Increased detection of short-chain perfluoroalkyl substances (PFAS) in environmental waters has revealed a need for selective sorbents capable of simultaneous removal of short- and long-chain PFAS. To address this need, a multi-PFAS imprinted polymer composite (mMIP) was synthesized on biochar (BC) substrates (so-called “BC@mMIP”). Use of multiple PFAS templates resulted in increased imprinting factor and PFAS removal compared to single-template MIPs. In ultrapure water containing 100 ng/L (each) of nine PFAS, the BC@mMIP removed three PFAS to below their EPA maximum contaminant level. Column testing in real wastewater effluent revealed competition between sulfate and PFAS for templated sorption sites. Thus, the BC@mMIP is recommended for use treating waters with low total dissolved solid levels, or in series with a pre-treatment process for optimal performance.
Article
Immobilization of a Multi-Template Imprinted Polymer on Biochar for Sorption of Short- and Long-Chain Per- and Polyfluoroalkyl Substances (PFAS)
https://doi.org/10.21203/rs.3.rs-4902900/v1
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PFAS
wastewater reuse
biochar
selective sorption
molecularly imprinted polymer
Per- and polyfluoroalkyl substances (PFAS) are a class of synthetic, fluorinated chemicals identified globally in human, animal, and environmental reservoirs including drinking1–3 and wastewater.4–8 Regulatory action has banned use of long-chain perfluorooctanesulfonate (PFOS) and perfluorooctanoate (PFOA)9–12 causing manufacturers to shift to comparable short-chain PFAS, and a corresponding increase in the prevalence and concentrations of short-chain PFAS in the environment and in (waste)water.6,13 Sorption with activated carbon or ion exchange resins has long been the standard for PFAS separation in water;14–16 however, removal efficiencies for short-chain and carboxylate PFAS are limited.1,17 The presence of organics18 and counterions16 can also reduce PFAS removal efficacy for these sorbents.16,19,20 Thus, there is a need for more selective PFAS sorbents with greater affinity for short-chain PFAS.
Molecularly imprinted polymers (MIPs) are customizable sorbents with high specificity for one or more target “template” compounds.21–23 Functional monomer(s) (i.e., MIP building blocks) undergo weak interactions (e.g., electrostatic, hydrophobic) with the template to form binding sites with tailored size, shape, and affinity. MIPs have recently been investigated as PFAS sorbents.22–25 Studies commonly immobilize the MIP onto granular, carbonaceous substrates22,23 (e.g., biochar) to enable use during large-scale water treatment. Biochar is an ideal substrate due to its high surface area,15,26–28 benign properties, and abundance of modifiable surface functional groups.29,30 Studies have focused on single template imprinting with long-chain PFAS (e.g., PFOS).22–24,31−33 These MIP composites have high templating success due to the high affinity of long-chain PFAS for hydrophobic, cationic functional monomers; however, sorption and templating success of short-chain PFAS by the composites remains low.22,24 For example, our previously synthesized PFOS-imprinted MIP-biochar (100 mg/L) demonstrated 10 times higher loading of PFOS compared to perfluorobutanesulfonate (PFBS)—its short-chain analogue—during competitive sorption of three PFAS (1500 µg/L each).22 Thus, further improvements are needed to increase affinity for short-chain PFAS.
Multi-template MIPs (mMIPs) have been leveraged for effective removal of multiple compounds within the same structural class.34–40 Gao et al. prepared a silica-based mMIP for selective removal of endocrine disrupting compounds (EDCs)37 using 17β-estradiol, estriol, and diethylstilbestrol as templates. Their mMIP composite demonstrated high imprinting factors (IFs, IF = QMIP/Qnon−imprinted polymer) of 1.80 to 6.38 for the templates and four analog EDCs.37 This approach is advantageous because compounds within a class often co-occur and have similar toxicological endpoints. For example, over 240 PFAS analytes comprising 57 sub-classes were identified in groundwater near aqueous film forming foam-impacted sites,41 including precursor compounds which transform to perfluoroalkyl acids (PFAAs) in the environment.42 Therefore, the objectives of this study were threefold: First, to synthesize a mMIP-biochar composite (BC@mMIP) using three PFAS templates of varying head group, size [PFBS, perfluoropentanoate (PFPeA)], and extent of fluorination [6:2-fluorotelomer sulfonate (6:2-FTS)]. The 6:2-FTS template was chosen because of its structural similarity to the previously used PFOS template,22 and because it is a common aerobic biotransformation product of PFAS precursors.43 Second, to compare the physicochemical properties and removal capabilities of the BC@mMIP to BC and to analogous single-template and non-imprinted MIPs. Third, to evaluate in situ sorption and regeneration of the BC@mMIP at (waste)water-relevant PFAS concentrations and in real wastewater effluent.
2.1 Inclusion of multiple PFAS templates improves sorption of short- and long-chain PFAS
BC@mMIP (Fig. 1) exhibited high sorption of both short- and long-chain PFAS [figure 2 (A – D)] due to synergistic interactions between PFAS templates during synthesis (i.e., improved self-assembly of template and functional monomer). For all three templates, IFs were significantly higher for the BC@mMIP than their single-template BC@MIP(X) analogues [figure 2 (B – D) where X is the PFAS template used].
Using multiple PFAS templates lowered individual template selectivity but increased total PFAS removal compared to the single-template and non-imprinted composites. For 6:2-FTS and PFBS, the BC@mMIP selectivity coefficient (Kselectivity; section S5) values were around 1.0 or lower indicating minimal improvement in template selectivity over other PFAS. For PFPeA (which was poorly removed by the other MIP composites) BC@mMIP Kselectivity values were high for the perfluorosulfonic acids (PFSAs) and most perfluorocarboxylic acids (PFCAs) but very low for perfluorobutanoate (PFBA) which reflects a significant improvement in short-chain PFCA by the BC@mMIP [which can also be seen in Fig. 2 (A)]. One notable exception is the low Kselectivity value for PFPeA and PFBA sorption onto BC@MIP(F) which supports a hypothesis from our prior work22 that MIPs templated by long-chain PFAS exhibit poor removal for shorter PFAS, likely due to competition for templated sorption sites. The high sorption capacities for both long- and short-chain PFAAs makes BC@mMIP a better candidate for waters containing multiple PFAS.
Compared to the non-imprinted polymer, the BC@mMIP showed similar sorption of > C6 PFAS (i.e., PFOS, PFHxS, 6:2-FTS, and PFOA) but significantly improved sorption of short-chain PFAS. This finding suggests binding to templated sites is an important sorption mechanism for short-chain PFAS removal, while longer PFAS may participate in both specific and non-specific sorption. Competition at non-specific binding sites may be reduced due to size exclusion of longer PFAS at sites templated by shorter PFAS.
Despite the improvement in short-chain PFAS sorption by the BC@mMIP compared to single-templated composites, a hierarchy of removal: PFSAs > PFCAs and long-chain PFAS > short-chain PFAS was observed. Sulfonic acids have higher electronegativity than carboxylic acids44 which will strengthen electrostatic attraction to the functional monomer quaternary nitrogen moiety. Also, higher PFAS solubility will decrease the adsorptive driving forces for short-chain PFAS onto the hydrophobic BC@mMIP surface and promote their aqueous phase stability. A significant decrease in sorption was observed between PFHxA (C6) and PFPeA (C5), and between PFPeA and PFBA (C4) which corresponds to increased aqueous solubility from 27.1 mg/L (PFHxA) to 197 mg/L (PFPeA) and 1370 mg/L (PFBA, table S4). This observed PFAS sorption hierarchy should be considered when evaluating the BC@mMIP sorption capabilities.
2.2 PFAS head group and chain length significantly impact sorption capacity and mechanism
Isotherm sorption tests revealed multiple BC@mMIP sorption mechanisms: (i) binding to templated-specific sites, (ii) binding to sites templated by other PFAS, and (iii) non-specific sorption onto the composite. PFPeA sorption was described well by the Langmuir model (R2 = 0.989; Fig. 3) with low sorption capacity (7.69 mg/g) which may be due to mode (i) sorption via electrostatic attraction with minimal non-specific sorption from hydrophobic attraction to non-templated MIP surfaces. These results also suggest that PFHxA or another longer-chain PFCA may be a better template due to greater affinity for the functional monomer. PFBS sorption, by contrast, was described best by the Freundlich model (R2 = 0.996) indicating sequential sorption by mode (i) followed by mode (ii) then mode (iii). Preferential sorption at templated sites occurs due to: (a) electrostatic attraction between the PFBS head group and the VBTAC quaternary nitrogen moiety;20,22,23,45 (b) hydrophobic attraction between the PFBS fluorocarbon tail and the VBTAC aromatic ring;22–24,32 and, (c) size and shape selectivity at the templated site. Additional sorption at high PFAS initial concentrations likely occurred from PFAS‒PFAS micelle or hemimicelle formation on the sorbent surface—a phenomenon previously reported at equilibrium concentrations orders of magnitude below the critical micelle concentration32,46 (i.e., approximately 6600 mg/L at 32°C for PFBS).47 Finally, 6:2-FTS sorption was described well by both the Langmuir (R2 = 0.980) and Freundlich (R2 = 0.984) models, indicating a smaller difference in sorption affinity between templated and non-specific binding sites compared to PFBS or PFPeA. The high logKoc and low water solubility of 6:2-FTS (table S4) facilitate non-specific sorption onto a range of sorbents, as demonstrated in section 2.1. Thus, 6:2-FTS is likely to adsorb to both templated and non-specific binding sites across the range of concentrations investigated.
2.3 BC@mMIP removes (waste)water relevant concentrations of PFOA to below the EPA drinking water standard
BC@mMIP demonstrated high PFAS removal at (waste)water-relevant concentrations under ultrapure water, circumneutral pH conditions (Fig. 4). Over 90% removal was observed for four PFSAs (92.5% of PFOS, 99.4% of PFHxS, 91.7% of PFBS, and 97.7% of 6:2-FTS) and two long-chain PFCAs (99.4% of PFOA and 97.7% of PFHxA). Of particular note is the low equilibrium PFOA concentration (0.5 ng/L), well below the EPA maximum contaminant level (MCL) of 4 ng/L. Two additional PFAS included in the new MCL, PFBS (9.1 ng/L at equilibrium) and PFHxS (0.6 ng/L at equilibrium), were similarly treated to below their regulatory concentrations of 2000 ng/L and 10 ng/L, respectively. The equilibrium PFOS concentration (16.4 ng/L), however, was notably higher than the MCL of 4 ng/L which we attribute to higher than expected initial PFOS concentration (217 ± 93 ng/L) and to PFOS contamination observed during SPE concentration. Despite these challenges, it is clear that high PFOS removal is possible. Over 50% removal of three (ultra)short-chain PFCAs was also observed (57.8% of PFPeA, 54.0% of PFBA, and 88.5% of TFA). The decrease in the C6, C5 and C4 PFCA removal can be primarily attributed to their corresponding increase in water solubility.15,48 These results indicate the BC@mMIP could be used to meet PFAS regulatory requirements, making it a promising alternative to lower-affinity sorbents.
High PFAS recovery from spent sorbent (Fig. 4) and fast desorption kinetics (figure S3) indicate long material lifetimes may be possible—an essential characteristic given the higher expected production cost compared to conventional sorbents. PFAS recovery did not follow a specific trend with respect to PFAS chain length, although recovery of PFSAs was generally higher than PFCA recovery (e.g., 84.3% PFHxS recovered compared to 56.4% PFOA and 73.3% PFHxA). This result is in contrast to our previous study with higher recovery as PFAS chain length increased.22 Greater than 100% recovery of PFAS templates is attributed to the incomplete washing of templates during the synthesis process. Unrecovered template mass was relatively low (0.28 µg PFBS, 0.86 µg 6:2-FTS, and 0.15 µg PFPeA) and did not appear to leach during sorption tests. Thus, unrecovered template is not expected to significantly impact sorption performance of this material.
2.4 PFAS sorption to BC@mMIP in column testing was hindered by the presence of wastewater effluent organic matter and total dissolved solids
The high total dissolved solids (TDS) in the PFAS-spiked wastewater effluent during round one of the column testing, particularly from sulfate salts, contributed to lower PFAS removal by the BC@mMIP than was expected from batch test results (TDS = 286 mg/L and sulfate = 16.3 mg/L in round one; Table 1). Complete breakthrough of PFOS and PFOA occurred around 125 pore volumes for the BC@mMIP compared to a much slower breakthrough on F400 (261 pore volumes; Fig. 5 and S4). A similar observation was made for PFBS and PFBA: BC@mMIP breakthrough at 50 pore volumes and F400 breakthrough at more than 100 pore volumes. Competition for sorption sites between PFAS and ionic species (particularly sulfate) has been observed in previous studies of PFAS sorption by MIPs or resins containing quaternary nitrogen moieties.16,19,20 Effectively, high sulfate concentrations render the imprinting approach null. For example, PFAS sorption on the BC@mMIP was comparable to sorption on the non-imprinted biochar polymer composite (BC@NP). The performance of the F400 was less impacted by matrix effects because PFAS sorption onto F400 (hydrophobic attraction to non-specific binding sites) is not impacted by competition with ionic species.49 The impaired PFAS sorption results by the BC@mMIP composite indicate pretreatment of complex aquatic matrices with high sulfate concentrations (e.g., with ion exchange resins) is needed to realize the full PFAS removal potential.
| Collection Date | 4/3/2023 | 4/10/2023 | 4/17/2023 | 4/24/2023 |
|---|---|---|---|---|
| DOC (mg/L) | 9.4 ± 0.3 | 6.7 ± 0.2 | 9.3 ± 0.2 | 10.2 ± 0.5 |
| TSS (mg/L) | 3.6 | 2.0 | 2.4 | 4.8 |
| TDS (mg/L) | 286 | 255 | 130 | 442 |
| pH (S.U.) | 7.12 | 7.20 | 7.15 | 7.38 |
| Chloride (mg/L) | 113 | 64.4 | 71.4 | 192 |
| Nitrate (mg/L) | 5.49 | 0.94 | 0.79 | 1.49 |
| Sulfate (mg/L) | 16.3 | 11.5 | 14.2 | 22.3 |
| Sodium (mg/L) | 41.7 ± 2.4 | 36.2 ± 2.5 | 24.1 ± 2.1 | 53.5 ± 4.2 |
| Calcium(II) (mg/L) | 17.5 ± 0.1 | 16.1 ± 0.1 | 17.1 ± 0.3 | 20.9 ± 0.4 |
| Magnesium(II) (mg/L) | 9.8 ± 0.0 | 9.3 ± 0.1 | 8.7 ± 0.2 | 14.5 ± 0.3 |
| Lead(II) (mg/L) | ND | ND | ND | ND |
| Aluminum(III) (mg/L) | 0.01 ± 0.0 | 0.01 ± 0.0 | 0.01 ± 0.0 | 0.01 ± 0.0 |
| Total iron (mg/L) | 0.5 ± 0.0 | 0.1 ± 0.0 | 0.1 ± 0.0 | 0.1 ± 0.0 |
| ND: below limit of detection | ||||
Sorption of acetaminophen, benzotriazole, and sulfamethoxazole were higher than all PFAS compounds for all four sorbents, with greater than 80%, 70%, and 50% removal, respectively, observed for round one of treatment (Fig. 5). This can be attributed to their high organic carbon distribution coefficients (logDoc; table S4) at pH 7.0 (1.09 to 1.65) and higher sorption to the sorbents and to suspended solids that may be physically removed in the columns. Complexation of PFAS and co-organics with suspended solids and subsequent filtration is particularly evident in the results of the sand-only columns (Figs. 6 and S5) which showed higher than expected removal of all analytes. Fipronil is the exception to the trend observed for co-organics. Fipronil has a high logDoc of 3.77 but exhibited low sorption and fast breakthrough compared to other organic analytes in all columns. Poor removal of fipronil may also be due to its large molecular size that may hinder sorption at templated binding sites.
2.5 BC@mMIP maintains consistent PFAS removal over multiple use cycles while activated carbon loses PFAS affinity over multiple cycles
The PFAS removal capabilities of BC@mMIP were maintained over four cycles of sorption and media regeneration (Figs. 6 and S5). Notably, PFAS removal increased from the second to third sorption rounds, particularly for PFBS, PFBA (Fig. 6), and PFPeA (figure S5). This increase can be attributed to the relatively lower influent TDS in round three (130 mg/L compared to 255 mg/L in round 2; Table 1) which will reduce sulfate competition for PFAS-templated sorption sites. In contrast F400 sorption sites were not well regenerated in this test (total PFAS removal by F400 was more than 20% lower in rounds 2 and 3 compared to round 1). Thus, the consistent PFAS removal capabilities of the BC@mMIP composite is likely to experience longer material lifetimes than traditional activated carbons, potentially offsetting the higher production cost and making it a promising alternative PFAS separation treatment.
High-affinity sorbents are needed to address PFAS contamination—particularly short-chain PFAS—for water sources, including wastewater. The multi-templated BC@mMIP described herein addresses this need with demonstrated capability PFOA, PFHxS, and PFBS treatment below their respective MCLs. We hypothesize a synergistic interaction occurs between multiple template types and the functional monomer that cannot occur in single-template MIPs. Complex interactions between the templates and the BC@mMIP were reflected in sorption isotherm results. Electrostatic attraction dominated interactions with less hydrophobic PFAS (e.g., PFBS) while hydrophobic interactions dominated for more hydrophobic PFAS (e.g., PFPeA and 6:2-FTS). Column testing in real wastewater effluent revealed the need for pre-treatment to further evaluate the capabilities and limitations of this sorbent in high sulfate matrices; however, BC@mMIP is expected to be an excellent option for PFAS treatment in waters with low TDS and organic content such as drinking water treatment and some water reuse scenarios.
4.1 Synthesis of multi-template MIP immobilized on spent coffee grounds biochar
Single- and multi-template MIP-biochar composites were synthesized as outlined in Table 2. Briefly, spent coffee grounds biochar was produced via a two-step process of pyrolysis at 400°C followed by activation with a 1:1 weight ratio of potassium hydroxide at 800°C26 and sieved to obtain the #30–50 mesh fraction (595–297 µm). MIP was polymerized directly on the biochar surface as described in our prior study22 and detailed in section S2.
Table 2 MIP composite composition, BET specific surface area, pore size, and cumulative pore surface area.
4.2 Characterization of biochar-molecularly imprinted polymer composite
Specific surface area (SSA) was measured for all biochar sorbents using Brunauer-Emmett-Teller (BET) analysis with a Micromeritics 3Flex Version 5.00 (Norcross, GA) instrument. The Barrett-Joyner-Halenda (BJH) model was used to obtain average pore diameter and cumulative pore surface area (details in section S4).
The BC@mMIP was imaged using transmission electron microscopy (TEM, Tecnai G2 F20 SuperTwin; FEI, Hillsboro, OR) to visualize the MIP layer. First, a small amount of sample was embedded into resin (see section S3 for recipe), dried at 60°C for 24 h, and sliced into 80 nm sections using a Leica Ultracut 6 microtome and hard diamond knife.
4.3 Batch sorption testing
First, a 4-day sorption test was conducted in ultrapure water at pH 7.00 to evaluate the effect of multiple templates on removal of a range of PFAS. All Table 2 sorbents were included in this test (dose: 100 mg/L). Sorbates included short- and long-chain PFAAs and one polyfluoroalkyl acid (i.e., PFOS, PFHxS, PFBS, PFOA, PFHxA, PFPeA, PFBA, TFA, and 6:2-FTS) at 50 µg/L each. Detailed batch test methods and calculations are provided in section S5.
Second, 4-day sorption isotherms were conducted with each individual template PFAS on the BC@mMIP to understand its sorption mechanisms and capacities. All solutions contained 100 mg/L sorbent to treat initial sorbate concentrations of 5‒5000 µg/L of PFBS, PFPeA, or 6:2-FTS. Details on batch test methods are in section S6 with calculations for Langmuir and Freundlich isotherm model fits.
Third, a 4-day sorption test of approximately 100 ng/L each of the same nine PFAS mixture as before onto 100 mg/L BC@mMIP in ultrapure water (pH 7.00) was performed to probe PFAS removal at (waste)water-relevant concentrations. Following sorption, PFAS loaded on spent BC@mMIP were recovered using the template wash solution (see section S2) for 7 d to estimate PFAS recovery and composite lifetimes. Additional method details are provided in section S7.
Details on PFAS quantification via liquid chromatography tandem mass spectrometry (LC-MS/MS) as well as sample concentration via solid phase extraction are provided in section S13 and section S14.
4.4 Column testing
Column tests were performed using wastewater effluent spiked with the same nine PFAS mix plus four organic compounds (i.e., acetaminophen, fipronil, sulfamethoxazole, and benzotriazole) at approximately 20 µg/L each to determine sorption capabilities and opportunities for in situ media regeneration for wastewater treatment. Triplicate columns were run in upflow configuration at 1 mL/min for (i) sand (control), and 1 wt% of (ii) F400 (commercial AC), (iii) BC@NP, and (iv) BC@mMIP amendments with sand. PFAS sorption and recovery were evaluated over four sequential cycles of sorption (6-day treatment) and regeneration (low flow 6-hr extraction with 18-hr deionized water wash). Sorption and regeneration samples were collected as 1-mL composites and stored at -20°C prior to analyte quantification. Additional details on column assembly and test methods are discussed in section S8.
Author Contributions
The manuscript was written through contributions of all authors. All authors have given approval to the final version of the manuscript.
Funding Sources
This work is supported by the National Science Foundation Graduate Research Fellowship Program under Grant No. DGE-1762114 and by the CAREER Award (CBET 2143301). Any opinions, findings, and conclusions or recommendations expressed in this material are those of the authors and do not necessarily reflect the views of the National Science Foundation. Part of this work was conducted at the Molecular Analysis Facility, a National Nanotechnology Coordinated Infrastructure (NNCI) site ate the University of Washington with partial support from the National Science Foundation via awards NNCI-2025489 and NNCI-1542101. This work was also supported by the University of Washington Co-Motion Innovation Gap Fund (68-3871) and CoMotion STEP Grant (68-6769), and by the King County Wastewater Treatment Technology Research Award Program (GR027109).
Acknowledgements
The authors would like to thank: Dr. Martin Sadilek at the University of Washington Department of Chemistry Mass Spectrometry Facility for assistance and expertise with LC-MS/MS quantification; Ellen Lavoie for assistance and expertise with TEM imaging; Fanny Okaikue-Woodi, Anusha Srivastava, Rayna Morales Lumagui, Sophia Jessum and Joshua Chong for their assistance with data collection; Phillip Cox for his assistance with BET analysis; and collaborators at the King County Wastewater Treatment Division for their support in wastewater effluent sampling. Part of this work was conducted at the Washington Clean Energy Testbeds, a facility operated by the University of Washington Clean Energy Institute.
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There is NO Competing Interest.