHR.

Previous study has identified several natural laccase-mediators in soybean meal water extract solution (Liang, et al., Chemosphere, 181: 320- 327 (2017)). The total organic carbon (TOC) and different metal ion concentrations in these extract solutions were analyzed (Table 8). The results showed that SBM extract had the highest TOC of 1865 mg/L followed by MC extract of 269.9 mg/L, the HM extract contained 24.8 mg/L TOC. For PS laccase (Figure 6A and 6B), decrease of PFOA concentration was observed in both SBM and HM extract in the presence of HBT, while PFOA concentration remained consistent in PS only treatment during the entire incubation period. Similarly, PO HBT treatment also led to significant

degradation of PFOA (Figures 6D and 6E). The PO HBT treatment reaction

rates (k = -0.0083/day, r = 0.99 in SBM; k = -0.0058/day, r = 0.89 in HM) are generally higher than PS HBT treatment (k = -0.0057/day, r = 0.92 in SBM;

k = -0.004l/day, r = 0.91 in HM), while PFOA concentration in the treatment without HBT tended to reach a plateau after 12 days. One possible reason

could be the exhaustion of natural organic mediators in the SBM and HM

solutions. In addition to organic matters, multivalent metal ions were also

detected in all three extract solutions at different levels, in particular, SBM

extract consisting of appreciable amounts of multivalent metal ions such as

Cu ²⁺ (Table 9), which had been shown to play a significant role in PFOA

degradation by ECOHRs (Luo, et al., Environmental Science & Technology

Letters, 2 198-203 (2015)).

Table 8. Total organic carbon (mg/L), pH values, and different metal ion

concentrations (mg/L) in extract solutions of soybean meal, humate, and

mushroom compost.

Total organic

Extract Metal ion concentration (mg/L)

carbon pH

Solution

(mg/L) Na Mg Al Ca Fe Cu

Soybean <

1865.0 5.1 78.33 < 0.08 78.67 < 0.6 0.0411 meal 1.00

<

Humate 24.8 3.5 6.92 4.20 < 0.08 34.14 < 0.6

0.008

Mushroom

269.9 7.9 36.18 13.49 1.83 9.49 1.14 1.4704 compost

Table 9. Molecular formulas, theoretical and measured deprotonated

molecule weight, mass error (ppm) and possible structures of PFOA

degradation products from ECOHRs.

[M-H] Mass

No Formula error Possible structure Treatment

Theo. Exp. (ppm)

1 CHNF 147.0488 147.0483 -3.2

2 C6H70N2F3 179.0438 179.0445 4.1 o PS 60

CF ₃ (CH=CH) ₂ NHCNH ₂ PS 20/4wk ^b

10 CH ONF 339.0421 339.0410 -3.3 PS 60

11 C HONF 371.0284 371.0289 1.5 PS 60

12 CHOF 384.9740 384.9741 0.4 o

HC(CF ₂ ) ₅ CF=CFC0 ₂ H

a PO 60: 60 U of PO was added to the reactor at the beginning of the

experiment only.

b PS 20/4 wk: 20 U of PS was added every 4 weeks. Example 6: ECOHRs can facilitate PFOA degradation in soil slurry.

Materials and Methods

ECOHR in soil slurry

The soil used in this study was collected from Dempsey Research Farm, Griffin, GA, USA that was classified as sandy loam in texture. The physical-chemical properties of the soil were analyzed and summarized in Table 10. The soil was air-dried and processed through a sieve prior to sterilization by autoclaving three times, 60 minutes each, at 121 °C within a three-day period. The soil was incubated at room temperature for 24 hours after the first two sterilization events. To prepare PFOA spiked soil samples, 100 g autoclaved soil was spiked with 1 mL of 50 mg/L PFOA in acetone, and then left uncovered under a fume hood with vigorous mixing to evaporate the solvent, thus yielding a soil sample containing 0.5 pg/g PFOA.

The ECOHR experiment was carried out in 20-mL high-density polypropylene vials. Each vial contained l.O-g PFOA-spiked air-dried soil, 50 mg soybean meal, and 1.5 mL HPLC water. Different enzyme dosing strategies were evaluated, including one-time dosing of laccase at a high dosage (60 units) and sequenced multiple dosing of laccase at a low dosage (20 units once every 4 weeks). The control samples were processed along with the treatments with equal amount of water added without laccase. Each vial was thoroughly mixed using a mechanical shaker at the beginning of the incubation and then mixed manually twice a day during the entire incubation period. Five replicates of the treatment vials were scarified at selected time intervals along with five control vials for PFOA analysis.

Table 10. Physical-chemical properties of the soil used in the study

Parameter Value Parameter Value

Sand content (%) 66.00 Cation exchange capacity 3.70

(meq/lOO g)

Silt content (%) 23.60 Organic matter content (%) 2.90

Clay content (%) 10.40 Soluble salts (mmhos/cm) 0.04 pH 5.15 Base saturation (%) 65.9

Extraction of PFOAfrom soil slurry

Each vial was first spiked with 0.125 pg of M8PFOA as surrogate standard, frozen at -18 °C, and then freeze dried (FreeZone freeze dryer, Labconco). A 0.5-g portion of the freeze-dried sample was mixed with a 5- mL mixture solvent (dichlorome thane : methanol = 2 : 1, v : v),

ultrasonicated for 30 minutes (Cole Parmer, Ultrasonic Processor), and then centrifuged at 200 g for 10 min before collecting the supernatant. The same extraction procedure was repeated twice, with all supernatants combined and blown to dryness under a gentle nitrogen flow and reconstituted in 2 mL methanol. The extractant was then passed through a 0.45-pm cellulose acetate membrane preconditioned by 2-mL methanol (VWR International, Radnor, PA). The filtrates were then analyzed by HPLC-MS/MS. The PFOA recovery of this extraction method was found to be 79 ± 2.4% by tests of seven samples spiked at 0.5 pg/g.

Results

The soil slurry experiment was conducted with SBM supplemented to the PFOA-spiked soil as a natural mediator material, because the experiment with the water extracts had demonstrated the effectiveness of SBM in PFOA degradation (Figure 6A-6C). The change of PFOA concentrations in the soil slurry during 20 weeks (140 days) of incubation is shown in Figure 7A-7B. Two types of enzymes, i.e. PO and PS were tested, and each was dosed using two different addition modes: I) repetitive addition of 20 U enzyme/g soil every four weeks during incubation (named as PO 20/4 wk and PS 20/4 wk), and II) one single addition of 60 U enzyme/g soil at the beginning of incubation (named as PO 60 and PS 60). Controls without the addition of enzyme were also incubated and processed along with the treatment reactors. It can be seen from Figure 7A that the addition of laccase, regardless of the dosing modes and enzyme species, greatly facilitated PFOA degradation in the soil slurry. The statistic analysis results were summarized in Figure 7B. PFOA removal reached 29, 35, 35, and 40% for PO 20/4 wk, PS 20/4 wk, PO 60, and PS 60 respectively after 20 weeks of reaction, whereas no decrease of PFOA concentrations was detected in the controls at the end of incubation.

PFOA removal in the soil slurry followed pseudo-first order kinetics, and the apparent reaction rate constants (k) were O.OOlO/day (r = 0.79) for PO 20/4wk, 0.00l3/day (r = 0.89) for PS 20/4wk, 0.00l2/day (r = 0.69) for PO 60, and 0.00l7/day (r = 0.94) for PS 60. It is worth noting that the PO 60 treatment exhibited significant PFOA removal within the initial four weeks, and then slowed down; while the PS 60 treatment induced more steady PFOA degradation throughout the 20 weeks of incubation. This was probably due to the different behaviors of the two enzymes in soil slurry. Example 7: Laccase activity in soil slurry.

Materials and Methods

An additional experiment was carried out to monitor the change of laccase activity in soil slurry. Each reactor contains a mixture of 1.0 g of autoclaved soil and 1.5 mL of laccase solution and was incubated at room temperature for 84 days. Triplicate of reactors were sacrificed at

predetermined time intervals, and centrifuged at 8288 g to separate the liquid and solid phases. Enzyme activity in liquid phase and solid phase were determined separately. The liquid phase enzyme activity which includes the enzyme in the liquid supernatant and enzyme loosely adsorbed on soil surface were determined using DMP as substrate as described previously (Luo, et ah, Environmental Science & Technology Letters, 2 198-203 (2015)). The solid phase enzyme activity was quantified by the rate of absorbance change of ABTS at 420 nm, as described in more detail below. Solid Phase Extraction.

Each 0.5-mL aqueous sample was first spiked with 0.5 mL of 0.5 mM M8PFOA in HPLC water as the surrogate standard. The solid phase extraction (SPE) cartridge (Waters, Oasis HLB) was then preconditioned with 3 mL methanol, 3 mL of HPLC water, and another 3 mL HPLC water sequentially, followed by loading of 0.9-mL sample. The cartridge was then washed with 3 mL HPLC water and then blown to dry for 10 min under vacuum. The cartridge was then eluted with 1 mL methanol for three times and then 1 mL acetonitrile for two times to elute perfluorocarboxylic acids (PFCAs). All eluents were combined, and the mixture was blown to 1 mL with nitrogen gas for PFCAs quantification. This extraction procedure was developed based on published methods with minor changes (So, et al., Environmental Science & Technology, 38, 4056-4063 (2004), Yu, et al., Water Research, 43:2399-2408 (2009)).

Assessment of enzyme activity.

Each reactor contained 1.0 g autoclaved soil, and 0.2 mL laccase solution, and 1.3 mL HPLC water to bring the total volume of water to 1.5 mL. Samples were collected and measured for enzyme activities at predetermined time intervals by sacrificing the reactors. At each sampling event, liquid and solid phases were separated by centrifuging at 8228 g for 5 minutes. The supernatant was then removed with pipette to determine the laccase activity in the liquid phase as described previously (Park, et al., Environmental Science & Technology, 33:2028-2034 (1999)). Briefly, laccase activity was measured via the oxidation of 1 mM of DMP in a citrate phosphate buffer (pH 3.8) with the absorbance change was measured at 468 nm. One unit is defined as the amount of enzyme that causes a unit absorbance change per minute in 3.4 ml of this solution in cuvette with 1 cm light path. The solid phase was then rinsed three times with HPLC water.

The enzyme activity was measured each time until there is no activity was detected in the rinse water. The combined activity in the supernatant and rinse water were equal to the enzyme activity remain in the liquid phase. In order to measure laccase activity in the solid phase, 3 ml of 100 mM ABTS in phosphate buffer (pH 6.0) was added directly to the solid phase sample after the rinses in each tube, which was then vortexed and centrifuged to allow for 10 minutes of color development. Laccase activity was then quantified by the rate of absorbance change at 420 nm (Canas, et al., Environmental Science & Technology, 41 (2007) 2964-2971 (2007)). The laccase activity measured in the ABTS unit was then converted to the DMP unit by a calibration curve prepared by measuring known amounts of laccase using both DMP and ABTS methods. The equation below showed the relationship between the laccase activities measured by these two methods: Activity ( DMP unit) = 2.65 x Activity (ABTS unit).

Results

Laccase is one of the key factors in ECOHRs, therefore

understanding the stability of different laccase species in soil is important in optimizing PFOA degradation efficiency. An additional laccase incubation experiment was conducted to investigate the behavior of PO and PS laccases in soil slurry, and the data are displayed in Figures 8A-8B. The apparent total enzyme activity at time zero was 0.48 U for PO and 0.83 U for PS in soil slurry, although a total of 1.71 U of PO or 1.78 U of PS was added. This is probably because that a portion of laccase was adsorbed to soil, and it is known that the apparent activity of enzyme usually declined in adsorption status (Mateo, et ak, Enzyme and Microbial Technology, 40: 1451-1463 (2007)). As shown in Figure 8A, the liquid phase activity accounted for the majority of the total apparent activity in soil slurry for PS laccase, while, for PO laccase, the contribution of liquid and solid phases tended to be equal (Figure 8A). PO activity became undetectable after about two months of incubation in the soil slurry, while after 84 days the total activity of PS still remained 0.18 U (Figure 8B). The relative less inactivation of PS than PO in soil slurry may explain the different PFOA degradation patterns shown in Figure 7A-7B. Example 8: PFOA degradation products.

Materials and Methods

Chemical analysis.

Seven short-chain PFCAs (Table 11), PFOA, M8PFOA, and HBT were quantified using a Waters AQCUITY I-Class UPLC system coupled with a XEVO TQD mass spectrometer (Waters, Milford, MA) as reported in a previous study (Luo, et al., Environmental Science & Technology Letters, 2 198-203 (2015)). The separation was carried out on a Waters UPLC BEH C18 column (2.1 x 100 mm, l.7pm, Waters, Milford, MA) at 40 °C using a gradient composition of 5 mmol/L ammonium acetate in HPLC water (solvent A) and 5 mmol/L ammonium acetate in methanol (solvent B). Targeted compounds were monitored simultaneously using multiple reaction monitoring, with the precursor and transition m/z values of all these compounds listed in Table 11 along with detection limits. Detail HPLC gradient program and MS operation parameters are described in more detail below.

Quantitative analysis of PFCAs, M8PFOA, and HBT was performed with a Waters ACQUITY I-Class UPLC system coupled with the XEVO TQD mass spectrometer (Waters, Milford, MA). The separation was carried out on a Waters UPLC BEH C18 column (2.1 x 100 mm, 1.7 pm, Waters, Milford, MA) at 40 °C using a gradient composition of solvent A (5 mmol/L ammonium acetate in HPLC water) and solvent B (5 mmol/L ammonium acetate in methanol). The flow rate was 0.3 mL/min with a gradient program lasting for 10 min: 0-0.5 min, hold at 10% B; 0.5-8 min, linearly increased B from 10% to 95%; 8-8.1 min, a linear increase from 95% to 100%; 8.1-9 min, linearly decreased B to 10%, and then equilibrium at 10% B for 1 min. Electrospray ionization was operated in a negative mode with the parameters set as the capillary voltage at -1.0 kV, desolvation temperature at 400 °C, and source block temperature at 150 °C. Nitrogen (> 99.999% purity, Airgas) was used as desolvation gas with a flow rate of 550 L/hour. Table 11. The molecular formula, retention time, precursor ion, transition ion, and detection limit of the PFCAs monitored in UPLC-MS/MS analysis.

Precursor Transition Detection

Molecular RT

Chemicals ion ion limit

Formula (min)

m/z m/z (pg/L)

F-

PFBA (C4) 2.6 213 169 0.018

(CF ₂ ) ₃ COOH

PFPeA F-

4.5 263 219 0.027

(C5) (CF ₂ ) ₄ COOH

PFHxA F-

5.5 313 269 0.022

(C6) (CF ₂ ) ₅ COOH

PFHpA F-

6.1 363 319 0.008

(C7) (CF ₂ ) ₆ COOH

F-

PFOA (C8) 6.6 413 369 0.011

(CF ₂ ) ₇ COOH

F-

PFNA (C9) 7.02 463 419 0.15

(CF ₂ ) ₈ COOH

PFDA F-

7.32 513 469 0.10

(C10) (CF ₂ ) ₉ C00H

PFUA F-

7.63 563 519 0.012

(Cl l) (CF ₂ ) _IO COOH

PFBA: perfluoro-n-butanoic acid; PFPeA: perfluoro-n-pentanoic acid;

PFHxA: perfluoro-n-hexanoic acid; PFHpA: perfluoro-n-heptanoic acid; PFNA: perfluoro-n-nonanoic acid; PFDA: perfluoro-n-decanoic acid; PFUA: perfluoro-n-undecanoic acid.

Identification of degradation products.

In order to identify possible products resulting from the ECOHR treatment of PFOA in soil, the extractants obtained through the above extraction and filtration procedure from selected soil slurry treatment samples and correspondent negative and positive controls were analyzed by high-resolution mass spectrometry (HRMS). The negative control was the soil slurry without PFOA spiked but with laccase and SBM, and the positive control was the soil slurry spiked with PFOA but without laccase or SBM. All samples were incubated together under the same conditions. HRMS was performed on an Orbitrap Elite high-resolution mass spectrophotometer (Thermo Scientific, San Jose, U.S.) with full scan from m/z = 100 to 1000 (resolution R= 250,000 at m/z 400) at ESI negative mode (Luo, et ak, Environmental Science & Technology Letters, 2 198-203 (2015)), and selected ions were further subjected to tandem mass fractionation (MS/MS) (resolution R = 60,000 at m/z 400).

Results

In order to screen possible PFOA degradation products in soil slurry, the samples from PS 60 and PS 20/4wk treatments were subjected to HRMS analysis at the end of 140 days of incubation, along with correspondent positive and negative control samples. The m/z peaks that were only detected in the reaction samples but not in the negative or positive controls were assigned as possible products. The elemental composition of each m/z was determined by Formula Finder in the Thermo Xcalibur with a limit of 5 ppm mass error. A total of 12 tentative products were found in this manner and are listed in Table 9. Three out of the 12 products (product No. 2, 3, and 6) were found in both PS 60 and PS 20/4 wk treatments. The molecular structures of the tentative products were further elucidated from the fragment ion spectra obtained from high-resolution tandem mass spectrometry.

The products do not include any short-chain length PFCAs, as those found in the other oxidative treatments such as photolysis and electrolysis (Jin, et ak, Journal of Hazardous Materials, 271: 9-15 (2014); Lin, et ak, Water Research, 46: 2281-2289 (2012)), but are primarily partially fluorinated compounds containing perfluoroalkyl moieties, similar to those found in studies about ECOHRs in aqueous phase (Luo, et ak,

Environmental Science & Technology Letters, 2 198-203 (2015); Luo, et ak, Environmental Pollution, 224:649-657 (2017)). These products in soil could be categorized into three types I) Product 2, 4, 8, 9 are formed by combining the perfluoro-end of PFOA with a non-fluoro moiety, II) Product 3, 5, 6, 12 seem to form from coupling the acid-end of PFOA with a non-fluoro moiety, and III) Product 1, 7, 10, 11 that contain either multiple non-fluoro moieties (1, 7) or multiple perfluoro moieties (10, 11).

Example 9: PFOA degradation mechanisms.

The types of the products identified in the system indicate a free radical chain reaction and cross-coupling mechanism schematically shown below. First, natural mediators (M) are activated by laccase to form free radicals (M·). The free radicals may attack the C-C bonds in PFOA, leading to the formation of a short-chain length perfluoroalkyl free radical, with or without the carboxylic functionality. Some other non-fluorinated organic compounds present in the reactor can also be converted to radicals by the activated natural mediators during free radical propagation process. The perfluoroalkyl free radicals and the non-fluorinated radicals could couple to produce the degradation products (Equations 4-7). Free radical

rearrangements may occur during the cross-coupling process to form structures like -CF=CF- in the products (Equation 8). Such free radical chain reaction and coupling process may happen repetitively, leading to products with multiple fluorinated and non-fluorinated moieties embedded with each other (Equation 9).

The non-fluorinated compounds involved in the reaction are the impurities present in the reactor, such as free amino acids or fermentation by-product acids and alcohols in laccase solution and SBM. The HRMS analysis results confirmed the presence of the non-fluorinated compounds. For example, the accurate molecular weights of amino acids such as histidine, tyrosine, and aspartic acid, which were involved in the formation of product 3, 8, and 10 respectively, were found in the sample extractants of the negative control. Without quantification of the amount of each product, due to the lack of standards, it is unable to determine which reaction pathway is predominant during PFOA degradation. However, 3 out of 12 products contained a C2F4CO2 fraction (i.e. product No. 3, 5, 6).

The results verified the effectiveness of laccase-SBM induced ECOHR in the degradation of PFOA in soil. SBM, as a cheap and eco- friendly natural organic material, could serve as an ideal natural organic

mediator due to its composition of high organic matter and various

multivalent metal ions. The purified PO laccase and crude PS laccase

demonstrated similar efficiency in PFOA degradation, while using crude PS laccase could significantly reduce the cost of ECOHRs application in soil

remediation. Based on the structures of the degradation products, PFOA

degradation during ECOHRs is proposed to be initialed by activated

mediators direct attack of the C-C bonds in PFOA with subsequent

fragmentation, rearrangement and cross-coupling.

The reaction mechanism of ECOHR induced PFOA degradation via direct dissociation at the C-C bond.

Initiation

Laccase

Mediator (M) Mediator radcial (M ^* ) 1

Propagation

M ^" + other organic compounds (R) - ► R

M ^’ or R ' + C ₇ F ₁₅ C0 ₂ C _n nF ^r 2n+l + C _m F _2m CO _;

Termination

CF ₃ ^' + R ^' — ^► CF ₃ R (Product 2) Q C ₂ F ₅ ^‘ + R ^' —► C ₂ F ₅ R (Product 8) © C ₃ F ₇ ^’ + R ^' —► C ₃ F ₇ R (Product 4) © C ₂ F ₄ C0 ₂ + R - ^► RC ₂ F ₄ C0 ₂ (Product 3, 5, 6) © C ₇ F ₁₅ C0 ₂ + R Rearrangement _{(Prodl]ct 12)} 8

Multiple Reactions

Rearrangement

C ₂ F ₅ R + M ^' C ₂ F ₅ R + R'· RC ₂ F ₄ R’ (Product 7, 1 1 ) Example 10: Coupling granular activated carbon with laccase catalyzed humification reaction for separation and destruction of PFAS

Materials and Methods

Chemicals and water samples

The perfluorooctanoic acid (PFOA) and perfluorooctanesulfonic acid (PFOS) were purchased from Sigma Aldrich (St. Louis, MO). The surrogate standard perfluoro-n-[ ¹³ C8]-octanoic acid (M8PFOA) and perfluoro-n- [ ¹³ C8]-octanesulfonic acid (M8PFOS) were obtained from Wellington Laboratories (Ontario, Canada). Sand, l-hydroxybenzotriazole (HBT), and 2,6-dimethoxyphenol (DMP) were purchased from Sigma Aldrich (St. Louis, MO). HPLC grade acetonitrile, methanol, and dichloromethane were from Fisher Scientific (Waltham, MA). All solutions were prepared in nanopure water generated from Bamstead NANOpure® water purification system (Thermo Scientific, Waltham, MA). The 0.2pm Cellulose Acetate Membrane and 4-mL Reservoir were purchased from VWR International (Radnor, PA). Laccase was produced from fermentation with fungal strain Pycnoporus sp SYBC-L3 (PS) in the form of a crude concentrate (-900 U/mL) of the fermentation broth (Liu, et al., Bioresource Technology, 135(0), 39-45 (2013)). Studies to characterize this enzyme and to study its uses in decontamination and biofuel production are described in (Liu, et al., Bioresource Technology, 135(0), 39-45 (2013), (Liu, et al., Bulletin of Environmental Contamination and Toxicology, 89(2), 269-273 (2012)). The granular activated carbon (GAC) obtained from Sigma Aldrich (St. Louis, MO) was 20-40 mesh particle size, and DSR A 8x40 GAC was acquired from CalgonCarbon (Pittsburgh, PA). Groundwater samples were collected at the Wurtsmith U.S. Air Force Base.

Batch experiments

The experiment with PFOA-spiked water samples was conducted in 20 mL polypropylene vials containing 15 mL reaction solution and 20 mg granular activated carbon purchased from Sigma Aldrich. The original concentrations of PFOA, HBT and laccase in each treatment container are listed in Table 12. The solution was incubated in dark at 30°C for 60 days with triplicate water samples taken at 0, 15, 30, and 60 days. The HBT and

enzyme were re-dosed at the original dosages after 30 days of incubation.

After the incubation was completed, PFOA concentrations in the aqueous

phase and sorbed on GAC were quantified as described below, and the total

amount remaining was calculated and expressed as concentration as if all in

solution.

For the batch experiment with groundwater, the test was conducted in

50 mL polypropylene centrifuge tube that contained 45 mL groundwater

containing 10 mg GAC (from CalgonCarbon), 2 Unit/mL PS laccase, and

w/wo 20 mM HBT. Control reactors were prepared and tested at the same

time that contained 10 mg GAC but not HBT or laccase in 45 mL

groundwater. The reactors were incubated at 30°C for 60 days, and triplicate

reactors were sampled on the last day of incubation, and analyzed as

described above for PFOA concentrations.

Table 12. The treatment conditions in batch reactor using spiked PFOA

solution.

PFOA PS HBT Fe ³⁺

Treatment

(mhioΐ L ¹ ) (Unit mL ¹ ) (pmol L ¹ ) (mmol L ¹ )

PFOA 1 0 0 0

PFOA+PS+HBT 1 4 40 0

PFOA+PS+HBT+Fe 1 4 40 10

GAC extraction

To quantify the amount of PFOA adsorbed on GAC, a multi-step

extraction was performed on the GAC collected from each sample. Firstly,

the GAC particles were dried in an oven at 45 °C for 24 hours until reaching

constant weight. Ten or twenty milligrams of GAC were extracted with 4 mL

mixture of dichloromethane and methanol (V:V=2: l) in a l0-mL glass tube,

and then the tube was held in the ultrasonic processor (Cole Parmer, Vernon

Hills, IL) for 45 min in a water to ensure the temperature of solution below

50 °C through the extraction process. After extraction, the extracted slurry

was centrifuged at 4200 rpm/min for 20 min to collect the supernatant. The

GAC residue was extracted again for another two times as described above, and all supernatants collected during extraction were combined to about 10 mL. The supernatant was then concentrated to nearly dry using a nitrogen flow. The resulting solution was mixed with 2 mL methanol (HPLC grade Methanol final concentration, 99.9%) and filtered through 0.2 pm cellulose acetate membrane before being transferred to HPLC vial for subsequent analysis. The average recovery was 56%-60% with a Relative Standard Deviation (RSD) of 3.23%.

Micro-column experiments

The setup of the flow-through column experiment is schematically represented in Figures 9A-9C. First, a micro cartridge was packed with 0.1 g GAC and 0.5 g quartz sand, and the filling materials were compacted by two frits on both ends. The column was periodically tapped and the water level in the column was maintained above the solid fillings through the packing process to ensure packing quality. After packing, the column was flushed with HPLC water at 0.1 mL/min for 24 hours before use. For the experiment with PFOA spiked solution, a solution containing 1 pmol/L PFOA, 20 pmol/L HBT, and 4 or 2 Unit/mL PS laccase was continuously passed through a column at a flow rate of 0.1 mL/min driven by a Cole Parmer Masterflex L/S pump. A control treatment was also conducted at the same time under the same condition with a solution containing only 1 pmol/L PFOA. The effluent was collected at pre-selected time intervals for PFOA quantification. A total of 6 L solution was passed through the column, and the procedure lasted for about 40 days. PS enzyme and HBT were used as a model enzyme and mediator in this study.

Column experiments were also conducted to test effectiveness of using the GAC column with and without ECOHR to treat low concentrations levels of PFOA in groundwater collected at Wuftsmith Air Force Base in Michigan, which was intended to provide information for designing a pilot scale column test on site. Each column was packed with 0.5 g CalgonCarbon GAC, and the flow rate was 0.14 mL/min. Instead of mixing prior to loading onto column, the groundwater, the solutions containing HBT, and the solution containing PS laccase were passed through GAC column in sequence at flow rate of 0.14 mL/min until the effluent concentration of each component reached 80% of its influent concentration. Meanwhile, control columns were prepared tested by passing groundwater and PS laccase (without HBT), or groundwater and HBT (without PS). In such scenarios, it is expected that PFAS, HBT and laccase are distributed evenly in GAC column, which ensures better contact between reactants and effective reaction, and when laccase was loaded to the column as the last component, ECOHR was initiated to degrade PFASs. After loading process, all columns were set aside to allow laccase degradation of PFAS to happen on GAC. After 45 days, the columns were flushed by the original groundwater sample, and the different elution curves between treatment and control were used to indicate the PFOA/S degradation, if any, happened on GAC.

Quantification methods

PFOA and PFOS were quantified using the Waters ACQUITY I-class UPLC coupled with XEVO TQD tandem quadrupole mass spectrometry (Milford, MA). An Acquity UPLC BEH 1.7 pm C18 column (Waters, Milford, MA) was used for UPLC separation with a mobile phase consisting of 5 mM ammonium acetate in water (A) and 5 mM ammonium acetate in methanol (B) at a flow rate of 0.3 mL/min by a lO-minute gradient program: 90% A and 10% B at time 0, linearly change to 10% A and 90% B, and then held for 2 min. Electrospray ionization was operated in a negative mode for PFOA and PFOS detection with the capillary voltage at 3 kV and the source temperature at 400 °C. Ultra high purity nitrogen was used as the desolvation and cone gas with the flow rates at 550 and 50 L/hour, respectively. Multiple reaction monitoring (MRM) was used to quantify PFOA, PFOS, and their isotope labeled standards based on the transition patterns m/z = 413 > 369 for PFOA, m/z = 421 > 376 for M8PFOA, m/z = 499 > 99 for PFOS, and m/z = 507 > 99 for M8PFOS. Concentrations were calculated based on the ratio between PFOA or PFOS and its isotope labeled standard in reference to a five-point calibration curve. Solid phase extraction

The effluent samples were purified and concentrated by Waters Oasis HLB cartridges (60 mg) before chemical analysis for PFAS and HBT. The SPE cartridges were conditioned with 3 mL methanol, 3 mL HPLC water, and then another 3 mL HPLC water before loading samples. Thirty milliliter of water samples were loaded on the cartridge slowly, and the cartridges were washed by 3 mL HPLC water to elute interferences. The SPE column was then vacuum dried for 10 minutes before being eluted with 3 x 1 mL methonal and 2 x 1 mL acetonitrile to collect the analytes. At the end of elution, vacuum was applied to the system for another 10 minutes to allow complete elution. The effluents were combined, and then dried under a gentle nitrogen flow, followed by reconstitution to 300 pL with methanol. The sample for HPLC-MS/MS quantification was prepared by mixing 100 pL reconstitute sample with 50 pL M8PFOA and 50 pL M8PFOS.

Laccase Activity Measurements.

Laccase activity was determined spectrometrically by adding 20 pL of sample to 3.4 mL of 1 mM 2, 6-dimethoxyphenol in a citrate phosphate buffer (pH 3.8) in a cuvette with 1 cm light path, the absorbance of which was measured at 468 nm, and one unit is defined as the amount of enzyme that causes a unit absorbance change per minute (I.-Y. Lee, Jung, Lee, & Park, 1999).

Product identification

Product identification was performed on the supernatant sample extracted from GAC samples that have been used for different treatments.

MS full scan (m/z 50 - 1000, resolution R = 120,000) were performed on both treatment and control samples using an Orbitrap Elite high-resolution tandem mass spectrometer (Thermo Scientific, Waltham, MA) in electrospray negative (ESI-) mode. Mass filtration function (Xcalibur 2.1, Thermo Fisher Scientific) was then performed to identify possible products by screening peaks only present in treatment samples but not in controls. For each m/z peak identified, a molecular formula was assigned using Formula Finder with 5 ppm mass error allowed. Subsequently, tandem MS (MS/MS) was performed on the possible products (resolution R = 30,000), and their molecular structures were deduced based on the fragment peaks and general MS fragmentation rules (McLafferty & Turecek, 1993).

GAC characterization

The chemical functionality of GAC samples after batch incubation was qualitatively identified by Fourier Transform Infrared Spectroscopy (FTIR). FTIR spectra were recorded between 4000 and 500 cm ¹ using an AVATAR 360 spectrophotometer (Thermo Nicolet Co., Waltham, MA). Results

Poly- and perfluorinated substances (PFASs) compounds are pollutants of emerging concern due to their world-wide distribution, environmental persistence and bioaccumulation potential (Fujii, et ak, Journal of Water Supply: Research and Technology-AQUA, 56(5):3l3-326 (2007). Giesy, et a , Environmental Science and Technology, 35(7): 1339- 1342 (2001)). Wildlife and human monitoring services have identified perfluoroalkyl acids (PFAAs) in mammal populations across the globe (Lau, et ak, Toxicological Sciences, 99(2):366-394 (2007)). Toxicological research of PFASs in mammals is still in its infancy, but studies indicate they are particularly related to developmental toxicity, immunotoxicity, and hepatotoxicity Lau, et ak, Toxicological Sciences, 99(2):366-394 (2007)). PFASs have been used in industrial applications since the latter half of the 20th century primarily as surfactants suited for harsh conditions because of their extreme chemical and thermal stability (Guo, et ak, Industrial & Engineering Chemistry Research, 47(3):502-508 (2008)). Since the l950's, USAF bases and commercial airports have used AFFF (aqueous film forming foams) to combat fuel fires. Many AFFF products contain PFASs as well as many other organic pollutants. USAF base Wurtsmith, Michigan has used these AFFFs in training exercises since 50's till its decommission in 93. Studies in early 2000s showed levels of groundwater contamination from 3 to 120 pg L ^_1 of perfluorooctanesulfonate (PFOS), perfluorohexanesulfonate (PFHxS); perfluorooctanoate (PFOA), and perfluorohexanoate (PFHxA) (Moody, et ak, Journal of Environmental Monitoring, 5(2):34l-345 (2003)). Treatment methods for the removal of PFAS surfactants from industrial effluents are needed to minimize the environmental release of these pollutants, but have been particularly difficult to develop and apply. Due to the stability of PFASs, a combination of multiple treatment technologies will likely be required to effectively address PFAS

contaminations which often exist in mixtures with co-contaminants present in complex environmental matrices (Merino, et al., Environmental

Engineering Science, 33(9):6l5-649 (2016)). Removal of PFAS surfactants from contaminated groundwater by sorption onto granular activated carbon (GAC) has been heavily investigated (Ochoa-Herrera, et al., Chemosphere, 72(10), 1588-1593 (2008)).

The presence of a variety of functional groups on the GAC’s surfaces allow negative charge-assisted H-bonds to form with the acid groups of PFAAs, thus facilitating sorption (Zhang, et al., Chemosphere, 144:2336- 2342 (2016)). However, GAC sorption is not a destruction approach, requires frequent change-outs, and the spent GAC needs to be disposed through incineration. The high stability of PFAA renders current in situ treatment technologies involving oxidation and microbial degradation not so effective for their destruction (Vecitis, et al., Frontiers of Environmental Science & Engineering in China, 3 (2) : 129- 151. doi: l0.l007/sl 1783-009- 0022-7 (2009)). Advanced oxidation processes (AOPs), such as alkaline ozonation, peroxone, or Fenton’ s reagent, have been shown effective to degrade a wide range of organic contaminants. However, those techniques have very limited success in destructing PFASs from studies reported so far. A couple of direct and indirect photolytic oxidation pathways (Hori, et al., Environmental Science & Technology, 39(7):2383-2388. doi:

l0.l02l/es0484754 (2005), Kormann, et al., Environmental Science & Technology, 25:494-500 (1991), Kutsuna, et al., International Journal of Chemical Kinetics, 39(5):276-288. doi: 10.1002/kin.20239 (2007)), sonochemistry (Cheng, et al., Environmental Science & Technology, 42(21) (2008), Cheng, et al., Environmental Science & Technology, 44(1), 445-450. doi: !0.l02l/es90265lg (2010)), and reductive dehalogenation achieved some PFAS degradation, but are not applicable for in situ treatment applications.

This study was designed to investigate the degradation of PFASs in spiked water samples and PFASs-containing groundwater with GAC and humification enzyme added to induce ECHOR in a batch reactor experiment. Flow through column experiment was also performed with GAC micro column with either spiked solution or groundwater samples, with

humification enzymes loaded to the columns in some treatments to investigate the degradation of PFASs by ECOHR on GAC. A pilot-scale GAC column test was conducted at Wuftsmith Air Force Base in Michigan to study the effectiveness of the treatment scenarios with or without ECOHR induced. GAC sorption isotherms and column breakthrough behaviors were characterized for selected PFAAs to provide necessary fundamental information to assist data interpretation, and identification of degradation products were attempted using High Resolution Mass Spectrometry for samples where ECHOR was induced. In addition, FTIR spectra of GAC with or without ECOHR treatment was obtained to help elucidate the role of functional groups on the GAC surface.

Granular activated carbon, contains various types of functional groups (e. g., free and hydrogen bonded OH groups (Wasewar, et ak, CLEAN-Soil, Air, Water, 37(11):872-883 (2009)). The coupling of GAC and ECOHR treatments is a“Trap and Treat” setup in nature. Since PFOS and PFOA concentrations in groundwater are at ppb levels, a“trap” technique, like GAC sorption, will be important to concentrate PFOS and PFOA in groundwater, making the ECOHR“treat” process more effective. The GAC trap treatment removes PFOA and PFOS from the groundwater, and concurrent/subsequent ECOHR treatment breaks down the PFOA and PFOS adsorbed on GAC to achieve detoxification of the contaminants, and potentially extend the service time of GAC.

Experiments were also designed to determine if GAC can serve the role as a laccase mediator, and therefore humification enzymes may be loaded to GAC and induce ECOHRs that can lead to PFAA destruction. Results

Sorption isotherm

Two types of granular activated carbon have been used in this study for different experiments. GAC purchased from Sigma Aldrich was used to treat spiked solution, because this GAC has been used widely in previous studies to enable data comparison. GAC DSR A 8x40 produced by Calgon Carbon is a product having been used in numerous groundwater remediation applications. Therefore, lab scale tests were performed using this GAC to offer guidance for pilot scale field test. The sorption isotherms of PFOA on both Sigma Aldrich GAC and Calgon Carbon GAC were studied, and the results were fitted well by the Langmuir models (Figure 10A-10B). The Sigma Aldrich GAC (qm = 39.4 mg/g) presented a slightly higher sorption capacity than Calgon Carbon GAC (qe = 31.08 mg/g), but has a much lower affinity (Kl = x) than the latter (Kl = y). The sorption capacity values were slightly lower than the results of a previous report using GAC from

SigmaAldrich (Zhang, et al., Chemosphere, 144:2336-2342 (2016)) and this difference is likely due to variation in granule size.

Sorption of PS laccase on the two GACs was also investigated. The results (Figure 10C-10D) reveals a much larger sorption capacity on Calgon Carbon GAC (174.97 U/mL) than on Sigma Aldrich GAC (1.70 U/mL). Previous studies have demonstrated activated carbon as an effective adsorbent for enzymes through different proposed mechanisms such as ionic interaction, covalent binding, and physical adsorption (Daassi, et ak, Biodeterioration & Biodegradation, 90:71-78 (2014), Liu, et ak, Bioresource Technology, 115:21-26 (2012). Among those mechanisms, some may require addition of external binding agents to ensure stable immobilization. The high sorption capacity on Calgon Carbon GAC is beneficial because laccase can be immobilized in large amount without the use of binding agents. Sorption of laccase on activated carbon will not alter the structure and oxidative ability of the enzyme, and moreover, the enzyme activity remains stable after multiple usage (20 cycles) (Nguyen, et a , Bioresource Technology, 210:108-116 (2016)). Treatment of spiked PFOA solution with SigmaAldrich GAC in batch reactor

In batch reactors, the concentrations of PFOA remained stable in control at the end of l5d, 30d, and 60d incubation, while it was reduced by 15.6%, 30.7%, and 32.4% with the addition of PS laccase, HBT, and ferric ion (Figure 11). Laccase-mediator system has been proven effective in degrading numerous contaminants such as antibiotics, PAHs, and PCBs (Colosi,et al., Environmental Science and Technology, 4l(3):89l-896 (2007), Liang, et al., Chemosphere, 181:320-327 (2017), Weber, et al.,

Environmental Science and Technology, 37(18), 4221-4227 (2003)), and recently its capability to oxidize PFOA during long-term incubation was also shown (Luo, et al., Environmental Science & Technology Letters, 2(7): 198- 203 (2015)). Unlike other contaminants, which usually degraded to negligible levels within hours or a few days, the reduction of PFOA happened slowly over time. It took about 40 days to reach 30% reduction in aqueous sample (Luo, et al., Environmental Science & Technology Letters, 2(7): 198-203 (2015)), which was comparable to the present observation.

This is mainly ascribed to the chemical structures of perfluorinated compounds, which consist of no hydrogen for abstraction and no double bond for addition, leaving direct electron transfer the only possible mechanism. It was found that the presence of certain multi-valent cations (Fe ³⁺ and Cu ²⁺ ) is important for the laccase-mediator reaction systems being effective towards PFOA degradation, because they bridge negatively charged laccase and PFOA in solution so that the free radicals of mediator produced from laccase surface can reach and react with PFOA before they are scavenged (Luo, et al., Environmental Pollution, 224:649-657 (2017)).

Therefore, ferric ion was included in this ECOHR reaction system for the experiment.

Interestingly, the reduction in PFOA concentration was also observed in the treatment with the addition of PS laccase and HBT but no ferric ion, and the removal reached 2%, 25.2%, and 26.8% after l5d, 30d, and 60d of incubations, respectively. Except for l5-day results, the removal of PFOA in the samples without the addition of ferric ions was only slightly less than the removal in the treatments with them added. This is interesting because in previous experiments with ECOHR in aqueous solutions, PFOA did not degrade if an external multi-valent cation was not added (Luo, et al., Environmental Science & Technology Letters, 2(7): 198-203 (2015), Luo, et al., Environmental Pollution, 224:649-657 (2017)). Several possible reasons can be proposed: 1) GAC itself contains some metal ions (Clark, et al., Granular activated carbon: design, operation and cost (1990)) that can serve the role of bridging laccase and PFOA for the ECOHR reaction of PFOA to happen; and/or 2) the aromatic carbon structure in GAC, having affinity to both laccase and PFOA, can serve a similar role as the multi-valent cations to bridge laccase and PFOA.

Extending reaction time from 30d to 60d did not lead to significantly greater PFOA reduction in all the treatments (Figure 11). This indicated that one or more of the factors supporting ECOHR were exhausted after 30 days of incubation. In laccase-mediator systems, mediators was consumed during ECOHR, and the activity of laccase can be inactivated along the reaction as well (Viswanath, et al., Enzyme research (2014)). As found in the previous test with periodical laccase and HBT dosing, laccase activity reduced between two enzyme additions, and HBT was gradually consumed over time (Luo, et al., Environmental Science & Technology Letters, 2(7): 198-203 (2015)). However, laccase activity is likely to be more stable when immobilized on GAC (Liu, et al., Bioresource Technology, 115:21-26 (2012)), ( Nguyen, et al., Bioresource Technology, 210:108-116 (2016)).

Treatment of groundwater with Calgon GAC in batch reactor

An additional batch experiment was performed to test PFOA degradation by laccase-induced ECOHR in PFAS-containing groundwater and CalgonCarbon GAC. Based on the assumption that GAC, rich in phenolic surface functionalities, may provide mediators for ECOHR reactions, a treatment with only PS laccase and GAC, but not HBT, added to the groundwater sample was also tested. Considering the metal ions naturally present in groundwater and GAC, external metal ion was not introduced to any treatment. The PFAS concentrations in groundwater (GW) sample are listed in Table 13. After 60 days of reaction, the PFOA concentration in GW+GAC+PS treatment was the lowest among all four conditions, reaching 23.6% removal compared to the GW+GAC control, while the treatment with HBT addition led to a reduction of 15.2%. This confirms that GAC has components that can serve as mediators, whereas the added HBT did not further enhance ECOHR reactivity, but limited PFOA reactions, probably because the self-coupling of the mediator radicals was promoted as mediator concentrations increased beyond optimum (Luo, et al., Environmental Science & Technology Letters, 2(7): 198-203 (2015)). The finding of GAC itself provides ECOHR mediators is of great application importance, as it eliminates the necessity of introducing an external chemical as mediator for field application of ECOHR. However, the percent removal in groundwater sample was not as great as that in the spiked samples, which was most likely due to the low PFOA concentration (0.872 ppb) in groundwater that limited the contact between PFOA and free radicals, leading to slower reaction. It is therefore beneficial to pass a large volume of groundwater through GAC column and have the PFAS mass built up on the GAC surface to ensure better interactions with enzyme and the free radicals generated during ECOHR. See Figure 15.

GAC characterization

The FTIR spectrum of GAC with PFOA adsorbed and w/wo PS laccase treatment was shown in Figure 16. The spectrum obtained for GAC with only PFOA adsorbed displays absorption bands centered at 2350 cm ¹ and 1500 cm ¹ . The peaks at 2350 cm ¹ showed the CºC binding groups due to symmetric or asymmetric stretching of aliphatic bond or the presence of ketones (Cooke, et ak, Fuel, 65(9): 1254-1260 (1986)). The location of multiple small bands at 1700-1400 cm ¹ in both samples are compatible with the presence of the PFOA characteristic functional groups, namely CF ₂ , CF ₃ , and COO- (Chen, et ak, Rsc Advances, 7(2), 927-938 (2017), Lin, et ak, Environmental Science and Technology, 49(17): 10562-10569 (2015).

Compared to GAC+PFOA treatment, the addition of PS laccase resulted in two extra bands centered at 3000 cm ¹ and 1200 cm ² . According to reference, the alcohol/phenol OH-stretch absorb infrared radiation at 3500- 3200 cm ¹ and the carboxylic acid OH-stretch is at 3000-2500 cm ¹ (Giinzler, et al., IR spectroscopy (2002)). Hence, the band located at 3000 cm ¹ may be caused by either type of OH-stretch, which is most likely a result of GAC surface functionality oxidation by laccase. The band at 1200 cm ² is compatible with the presence of v(C-O-C) vibration, likely resulting from laccase catalyzed reactions to form ether (Weber, et al., Environmental Science and Technology, 37(18), 4221-4227 (2003)). The alterations in GAC surface functionalities indicate the sorption of PFOA on the carbon surface as well as the occurrence of laccase-catalyzed oxidation reactions.

Treatment of spiked PFOA solution with SigmaAldrich GAC micro-column

The sketch and picture of the micro-column packed with GAC are shown in Figures 9A-9B, and the setup is shown in Figure 9C. In this test with PFOA spiked solution (1 pmol/L), the enzyme dosages were 2 unit/mL or 4 unit/mL, and the mediator dose was 20 pmol/L. The breakthrough curves for the GAC micro-columns with or without ECOHR induced was displayed in Figure 12. For PFOA solution with the higher PS laccase dose (4 unit/mL) and HBT, the initial breakthrough of PFOA occurred at the very early stage of the experiment, indicating high concentration of enzyme may block some of the sorption sites. Enzyme immobilization at certain level will result in smaller surface area on GAC (Daassi, et al., Biodeterioration & Biodegradation, 90:71-78 (2014)). Therefore, having a high enzyme input is not always beneficial to the system.

The breakthrough of PFOA happened at similar points (~ 350 bed volume) for both the control treatment with only PFOA in solution and the ECOHR treatment with PFOA, laccase and HBT in solution. From the breakthrough points on, the effluent concentration of the control remained continuously higher than that of the ECOHR treatment. When PFOA reached 100% breakthrough with an effluent concentration of 1 pmol/L, the effluent concentration of the treatment column was still 0.4 pmol/L. This reduced breakthrough concentration demonstrated that PFOA has degraded during ECOHR on GAC column.

Treatment of groundwater with Calgon GAC micro-column

A micro-column experiment was also conducted with Wurftsmith AFB groundwater on CalgonCarbon GAC. Because enzyme sorption on GAC is much less strong than PFOA, loading enzyme in the feed water along with PFOA is less practical in an application scenario, as enzyme will break much earlier than PFOA and be wasted. As such, a different loading scheme was tested in this experiment, where groundwater was first passed through the column until 80% breakthrough of both PFOA and PFOS, and then HBT and PS were each mixed in groundwater and passed through the column in sequence until both reached 80% breakthrough (Figures 13A- 13C). After loading all the reagents, the columns were set aside for reaction with parafilm covering both ends to allow air diffusion. After 45 days of incubation, the columns were flushed by groundwater again, and the PFOA and PFOS concentrations in effluent were determined and shown in Figure 14A-14B. For PFOA concentrations in effluents, no significant difference was observed among treatments. This was most likely due to the low PFOA concentration in groundwater, which limited the sorption on GAC and contact with mediator and enzyme. As for PFOS, the treatment effects were not remarkable within the first 400 BV, but the concentrations were much lower in PFOS+PS and PFOS+PS+HBT treatments than the control from 400 BV to 700 BV (~ 100% breakthrough). These results indicate that laccase treatment will delay the breakthrough of PFOS, and thus extend the service life of GAC column by 1/3 even without mediator added.

Conclusions

In this study, a coupling technology combines GAC adsorption with enzyme catalyzed oxidation reaction was tested in two application scenarios, batch reactor and column reactor, to remove and degrade PFOA and PFOS from water. Both laboratory prepared spiked solution and groundwater collected at contaminated sites were subject to testing. Two kinds of granular activated carbon, including GAC obtained from a scientific supplier and GAC used for common remediation practice were applied in the study to assess the universal applicability of this technology.

In the batch reactor study, more than 20% reduction in PFOA concentration was found using either spiked solution or groundwater after 60 days of incubation. FTIR spectra of GAC before and after incubation indicated the occurrence of oxidation reaction on GAC surface, leading to enriched functional groups on GAC after incubation with enzyme and PFOA solution. The column study using spiked PFOA solution showed lower PFOA effluent concentration in ECOHR implemented treatment than control until full breakthrough. However, when transferring the same setup to groundwater, the impact of ECOHR treatment was not significant, which was most likely due to the extremely low PFOA and PFOS concentrations in groundwater. In general, this environmentally benign coupling technology can be applied in future PFAS remediation practice to destroy PFAS in situ and potentially extend the service life of GAC vessel.

Table 13. Groundwater PFAS concentrations and other properties.

Molecular Detection limit Concentration

Chemicals

Formula (ng/L) (Pg/L)

PFBA (C4) F-(CF ₂ ) ₃ COOH 0.18 0

PFPeA (C5) F-(CF ₂ ) ₄ COOH 0.27 0.212

PFHxA (C6) F-(CF ₂ ) ₅ COOH 0.22 0.487

PFHpA (Cl) F-(CF ₂ ) ₆ COOH 0.08 0.085

PFOA (C8) F-(CF ₂ ) ₇ COOH 0.11 0.872

PFNA (C9) F-(CF ₂ ) _S COOH 1.5 0

PFDA (C10) F-(CF ₂ ) ₉ COOH 1.0 0

PFUA (Cl l) (CF ₂ ) _IO COOH 0.12 0

PFOS (C8) F-(CF ₂ ) ₈ S0 H 0.82 3.917

HBT C ₆ H ₅ N 0 0.19 0

TOC (ppm) 4.6

Conductivity

258.5

(ps/cm)

Table 14. Wave numbers and ascription of the principle bands in activated carbons with PFOA or PFOA+PS adsorbed.

Wavenumbers

Assignments PFOA+GAC PFOA+GAC+PS

(cm ¹ )

alcohol/phenol OH- stretch

3050-2850 carboxylic acid OH- stretch

C=C binding or

2400-2300 * *

ketone

PFOA characteristic

1700-1400

functional groups

1100-1000 v(C-O-C) vibration

Unless defined otherwise, all technical and scientific terms used herein have the same meanings as commonly understood by one of skill in the art to which the disclosed invention belongs. Publications cited herein and the materials for which they are cited are specifically incorporated by reference.

Those skilled in the art will recognize, or be able to ascertain using no more than routine experimentation, many equivalents to the specific embodiments of the invention described herein. Such equivalents are intended to be encompassed by the following claims.