Skip to main content
NCBI home page
ACS AuthorChoice logoLink to ACS AuthorChoice
. 2022 Dec 28;3(1):121–130. doi: 10.1021/acsestengg.2c00256

Dissolved Methane Recovery and Trace Contaminant Fate Following Mainstream Anaerobic Treatment of Municipal Wastewater

Stephen M Galdi ‡,§,*, Aleksandra Szczuka §,∥, Chungheon Shin ‡,§, William A Mitch †,§, Richard G Luthy †,§,*
PMCID: PMC9841518  PMID: 36660091

Abstract

graphic file with name ee2c00256_0006.jpg

Anaerobic treatment of municipal wastewater with the staged anaerobic fluidized bed membrane bioreactor (SAF-MBR) shows promise to transform secondary wastewater treatment into an energy-positive process. However, the dissolved methane in SAF-MBR effluent needs to be recovered to reach net energy positive. To recover this methane for energy generation, an air stripping system was constructed downstream of a pilot-scale SAF-MBR facility and operated for over 80 days. The process removed 98% of effluent dissolved methane, and with the addition of intermittent disinfection recovered an average of 90% of the dissolved methane. The exit gas from air-stripping comprised 1.5–2.5% methane and could be utilized by blending with biogas produced from primary solids digestion and the SAF-MBR in an on-site combustion process. The direct energy costs for air stripping methane are <1% of the energy recoverable from the dissolved methane, not accounting for siloxane or sulfide scrubbing. Only siloxanes were observed at levels impacting combustion in this study, with 1.6 mg Si/m3 present in the blended biogas and air stripping mixture. The fate of a subset of trace organic contaminants was examined across the air stripping unit to check for aerobic degradation by methanotrophs or other opportunistic aerobes. Only 1,4-dioxane and benzotriazole showed statistically significant removal among 17 compounds screened, with 0.53 ± 0.13 and 0.34 ± 0.15 fraction removal, respectively. Our results indicate that air stripping is an energy efficient and robust technology for dissolved methane removal and onsite utilization for heat and electricity generation from anaerobic treatment of municipal wastewater.

Keywords: anaerobic, wastewater, energy, siloxanes, methanotroph

Introduction

Mainstream anaerobic treatment has gained attention for its reduction in energy and solids handling costs compared to conventional activated sludge processes.1,2 One such technology is the staged anaerobic fluidized bed-membrane bioreactor (SAF-MBR), which has demonstrated potential to be the first net energy positive, secondary wastewater treatment process.3,4 The SAF-MBR process uses an expanded bed of granular activated carbon with attached anaerobic microbial biofilms followed by ultrafiltration (40 nm pore size) membranes that retain suspended solids and colloidal organic matter. The SAF-MBR treats primary effluent at hydraulic residence times (5.3–10 h) comparable to conventional activated sludge treatment.4 Previous research has indicated that concentrations of pharmaceuticals and disinfection byproducts precursors were lower in a pilot-scale SAF-MBR effluent than in a parallel full-scale activated sludge system.5

As in other high throughput anaerobic treatment systems, the SAF-MBR effluent has a significantly higher proportion of dissolved methane than conventional anaerobic digestors (ADs).6 Between 15 and 50% of produced methane from the SAF-MBR process may remain dissolved in the effluent. This must be recovered for net energy production and to avoid methane emissions with high greenhouse impact.7 Current efforts on dissolved methane recovery are focused on membrane contactors,8,9 due to the diversity of utilization options for the concentrated gas stream and high separation efficiency. Other energy efficient options being examined include de-gassing membrane-associated biological contactors,10 vacuum degasification,11 air stripping separation,12 and down-hanging sponge reactors.13 Of these, air stripping and down-hanging sponge reactors deliver the lowest recovered methane purity but with the advantage of low energy costs and high (>99%) methane removal efficiency. In addition, the transition from an anaerobic reactor to an aerobic condition may provide a favorable environment for methanotrophs whose metabolisms utilize non-specific mono-oxygenase capable of degrading a variety of emerging contaminants.14,15 While a down-hanging sponge reactor may be the best choice for encouraging biodegradation, the 2 h residence time required makes it difficult to scale.13 Finally, few studies of mainstream anaerobic systems have begun to assess the needs for siloxane or other gas impurity polishing in biogas or recovered dissolved methane.8 Siloxanes appear in wastewater due to their use in many personal care products. Siloxanes are volatile and poorly soluble,16 and these compounds may cause issues for energy recovery by forming silica scale inside engines and exhaust pipes.17−19

To achieve methane recovery without using complex membrane separation processes, an air stripping methane recovery system with countercurrent air flow was evaluated at pilot-scale downstream of a SAF-MBR facility. The pilot work was divided into two phases depending on the disinfection regime to control biological growth and methane oxidation in the air stripping unit: (1) bi-weekly disinfection dosing to clear biological blockages and (2) daily disinfection dosing to minimize methane oxidation in the column. During phase 2, a spiking experiment of a subset of pharmaceuticals and other contaminants of concern was conducted to examine whether they are removed beyond expected volatilization, indicating possible biodegradation. Additionally, screening for siloxanes in the SAF-MBR biogas and air stripped gas was conducted to determine the importance of siloxane presence in recovered dissolved methane. Phase 2 was interrupted due to the onset of the COVID-19 pandemic and was divided into a short-term high frequency test and a long-term experiment.

Materials and Methods

Air Stripping Pilot Unit

The air stripping unit received effluent from a pilot-scale SAF-MBR system that treats domestic sewage scalped from a sewer at the Codiga Resource Recovery Center on the Stanford University campus as described previously.20 The sewage was treated using a microscreen prior to the SAF-MBR. The air stripping system was constructed using a 20.3 cm (8 in.) diameter column with 1.5 m of 1.6 cm polypropylene Pall Rings. The media was supported by 13 mm (1/2 in.) 316 stainless steel form-soldered mesh between two gasketed flange connections (McMaster-Carr). A schematic of the air stripping pilot is shown in Figures 1, and S.1 shows a picture of the SAF-MBR facility and the air stripping pilot unit. The system treated an average of 3.8 L/min (∼1 gal/min) of SAF-MBR effluent. SAF-MBR effluent was received for 6–8 min at a time followed by 2 min relaxation intervals designed to allow the membrane permeate pump to moderate transmembrane pressure and biofouling.21 A flow distribution plate at the top of the air-stripping unit was constructed out of a 1.6 mm (1/16 in.) polyvinyl chloride (PVC) sheet perforated with 3.2 mm (1/8 in.) holes every 38 mm (1.5 in.). Standing water (30.5 cm or 12 in.) was maintained at the bottom of the unit to direct an average of 6.14 L/m air flow (1.66 average air/water volumetric ratio) into the base of the column via 3.2 mm (1/4 in.) PVC air tubing and an aquarium pump (Hygger 6 W). Off-gas was routed away from the work site through an elevated dispersion stack with a Luer lock sampling port for methane sampling. A submerged sampling port was used to sample effluent water before exiting the unit and another sampling port was placed on a T-joint in the influent pipe just upstream of the column entrance to sample influent water from the SAF-MBR. Methane leaks were checked with a Sensit HXG-2d portable sensor and any leaks at through-wall tubing connections were sealed with a polytetrafluoroethylene sealant.

Figure 1.

Figure 1

Open in a new tab

Schematic of the air stripping pilot unit constructed at the SAF-MBR facility at the Codiga Resource Recovery Center. The design packing depth was 1.5 m for an effluent flow rate of 3.8 L/m. Hypochlorite and potable water were used to flush the column with disinfectant during system downtime.

Design

The counter-flow system design employed the relationship between tower height and methane removal using the relationship shown in eq 1, based on the two film theory of mass transfer with mass transfer rates calculated using the Onda correlation.22−24

graphic file with name ee2c00256_m001.jpg 1

where L = tower height [m], Q = water flow rate [m3/s], A = tower area [m2], C0 = influent concentration [mg/L], Ce = desired effluent concentration [mg/L], Cz = concentration at height z [mg/L], Cs* = equilibrium concentration [mg/L], KL·a = mass transfer rate [1/s], and S = stripping factor = Inline graphicHCC = dimensionless Henry’s constant

The size of the unit, 1.5 m in height, was based on a design flow of 3.8 L/m (∼1 gpm), 1.5 air/water ratio, 20 cm column diameter, 1.6 cm Pall Ring packing, and a desired 97% removal of dissolved methane. Full design assumptions and limitations are listed in the Supporting Information, S.2.

Methane Sampling

Dissolved methane from the air stripping unit effluent was sampled by flushing and filling 525 mL borosilicate serum bottles (Wheaton 223952) and sealing them with 30 mm aluminum seals (Wheaton 224187-01) and bromobutyl rubber stoppers (Wheaton W224100-342) as described by Shin et al.7 Effluent dissolved methane was extracted into the gas phase by replacing 6 mL of the sealed 525 mL water sample with nitrogen gas and shaking the bottles 1 min by hand, equilibrating for at least 2 h at 30 °C to decrease solubility variation, and shaken 1 min by hand again before analysis. Influent dissolved methane was sampled using 10 mL glass BD Vacutainers (BD 366430) and 10 mL borosilicate crimp top vials (Wheaton 223686), which were injected with 5 mL of sample water that rapidly degassed, as described by Alberto et al.25

Extracted methane in the sample headspace was measured using gas chromatography with a thermal conductivity detector (GOW-MAC series 580). One mL of headspace gas was injected. The system used an Altech CTR 1 column, helium carrier gas, and featured a detection limit of 0.5% volume. Methane concentrations in the dissolved phase were then calculated using Henry’s law adjusted for temperature,26 resulting in effective detection limits for dissolved methane of 0.15 and 6.4 mg/L for the borosilicate bottles and vacutainer vials, respectively.

Field Sensors

In addition to methane, water samples were collected and analyzed for dissolved oxygen (YSI ProODO) and pH (Oakton UX-35805-05). Samples were collected in sealed amber vials (headspace free) and analyzed within half an hour of collection to minimize oxygen consumption. Digital sensors were deployed to collect flow and temperature data in real time at the air stripping unit. Results were recorded via an Arduino Uno control board onto an SD card at 1 min intervals. Temperature data were averaged over 20 s intervals to reduce noise. Effluent water temperature was measured via a MAXIM DS18B20 digital temperature probe. A PULSAFEEDER MTR-103G mechanical flow meter with reed switch measured flow rate during phase 1, which was set by a gate valve 10 pipe diameters downstream. During phase 2, flow was measured via a Georg Fischer (GF) 3-2551-P0-12 magmeter and controlled with Bioforcetech’s Plexus control board and a GF EA25 actuated ball valve with a positioner board. Pressure drop in the column was monitored with a 3.2 mm (1/8 in.) diameter manometer installed between the air inlet and media support grate.

Column Disinfection

Disinfection during phases 1 and 2 were conducted with sodium hypochlorite (NaOCl) at concentrations of 350 and 300 mg/L as Cl2, respectively. For phase 1, 150 L [40 gal] of 350 mg/L as Cl2 NaOCl in tap water were pumped down through the column at 7.6 L per minute (L/min) [2 gpm] once every two weeks for 20 min. Once during phase 1 and again between phases 1 and 2, a physical cleaning of the air stripping media was performed to reduce biomass clogging. The media was removed from the column, soaked for 12 h in 350 mg/L Cl2 solution, agitated with a brush, and rinsed with potable water. To improve performance and limit biomass buildup, the procedure during phase 2 employed an automated daily dosing system controlled by an Arduino Uno microcontroller that pumped 300 mg/L as Cl2 NaOCl in tap water at 7.6 L/min down through the column for 5 min daily. In both phases, the wastewater effluent flow was halted during disinfection. Phase 1 continued for 120 days before changing the disinfection procedure in phase 2, which continued for 80 days of intermittent testing.

Energy Consumption

The energy requirement (Eair, kW h/m3) of the pilot-scale air stripping tower was estimated with eq 2.

graphic file with name ee2c00256_m003.jpg 2

where kWad is the power in kW, Qliquid is the influent flow rate of the stripping tower in m3/h, and εairpump is the air pump efficiency (65%).27

The adiabatic power requirement of the air pump (kWad) was calculated with eq 3.28

graphic file with name ee2c00256_m004.jpg 3

where k is the adiabatic constant of 1.4 for air, Qair is the volumetric flow rate of air in m3/h, p1 is absolute inlet pressure in kPa, and p2 is absolute discharge pressure in kPa.

Carbon Footprint

The carbon footprint of the pilot-scale air stripping tower was assessed by considering direct and indirect emissions. The direct emissions include dissolved methane concentration in effluent with its respective 100 year global warming potential (GWP100) value of 34 kg-CO2 equiv/kg-CH4.29 The indirect emissions include the use of sodium hypochlorite used for column disinfection, which has an emission factor of 3.12 kg-CO2 equiv/kg-NaOCl30 and the use of electricity (0.38 kg-CO2 equiv/kW h, US EIA, 2020).31

Recovery Combustion Assumptions

Blending requirements for biogas were calculated for the stoichiometric combustion of primary, secondary, and dissolved effluent produced methane as observed at the Codiga Resource Recovery Center. Two scenarios of off-gas oxygen concentration were examined: maximum oxygen flux within the column and atmospheric oxygen, with the flux calculated using two film mass transfer theory and an assumed aqueous concentration of 0 through the entire column. Further assumptions and equation descriptions are available in the Supporting Information (eq S2.8). Electricity recovery was calculated assuming a 56% generation efficiency using a combined cycle gas generator.32 Hydrogen sulfide was assumed negligible based on previous data on the wastewater consistency at the Codiga Resource Recovery Center.4,33

Trace Organic Contaminants

Removal of 17 contaminants of emerging concern was evaluated through the pilot-scale SAF-MBR and the downstream air stripping unit. The contaminants targeted are frequently detected in sewage and represent a range of micropollutants that are either poorly removed through sorption and biological degradation during traditional wastewater treatment.34 The following contaminants were targeted: 1,4-dioxane, atenolol, benzotriazole, bezafibrate, carbamazepine, ciprofloxacin, diclofenac, diuron, fipronil, hydrochlorothiazide, ibuprofen, naproxen, N-nitrosodimethylamine (NDMA), oryzalin, ranitidine, sucralose, and sulfamethoxazole.

Contaminant removal was evaluated by comparing influent and effluent concentrations across the air stripping column. Due to the low concentrations of trace organic contaminants in the SAF-MBR effluent, a premixed contaminant stock solution made with SAF-MBR effluent was used to increase the concentration of contaminants entering the air stripping unit. The unit was operated continuously for 1 h during week 2 of phase 2 in this configuration with target influent contaminant concentrations of 250 nM for most contaminants, representing the high end of typical concentrations in secondary effluent.34 Three exceptions were NDMA, which was supplemented at a concentration of 4 nM, 1,4-dioxane, which was supplemented at a concentration of 1000 μM, and sucralose, which was not supplemented. Samples (500 mL) were taken at the air stripper influent and effluent.

Contaminant concentrations were determined by gas or liquid chromatography mass spectrometry. 1,4-Dioxane and NDMA were analyzed by gas chromatography mass spectrometry (GC–MS) using methods described previously,35,36 with detection limits of ∼10 and ∼0.1 nM, respectively. Briefly, 40 mL samples were extracted into 3 mL of MtBE for 1,4-dioxane quantification, and 50 mL samples were extracted into 4 mL of dichloromethane three times (the three extracts were combined and blown down to ∼1 mL) for NDMA quantification. The remaining compounds were analyzed by liquid chromatography tandem mass spectrometry (LC–MS/MS) with methods adapted from McCurry et al. and Szczuka et al.5,36 Sucralose samples were filtered with 0.7 μm glass fiber filters and analyzed directly with a 0.5 μg/L detection limit. Fourteen compounds were concentrated by solid phase extraction. Briefly, 250 mL samples were passed through 6 mL Oasis HLB cartridges (Waters Corp.) pre-rinsed with methanol and eluted with 12 mL of methanol; the resulting extracts were analyzed, and the method detection limits ranged from 0.5 to 2 nM. The LC–MS/MS method and detection limits are detailed in S.3.2.

Siloxanes

Three grab samples were taken of SAF-MBR biogas and air stripping off-gas during the automatic disinfection phase of air stripping operation. Six-liter samples were collected on an XAD-2 solid phase sorbent tube at a flow rate of 0.2 L/min and sent to ALS Global for extraction and GC–MS analysis per ASTM method D8230-19. Detection limits and data are available in S.6.1.

Results

Dissolved Methane Recovery

Dissolved methane was removed from the SAF-MBR effluent at efficiencies that exceeded the design efficiency for both phases of the pilot experiment. Methane removal averaged greater than 98%. Methane recovery in the gas state varied depending on the frequency of stripping column disinfection and cleaning. During phase 2, which increased disinfection frequency from biweekly to daily, average methane recovery increased significantly (Welch’s t-test P = 0.0004) from an average of 58 to 90%, as shown in Table 1.

Table 1. Methane Removal and Recovery during Testing with Biweekly and Daily Disinfection to Control Biological Growth and Methane Oxidationa.

parameter phase 1 biweekly disinfection phase 2 daily disinfection replicates phase 1, 2 (n)
methane % removal 99.2 ± 1.1 98.7 ± 0.7 45, 54
methane % recovery 58.0 ± 48.1 90.3 ± 30.0 39, 54
air stripping influent concentration [mg/L] 18.8 ± 8.3 25.0 ± 5.5 45, 54
air stripping effluent concentration [mg/L] 0.145 ± 0.201 0.327 ± 0.160 45, 54
off-gas concentration [% v/v] 1.28 ± 0.71 2.11 ± 0.48 39, 54
effluent temperature [°C] 23.9 ± 3.3 18.6 ± 1.6 38,000, 41,000
water flow rate [L/min] 4.81 ± 1.87 3.83 ± 0.36 38,000, 41,000
Open in a new tab
a

Temperature and flow data from continuous data loggers.

While there were changes between the two testing phases (i.e., disinfection schedule, flow control, seasonal effects on temperature, and SAF-MBR dissolved methane concentration), the temporal visualization of recovery rate and disinfection events for phase 1 (Figure 2a) and the first days of phase 2 (Figure 2b) indicates a correlation between column disinfection and improved methane recovery. Two drops in methane recovery greater than 30% occurred during phase 1: the first after one month of operation and the second one month after manual cleaning of the column to alleviate biomass clogging in the air stripping media. In both cases, the removal of methane did not decrease as substantially, suggesting the drop in dissolved methane was due to transformation rather than volatilization. Additional temporal data from the extended portion of phase 2 can be found in Supporting Information S.5.

Figure 2.

Figure 2

Open in a new tab

Disinfection events during phase 1 (a) and the start of phase 2 (b), plotted with methane removal and recovery in the off-gas. Biofilm clogging required manual cleaning in addition to biweekly disinfection as highlighted in phase 1. Disinfection dosing was interrupted for two days between March 15–17 during phase 2, which correlated with a decrease in methane recovery but not removal efficiency. Methane fractions are calculated based on the mass flow of dissolved methane leaving the SAF-MBR at the time sampled.

Energy Balance and Carbon Footprint

The pilot-scale air stripping tower energy requirement was assessed in phase 2 with the automated disinfection system. Under operating conditions with influent flow rate of 3.8 L/min and volumetric air flow rate of 6.2 L/min, the measured pressure drop across the column measured by the manometer was 276 Pa (2 cm water), resulting in an energy requirement of 2.1 × 10–4 kW h/m3. This energy requirement is less than the energy required for other strategies of dissolved methane recovery: 7.2 × 10–3 to 9.0 × 10–3 kW h/m3 for degassing membranes37−39 and 6.9 × 10–3 kW h/m3 for vacuum degasification.11 All these energy costs are less than the potential electricity generated from the recovered methane: 0.156 kW h/m3 for 56% efficient combined cycle generation,32 20 mg/L dissolved methane, and 90% recovery. The air stripping system requires ∼90% less energy than alternative methane separation processes, and uses less than 1% of the potential electricity generated. The achieved dissolved methane removal efficiency through the air stripping tower (98.7%) is also greater than high purity separation processes: 83 and 90% for degassing membranes with unit energy estimates38,39 and 94% for vacuum degasification.11 These different energy requirements and dissolved methane removal efficiencies result in different carbon footprints (kg-CO2 equiv per m3 of treated water). Figure 3 summarizes the carbon footprint of each process. The experimental removal is on par with previous air stripping and down hanging sponge reactor work, with ∼99% removal of dissolved methane.12,13 While the methane removal and recovery are the same in membrane and vacuum separation, the down hanging sponge had lower recovery (76.813 vs 90% in this study) and the previous air stripping work had no planned recovery. These different energy requirements and dissolved methane removal efficiencies result in different carbon footprints (kg-CO2 equiv per m3 of treated water). Figure 3 summarizes the carbon footprint of each process.

Figure 3.

Figure 3

Open in a new tab

Carbon footprint and methane recovery of air stripping (current study), degassing membranes and vacuum separation. For the vacuum degasifier, a mean energy requirement of 8.4 × 10–3 kW h/m3 was considered from Lee et al., 2020,11 and a mean dissolved methane removal efficiency of 86.5% was considered from Evans et al., 201938 and Rongwong et al., 2018.39 For all cases, an influent dissolved methane concentration of 20 mg-CH4/L was assumed. The yellow dotted box on the degassing membrane column represents the possibility of additional carbon footprint from chemical membrane fouling controls.

Although the air stripping operation required sodium hypochlorite (NaOCl) for column disinfection, resulting in an additional carbon footprint of 0.006 kg-CO2 equiv/m3, the overall carbon footprint of the air stripping tower (0.016 kg-CO2 equiv/m3) is smaller than for the other options (0.043 kg-CO2 equiv/m3 for a vacuum degasifier and 0.095 kg-CO2 equiv/m3 for a degassing membrane), due to the higher average dissolved methane removal efficiency of air stripping. This implies that the dissolved methane removal efficiency is the most sensitive factor in carbon footprint in terms of curtailing fugitive emissions. It is worth noting that there is no other study that assessed chemical requirements for fouling controls for degassing membranes to date, which could increase the total carbon footprint as for the air stripper.

The off-gas from the air stripping tower, containing O2 and stripped CH4, needs to be connected to a combined heat and power (CHP) unit to completely combust the stripped CH4 with gaseous CH4 from the SAF-MBR and CH4 from an AD treating primary settled solids. To achieve complete combustion of methane, the O2-to-CH4 molar ratio for the CHP needs to be at least 2. Air stripping operations at high air-to-water ratios should be avoided because excess O2 in the stripping tower off-gas can dilute CH4 concentrations below levels needed to sustain combustion, resulting in CH4 release to the atmosphere. To assess the stoichiometry of O2 and CH4 for co-combustion through a CHP unit, we considered the pilot-scale SAF-MBR system4 that provided influent for the pilot-scale air stripping tower in this study.

The pilot system, as shown in Figure 4, consisted of a primary treatment system followed by the SAF-MBR for secondary treatment. The system treated 800 mg/L COD domestic wastewater. We assumed the SAF-MBR flow rate (Qw) was the same as that to the air stripper for the purposes of this study (3.8 L/min). Among the 300 mg/L of COD removed during primary treatment, 90% of the COD is assumed to be converted to gaseous CH4 in a side stream AD, yielding a molar production rate of 0.016 mol-CH4/min. The SAF-MBR converted 40% of the remaining COD to gaseous CH4, yielding a molar production rate of 0.012 mol-CH4/min. SAF-MBR solids digestion was assumed negligible. The air stripping tower achieved an average of 90% recovery of dissolved CH4 in the SAF-MBR effluent (20 mg-CH4/L), for a molar production rate of 0.004 mol-CH4/min. The overall molar CH4 production rate of 0.032 mol-CH4/min (0.016 + 0.012 + 0.004) was considered as a fuel flow rate to the CHP. In phase 2, the applied air flow rate from the stripping tower was 6.6 L/min (corresponding to an air-to-water ratio of 1.7), which can supply oxidant to the CHP at a flow rate of 0.055 mol-O2/min. For this case, the O2-to-CH4 molar ratio to the CHP is 1.7, requiring an additional 0.009 mol-O2/min of oxygen from makeup air for completely combusting the produced CH4.

Figure 4.

Figure 4

Open in a new tab

Stoichiometry assessment of O2 and CH4 from the pilot-scale SAF-MBR (Shin et al., 2021)4 and stripping tower (this study) for co-combustion of CH4 via a CHP generator. Orange dotted lines represent the molar production rate of CH4, and blue dotted lines represent the molar flow rate of O2 to the CHP generator.

Trace Contaminant Fate

The trace contaminant spiking experiment comprised 14 compounds, with only 1,4-dioxane and benzotriazole showing statistically significant (P < 0.05) removal, as shown in Table 2. Although 1,4-dioxane has the greatest Henry’s constant of the contaminants removed, none of the contaminants tested, including 1,4-dioxane are known to be well-removed by air stripping alone.40

Table 2. Trace Contaminants with Dimensionless Henry’s Constants (Hcc), Spiked Concentrations, and Measured Removal Expressed As Effluent Concentration As Well As Fraction Removed by Passage through the Air Stripping Processa.

        [nM except 1,4-dioxane in μM]
  
name chemical formula MW Hcc at 25 °C spiked influent air stripped effluent fraction removal
1,4-dioxane[μM] C4H8O2 88.11 2.07 × 10–4b 1090 ± 14 525 ± 134 0.52 ± 0.13
benzotriazole C6H5N3 119.12 6.13 × 10–6c 222 ± 46 148 ± 52 0.34 ± 0.15
ciprofloxacin C17H18FN3O3 331.35 2.09 × 10–17c 250 ± 46 182 ± 51 0.28 ± 0.07
NDMA C2H6N2O 74.08 1.53 × 10–14d 3.9 ± 1.9 3.1 ± 1.1 0.18 ± 0.11
sulfamethoxazole C10H11N3O3S 253.28 7.44 × 10–5e 273 ± 47 241 ± 55 0.08 ± 0.33
ibuprofen C13H18O2 206.29 2.69 × 10–11e 214 ± 24 202 ± 18 0.05 ± 0.14
ranitidine C13H22N4O3S 314.4 6.11 × 10–5c 255 ± 7 241 ± 39 0.05 ± 0.18
bezafibrate C19H20ClNO4 361.82 1.40 × 10–13f 240 ± 8 240 ± 0 0 ± 0.03
diclofenac C14H11Cl2NO2 296.15 1.93 × 10–10c 163 ± 75 168 ± 95 0 ± 0.12
naproxen C14H14O3 230.26 1.39 × 10–8c 241 ± 22 238 ± 25 0.01 ± 0.07
diuron C9H10Cl2N2O 233.1 2.02 × 10–8e 318 ± 73 316 ± 70 0 ± 0.04
fipronil C12H4Cl2F6N4OS 437.15 3.36 × 10–8e 217 ± 21 216 ± 21 0 ± 0.09
atenolol C14H22N2O3 266.34 5.60 × 10–17c 267 ± 69 266 ± 81 0 ± 0.12
sucralose C12H19Cl3O8 397.64 1.61 × 10–17e 27 ± 3 26 ± 4 0 ± 0.23
hydrochlorothiazide C7H8ClN3O4S2 297.7 1.79 × 10–10c 264 ± 16 268 ± 7 –0.02 ± 0.04
oryzalin C12H18N4O6S 346.36 7.76 × 10–8e 129 ± 112 133 ± 120 0 ± 0.07
carbamazepine C15H12N2O 236.27 4.48 × 10–9e 205 ± 17 208 ± 16 –0.01 ± 0.06
Open in a new tab
a

Bolded compounds had statistically significant removal (t-test P < 0.05).

b

Ondo and Dohnal, 2007.41

c

US EPA, 2012.42

d

Haruta et al., 2011.43

e

Sander, 2015.44

f

Kehrer, 2008.45

Additional measurements of water quality in the air stripping column, shown in Table 3, highlight the possibility of a small but persistent community of active microbes in the column that reduced the measured COD during phase 2. There was no significant (two-tailed, paired t-test, P = 0.71) change in ammonia concentrations, suggesting that nitrification is limited by kinetics or disinfection. Heterotrophic bacteria accumulated in the column, and while the methane losses to oxidation averaged only 10%, the effluent oxygen levels were less than saturation (8.9 mg/L at 20 °C). The increase in pH by air stripping is expected for equilibration of anaerobic effluent with the atmosphere due to oversaturated carbon dioxide leaving the water, but the pH remained below the 9.3 pKa of ammonium, limiting ammonia volatilization.

Table 3. Water Quality Parameters Measured during Air Stripping Column Operationa.

    biweekly disinfection daily disinfection n
SAF-MBR effluent DO [mg/L] 0.27 ± 0.07 0.20 ± 0.09 27
  pH 6.77 ± 0.03 6.38 ± 0.03 30
  COD [mg/L] 103 ± 49 84 ± 24 11
  ammonia-N[mg/L]   55 ± 7 16
air stripping unit effluent DO [mg/L] 4.54 ± 0.16 6.23 ± 0.26 27
  pH 7.58 ± 0.02 7.50 ± 0.03 30
  COD [mg/L] 105 ± 78 55 ± 11 12
  Ammonia-N[mg/L]   52 ± 11 16
Open in a new tab
a

Dissolved oxygen and pH were monitored at the same time as dissolved methane, while COD and ammonia were measured intermittently due to high variability and low change across the column. Influent COD data from Shin et al.4

Siloxanes

While hydrophilic trace organic contaminants remain in the aqueous phase through the methane stripping process, siloxane compounds are quite hydrophobic and removed by air stripping. Significant quantities of siloxanes were detected in both the SAF-MBR gas as well as the air-stripping produced gas. Commonly occurring siloxane compounds were measured in three batch samples of the SAF-MBR biogas and air-stripped methane gas. Results are shown with relevant siloxane chemical properties in Table 4.

Table 4. Siloxane Species Measured in Biogas and Stripped Off-Gas with Average and Maximum Values Detecteda.

        observed μg/m3(n = 3)
  characteristics
 SAF-MBR biogas
 stripped methane
siloxane name chemical formula MW KHb average max average max
trimethylsilanol (TMS) (CH3)3SiOH 90.2 0.00577d 53.3 ± 92.4 160 27.3 ± 47.3 82
hexamethyldisiloxane (L2) C6H18OSi2 162.38 184c 20.3 ± 35.2 61 N.D.  
hexamethylcyclotrisiloxane (D3) C6H18O3Si3 222.46 2.62c 22.7 ± 39.3 68 N.D.  
octamethyltrisiloxane (L3) C8H24O2Si3 236.53 1175d 20 ± 34.6 60 N.D.  
octamethylcyclotetrasiloxane (D4) C8H24O4Si4 296.61 490d 203 ± 286 530 64.7 ± 57.5 110
decamethyltetrasiloxane (L4) C10H30O3Si4 310.68 2818d N.D. N.D.    
decamethylcyclopentasiloxane (D5) C10H30O5Si5 370.77 1350d 213 ± 370 640 60 ± 84.9 120
dodecamethylpentasiloxane (L5) C12H36O4Si5 384.84 4640e N.D. N.D.    
dodecamethylcyclohexasiloxane (D6) C12H36O6Si6 444.92 1020d 207 ± 247 480 68 ± 61.6 120
total converted to μg/m3 silicon 277 ± 388 720 76 ± 70.8 140
Open in a new tab
a

Full data available in S.6.2.

b

Dimensionless air/water partition at 20–25 °C.

c

ChemFinder, ChemIDplus Advanced, HSDB 2005, TemaNord 2005.

d

Xu and Kropscott, 2014.46

e

Mazzoni, Roy, and Grigoras, 1997, pg. 53–81.47

While the SAF-MBR biogas has higher observed concentrations of siloxanes, the relative contribution per unit of water treated is only 16 ± 12% from the SAF-MBR biogas. This is due to the lower volumetric flow of biogas relative to air stripping gas: biogas has a flow ratio to process water of 0.1 compared to 1.7 for the air stripping unit.

Discussion

Dissolved Methane Recovery via Air Stripping

While several environmental variables changed between the testing phases in this study, the air stripping operation and methane removals were robust. Seasonal impacts should reduce methane recovery at low water temperatures due to increasing methane solubility, as well as reduced biological activity. However, even with a reduction in temperature during phase 2 with daily disinfection, the removal of methane remained greater than 98% and evidence of microbial activity was reduced but not eliminated. The upper portion of the air stripping column remained clear, but the lower portion accumulated biomass as seen in photos in the Supporting Information S6. A reduced but significant difference between methane removal and recovery in off-gas persisted with daily disinfection. However, the loss of some methane to oxidation may increase the overall removal rate and could provide ancillary benefits for trace contaminant removal. Large-scale methane recovery in air stripping columns will likely require further optimization of disinfectant dosing and the addition of physical cleaning of media. Increased hydraulic loading in the column could increase the impact of biomass accumulation on column flooding and gas pressure requirements.

The choice of an air/water ratio for methane recovery was chosen to maintain the off-gas methane concentration at a safety factor of two below the lower explosion limit in oxygen-rich air of 5% by volume.48 The selected air/water ratio of 1.5 did not impact the modeled or observed methane recovery. Due to methane’s high Henry’s constant, any air/water ratio greater than 1 yields negligible improvement to recovery. Neither does this safety factor limit the recovery of all gaseous methane via blending with primary digestor biogas without exceeding stoichiometric combustion, which occurs above an air/water ratio of 2.1.

Energy Consumption and Greenhouse Impact Comparison

Energy costs are a minor contribution to greenhouse impacts among the processes considered for methane recovery. Air stripping is the most energy efficient process, though the inhibition of methane oxidizing microorganisms requires additional energy and environmental impacts from chemical cleaning. The largest factor for greenhouse impact of dissolved methane management is the methane removal efficiency, and the assumption that removed methane can be consumed in a way that prevents its release to the atmosphere. In all cases, the recovery of methane has a positive energy balance when utilized in a combined cycle process, and the energy balance is even greater with CHP, for which efficiencies between 80 and 90% are common.49 Air stripping appears to provide the most energy and climate-optimal recovery of dissolved methane from anaerobic membrane bioreactors (AnMBRs), but only under set utilization assumptions. Without on-site methane combustion and primary solids digestion, alternative methane separation and utilization technologies should be considered, as the stoichiometric limitations of blended combustion gas would require releasing a fraction of dilute methane off-gas to the atmosphere. Additionally, continued advancements in membrane exchangers can increase methane recovery up to 98% and may improve the favorability of these technologies, which still must consider the impacts of required physical and chemical cleaning.8

Situations where on-site combustion is infeasible may find the methane-rich, oxygen-free gas streams from hollow fiber membrane or vacuum degasifier separation easier to upgrade for alternative beneficial uses, such as biological feedstock or resale to gas utility lines. Upgrading biogas can be energy-intensive, particularly when separation of carbon dioxide is required as is the case for addition to natural gas lines.19

Trace Contaminants and Siloxanes

The removal of 1,4-dioxane is greater than expected based on volatilization alone, as air stripping has not been observed to be an effective removal mechanism for this highly soluble compound.40 Microbial degradation is a likely explanation for the observed removal, as the transition from anaerobic to aerobic conditions favors colonization by microbes with aerobic metabolisms. Monooxygenase-utilizing organisms such as methanotrophs have been observed to degrade 1,4-dioxane,15 as well as benzotriazole and sulfamethoxazole.14 A factor that may explain the comparatively higher removal of 1,4-dioxane could be the greater spike concentration utilized to account for the analytical method. The increased concentration of 1,4-dioxane in the water would be degraded faster by any microbes present in the column at the time of testing. If energy recovery is not feasible, biological oxidation of methane downstream of mainstream AnMBRs could provide additional benefits for trace organic contaminant removal while offsetting greenhouse emissions. Further investigation into the microbes and metabolisms present could determine if degradation were possible downstream of methane recovery or if co-metabolism of methane and trace organics by mono-oxygenase enzymes is achievable.

While the measured values of siloxanes at the pilot facility were highly variable, the presence of siloxanes in both the SAF-MBR biogas and air stripped gas have substantial implications for net energy recovery from AnMBRs. Acceptable values of siloxanes in gas combustion range from 0.3 mg Si/m3 in microturbines up to 2.5–5 mg Si/m3 for large turbines.50 Blending the biogas with off-gas from the air stripper will increase the effective loading of siloxanes to a minimum of 1.6 mg Si/m3, based on calculations shown in S.6.3 assuming no accumulation of siloxanes in primary biogas beyond those observed in SAF-MBR biogas. Siloxanes are likely problematic for vacuum degassing of methane as well but may be removed during hollow fiber filtration depending on membrane selectivity.

Abatement solutions for the additional loading of air stripped siloxanes, as well as water vapor or hydrogen sulfides where present, may be needed and could include those currently used for biogas purification. However, given the low concentration of methane and potential use of off-gas for oxygen supply to a turbine or furnace, the removal of carbon dioxide, nitrogen, or oxygen are unnecessary. This consideration narrows the options for air stripping gas polishing to those that specifically target the constituents of concern at the lowest energy per volume gas, as the gas volume to be treated is an order of magnitude greater than the biogas produced for a plant of similar size. The suitable treatments for air-stripper off-gas in this case are sorption technologies, utilizing either activated carbon, biochar, silica gel, zeolites, or other inorganic sorbents.18,51,52 Costs of activated carbon and poor thermal recovery from siloxane loading make inexpensive biochar or more robust silica gels or zeolite promising options, with zeolites being particularly appealing due to their compatibility with high moisture gases.52

Conclusions

Extended pilot testing of air stripping recovery of SAF-MBR dissolved methane indicates that counter flow mass transfer over randomly packed media is an energy efficient and robust technology for methane removal and onsite utilization of the air-stripped methane for heat and electricity generation. Frequent disinfection of the air stripping unit is required to prevent methanotrophs from severely hindering methane recovery, and an air water ratio of 1.5–2 is required to prevent effluent gas from exceeding 50% of the lower explosive limit when mixed with ambient air. Possible evidence of trace contaminant removal via monooxygenase co-metabolism with methane was observed, but such removal is limited under optimal methane recovery conditions. Biogas polishing for siloxanes, water vapor, and possibly hydrogen sulfide is likely required for most wastewater sources and the protection of downstream combustion units. Air stripping provides excellent removal of dissolved methane, using the energy equivalent of less than 1% of recoverable electricity. Air stripping reduces possible methane fugitive emissions from treated wastewater. Other methane recovery technologies using vacuum degasifiers or membranes have positive energy balances but have lower methane removal. While methane recovery from AnMBRs is clearly needed from an energy and greenhouse perspective, the choice of technology will depend on compatibility with operational constraints to maximize the utility of the recovered methane.

The future of a built-out system would include a micro-screen for primary treatment and digestion of solids, followed by the staged anaerobic fluidized bed membrane bioreactor (SAF-MBR) for biogas methane. Effluent from this anaerobic process contains dissolved methane, sulfide, and ammonia, which need to be treated before water reuse. The methane can be recovered by air stripping as described in this study. Residual COD and sulfide can be then aerobically removed by heterotrophic bacteria and sulfide-oxidizing bacteria. Effluent polishing for trace organic contaminants could include biochar or regenerated activated carbon filtration with UV disinfection for non-potable reuse with ammonia fertilizer. This process train will be evaluated in follow-on work, and results from this and future studies will expand the available options for incorporating mainstream anaerobic reactors rapidly and sustainably into existing wastewater treatment and reuse trains.

Acknowledgments

Funding for this study came from the Bill and Cloy Codiga Family, the Singapore Public Utility Board, the NSF Engineering Research Center for Re-inventing the Nation’s Urban Water Infrastructure (ReNUWIt, NSF ERC 1028968), and a Stanford Graduate Fellowship which supported Stephen Galdi. The assistance of Sebastien Tilmans, Zhizheng Qui, and Masoom Desai is acknowledged.

Supporting Information Available

The Supporting Information is available free of charge at https://pubs.acs.org/doi/10.1021/acsestengg.2c00256.

  • Experiment photos, model equations, analytical instrument settings, detection limits, and additional contaminant data (PDF)

The authors declare no competing financial interest.

Supplementary Material

ee2c00256_si_001.pdf (1.1MB, pdf)

References

  1. McCarty P. L.; Bae J.; Kim J. Domestic Wastewater Treatment as a Net Energy Producer-Can This be Achieved?. Environ. Sci. Technol. 2011, 45, 7100–7106. 10.1021/es2014264. [DOI] [PubMed] [Google Scholar]
  2. Gao H.; Scherson Y. D.; Wells G. F. Towards Energy Neutral Wastewater Treatment: Methodology and State of the Art. Environ. Sci.: Processes Impacts 2014, 16, 1223–1246. 10.1039/C4EM00069B. [DOI] [PubMed] [Google Scholar]
  3. Kim J.; Kim K.; Ye H.; Lee E.; Shin C.; McCarty P. L.; Bae J. Anaerobic Fluidized Bed Membrane Bioreactor for Wastewater Treatment. Environ. Sci. Technol. 2011, 45, 576–581. 10.1021/es1027103. [DOI] [PubMed] [Google Scholar]
  4. Shin C.; Tilmans S. H.; Chen F.; McCarty P. L.; Criddle C. S. Temperate Climate Energy-Positive Anaerobic Secondary Treatment of Domestic Wastewater at Pilot-Scale. Water Res. 2021, 204, 117598. 10.1016/j.watres.2021.117598. [DOI] [PubMed] [Google Scholar]
  5. McCurry D. L.; Bear S. E.; Bae J.; Sedlak D. L.; McCarty P. L.; Mitch W. A. Superior Removal of Disinfection Byproduct Precursors and Pharmaceuticals from Wastewater in a Staged Anaerobic Fluidized Membrane Bioreactor Compared to Activated Sludge. Environ. Sci. Technol. Lett. 2014, 1, 459–464. 10.1021/ez500279a. [DOI] [Google Scholar]
  6. Yeo H.; An J.; Reid R.; Rittmann B. E.; Lee H.-S. Contribution of Liquid/Gas Mass-Transfer Limitations to Dissolved Methane Oversaturation in Anaerobic Treatment of Dilute Wastewater. Environ. Sci. Technol. 2015, 49, 10366–10372. 10.1021/acs.est.5b02560. [DOI] [PubMed] [Google Scholar]
  7. Shin C.; McCarty P. L.; Bae J. Importance of Dissolved Methane Management When Anaerobically Treating Low-Strength Wastewaters. Curr. Org. Chem. 2016, 20, 2810–2816. 10.2174/1385272820666160517155831. [DOI] [Google Scholar]
  8. Centeno Mora E.; Chernicharo C. d. L. Use of Membrane Contactors for Removing and Recovering Dissolved Methane from Anaerobic Reactors Effluents: State-of-the-Art, Challenges, and Perspectives. Rev. Environ. Sci. Biotechnol. 2020, 19, 673–697. 10.1007/s11157-020-09546-w. [DOI] [Google Scholar]
  9. Velasco P.; Jegatheesan V.; Othman M. Recovery of Dissolved Methane From Anaerobic Membrane Bioreactor Using Degassing Membrane Contactors. Front. Environ. Sci. 2018, 6, 151. 10.3389/fenvs.2018.00151. [DOI] [Google Scholar]
  10. Crone B. C.; Sorial G. A.; Pressman J. G.; Ryu H.; Keely S. P.; Brinkman N.; Bennett-Stamper C.; Garland J. L. Design and Evaluation of Degassed Anaerobic Membrane Biofilm Reactors for Improved Methane Recovery. Bioresour. Technol. Rep. 2020, 10, 100407. 10.1016/j.biteb.2020.100407. [DOI] [PMC free article] [PubMed] [Google Scholar]
  11. Lee E.; Rout P. R.; Kyun Y.; Bae J. Process Optimization and Energy Analysis of Vacuum Degasifier Systems for the Simultaneous Removal of Dissolved Methane and Hydrogen Sulfide from Anaerobically Treated Wastewater. Water Res. 2020, 182, 115965. 10.1016/j.watres.2020.115965. [DOI] [PubMed] [Google Scholar]
  12. Huete A.; de los Cobos-Vasconcelos D.; Gómez-Borraz T.; Morgan-Sagastume J. M.; Noyola A. Control of dissolved CH4 in a municipal UASB reactor effluent by means of a desorption - Biofiltration arrangement. J. Environ. Manage. 2018, 216, 383–391. 10.1016/j.jenvman.2017.06.061. [DOI] [PubMed] [Google Scholar]
  13. Matsuura N.; Hatamoto M.; Sumino H.; Syutsubo K.; Yamaguchi T.; Ohashi A. Closed DHS System to Prevent Dissolved Methane Emissions as Greenhouse Gas in Anaerobic Wastewater Treatment by Its Recovery and Biological Oxidation. Water Sci. Technol. 2010, 61, 2407–2415. 10.2166/wst.2010.219. [DOI] [PubMed] [Google Scholar]
  14. Benner J.; De Smet D.; Ho A.; Kerckhof F.-M.; Vanhaecke L.; Heylen K.; Boon N. Exploring Methane-Oxidizing Communities for the Co-Metabolic Degradation of Organic Micropollutants. Appl. Microbiol. Biotechnol. 2015, 99, 3609–3618. 10.1007/s00253-014-6226-1. [DOI] [PubMed] [Google Scholar]
  15. Mahendra S.; Alvarez-Cohen L. Kinetics of 1,4-Dioxane Biodegradation by Monooxygenase-Expressing Bacteria. Environ. Sci. Technol. 2006, 40, 5435–5442. 10.1021/es060714v. [DOI] [PubMed] [Google Scholar]
  16. Kochetkov A.; Smith J. S.; Ravikrishna R.; Valsaraj K. T.; Thibodeaux L. J. Air-Water Partition Constants for Volatile Methyl Siloxanes. Environ. Toxicol. Chem. 2001, 20, 2184–2188. 10.1002/etc.5620201008. [DOI] [PubMed] [Google Scholar]
  17. Dewil R.; Appels L.; Baeyens J. Energy Use of Biogas Hampered by the Presence of Siloxanes. Energy Convers. Manage. 2006, 47, 1711–1722. 10.1016/j.enconman.2005.10.016. [DOI] [Google Scholar]
  18. Nyamukamba P.; Mukumba P.; Chikukwa E. S.; Makaka G. Biogas Upgrading Approaches with Special Focus on Siloxane Removal-A Review. Energies 2020, 13, 6088. 10.3390/en13226088. [DOI] [Google Scholar]
  19. Khan M. U.; Lee J. T. E.; Bashir M. A.; Dissanayake P. D.; Ok Y. S.; Tong Y. W.; Shariati M. A.; Wu S.; Ahring B. K. Current Status of Biogas Upgrading for Direct Biomethane Use: A Review. Renewable Sustainable Energy Rev. 2021, 149, 111343. 10.1016/j.rser.2021.111343. [DOI] [Google Scholar]
  20. Shin C.; Szczuka A.; Liu M. J.; Mendoza L.; Jiang R.; Tilmans S. H.; Tarpeh W. A.; Mitch W. A.; Criddle C. S. Recovery of Clean Water and Ammonia from Domestic Wastewater: Impacts on Embodied Energy and Greenhouse Gas Emissions. Environ. Sci. Technol. 2022, 56, 8712–8721. 10.1021/acs.est.1c07992. [DOI] [PubMed] [Google Scholar]
  21. Habib R.; Asif M. B.; Iftekhar S.; Khan Z.; Gurung K.; Srivastava V.; Sillanpää M. Influence of Relaxation Modes on Membrane Fouling in Submerged Membrane Bioreactor for Domestic Wastewater Treatment. Chemosphere 2017, 181, 19–25. 10.1016/j.chemosphere.2017.04.048. [DOI] [PubMed] [Google Scholar]
  22. Onda K.; Takeuchi H.; Okumoto Y. Mass Transfer Coefficients Between Gas And Liquid Phases In Packed Columns. J. Chem. Eng. Jpn. 1968, 1, 56–62. 10.1252/jcej.1.56. [DOI] [Google Scholar]
  23. Roberts P. V.; Hopkins G. D.; Munz C.; Riojas A. H. Evaluating Two-Resistance Models for Air Stripping of Volatile Organic Contaminants in a Countercurrent, Packed Column. Environ. Sci. Technol. 1985, 19, 164–173. 10.1021/es00132a010. [DOI] [Google Scholar]
  24. Crittenden J. C.; Trussell R. R.; Hand D. W.; Howe K. J.; Tchobanoglous G.. MWH’s Water Treatment Principles and Design, 3rd ed.; John Wiley & Sons, Inc.: Hoboken, NJ, USA, 2012. [Google Scholar]
  25. Alberto M. C. R.; Arah J. R. M.; Neue H. U.; Wassmann R.; Lantin R. S.; Aduna J. B.; Bronson K. F. A Sampling Technique for the Determination of Dissolved Methane in Soil Solution. Chemosphere: Global Change Sci. 2000, 2, 57–63. 10.1016/S1465-9972(99)00044-6. [DOI] [Google Scholar]
  26. Kayode Coker A.Physical Properties of Liquids and Gases. Ludwig’s Applied Process Design for Chemical and Petrochemical Plants; Elsevier, 2010; pp 757–792. [Google Scholar]
  27. Shin C.; Bae J. Current Status of the Pilot-Scale Anaerobic Membrane Bioreactor Treatments of Domestic Wastewaters: A Critical Review. Bioresour. Technol. 2018, 247, 1038–1046. 10.1016/j.biortech.2017.09.002. [DOI] [PubMed] [Google Scholar]
  28. Perry’s Chemical Engineers’ Handbook, 7th ed.; Perry R. H., Green D. W., Maloney J. O., Eds.; McGraw-Hill: New York, 1997. [Google Scholar]
  29. Myhre G.; Shindell D.; Bréon F.-M.; Collins W.; Fuglestvedt J.; Huang J.; Koch D.; Lamarque J.-F.; Lee D.; Mendoza B.; Nakajima T.; Robock A.; Stephens G.; Zhang H.; Aamaas B.; Boucher O.; Dalsøren S. B.; Daniel J. S.; Forster P.; Granier C.; Haigh J.; Hodnebrog Ø.; Kaplan J. O.; Marston G.; Nielsen C. J.; O’Neill B. C.; Peters G. P.; Pongratz J.; Ramaswamy V.; Roth R.; Rotstayn L.; Smith S. J.; Stevenson D.; Vernier J.-P.; Wild O.; Young P.; Jacob D.; Ravishankara A. R.; Shine K.. Anthropogenic and Natural Radiative Forcing. Climate Change 2013: The Physical Science Basis. Contribution of Working Group I to the Fifth Assessment Report of the Intergovernmental Panel on Climate Change; Cambridge University Press: Cambridge, UK and New York, NY, USA, 2013.
  30. Alvarez-Gaitan J. P.; Peters G. M.; Rowley H. V.; Moore S.; Short M. D. A Hybrid Life Cycle Assessment of Water Treatment Chemicals: An Australian Experience. Int. J. Life Cycle Assess. 2013, 18, 1291–1301. 10.1007/s11367-013-0574-4. [DOI] [Google Scholar]
  31. EIA—State Electricity Profiles. U.S. Energy Information Administration. https://www.eia.gov/electricity/state/unitedstates/ (accessed April 20, 2022).
  32. Breeze P.Greenhouse Gas Emissions and Power Generation. Electricity Generation and the Environment; Elsevier, 2017; pp 23–31. [Google Scholar]
  33. Szczuka A.; Berglund-Brown J. P.; Chen H. K.; Quay A. N.; Mitch W. A. Evaluation of a Pilot Anaerobic Secondary Effluent for Potable Reuse: Impact of Different Disinfection Schemes on Organic Fouling of RO Membranes and DBP Formation. Environ. Sci. Technol. 2019, 53, 3166–3176. 10.1021/acs.est.8b05473. [DOI] [PubMed] [Google Scholar]
  34. Margot J.; Rossi L.; Barry D. A.; Holliger C. A review of the fate of micropollutants in wastewater treatment plants. Wiley Interdiscip. Rev.: Water 2015, 2, 457–487. 10.1002/wat2.1090. [DOI] [Google Scholar]
  35. Zhang Z.; Chuang Y.-H.; Szczuka A.; Ishida K. P.; Roback S.; Plumlee M. H.; Mitch W. A. Pilot-Scale Evaluation of Oxidant Speciation, 1,4-Dioxane Degradation and Disinfection Byproduct Formation during UV/Hydrogen Peroxide, UV/Free Chlorine and UV/Chloramines Advanced Oxidation Process Treatment for Potable Reuse. Water Res. 2019, 164, 114939. 10.1016/j.watres.2019.114939. [DOI] [PubMed] [Google Scholar]
  36. Szczuka A.; Huang N.; MacDonald J. A.; Nayak A.; Zhang Z.; Mitch W. A. N-Nitrosodimethylamine Formation during UV/Hydrogen Peroxide and UV/Chlorine Advanced Oxidation Process Treatment Following Reverse Osmosis for Potable Reuse. Environ. Sci. Technol. 2020, 54, 15465–15475. 10.1021/acs.est.0c05704. [DOI] [PubMed] [Google Scholar]
  37. Crone B. C.; Garland J. L.; Sorial G. A.; Vane L. M. Significance of Dissolved Methane in Effluents of Anaerobically Treated Low Strength Wastewater and Potential for Recovery as an Energy Product: A Review. Water Res. 2016, 104, 520–531. 10.1016/j.watres.2016.08.019. [DOI] [PubMed] [Google Scholar]
  38. Evans P. J.; Parameswaran P.; Lim K.; Bae J.; Shin C.; Ho J.; McCarty P. L. A Comparative Pilot-Scale Evaluation of Gas-Sparged and Granular Activated Carbon-Fluidized Anaerobic Membrane Bioreactors for Domestic Wastewater Treatment. Bioresour. Technol. 2019, 288, 120949. 10.1016/j.biortech.2019.01.072. [DOI] [PubMed] [Google Scholar]
  39. Rongwong W.; Goh K.; Bae T.-H. Energy Analysis and Optimization of Hollow Fiber Membrane Contactors for Recovery of Dissolve Methane from Anaerobic Membrane Bioreactor Effluent. J. Membr. Sci. 2018, 554, 184–194. 10.1016/j.memsci.2018.03.002. [DOI] [Google Scholar]
  40. Odah M. M.; Powell R.; Riddle D. J. ART In-Well Technology Proves Effective in Treating 1,4-Dioxane Contamination. Remediat. J. 2005, 15, 51–64. 10.1002/rem.20049. [DOI] [Google Scholar]
  41. Ondo D.; Dohnal V. Temperature dependence of limiting activity coefficients and Henry’s law constants of cyclic and open-chain ethers in water. Fluid Phase Equilib. 2007, 262, 121–136. 10.1016/j.fluid.2007.08.013. [DOI] [Google Scholar]
  42. US EPA . Estimation Program Interface (EPI) Suite for Microsoft Windows, V 4.11; United States Environmental Protection Agency: Washington DC, USA, 2012. [Google Scholar]
  43. Haruta S.; Jiao W.; Chen W.; Chang A. C.; Gan J. Evaluating Henry’s law constant of N-Nitrosodimethylamine (NDMA). Water Sci. Technol. 2011, 64, 1636–1641. 10.2166/wst.2011.742. [DOI] [PubMed] [Google Scholar]
  44. Sander R. Compilation of Henry’s law constants (version 4.0) for water as solvent. Atmos. Chem. Phys. 2015, 15, 4399–4981. 10.5194/acp-15-4399-2015. [DOI] [Google Scholar]
  45. Kehrer A.Die Wirkung von Pharmaka und Pestiziden einzeln und in Kombination auf die Embryonalentwicklung des Zebrabärblings (Danio rerio). Ph.D. Dissertation, Technische Universität Dresden, Dresden, Germany, 2008. [Google Scholar]
  46. Xu S.; Kropscott B. Evaluation of the three-phase equilibrium method for measuring temperature dependence of internally consistent partition coefficients (K OW , K OA , and K AW ) for volatile methylsiloxanes and trimethylsilanol. Environ. Toxicol. Chem. 2014, 33, 2702–2710. 10.1002/etc.2754. [DOI] [PMC free article] [PubMed] [Google Scholar]
  47. Mazzoni S. M.; Roy S.; Grigoras S.. Chapter 3 Eco-Relevant Properties of Selected Organosilicon Materials. In Organosilicon Materials; Chandra G., Ed.; The Handbook of Environmental Chemistry; Springer Berlin Heidelberg: Berlin, Heidelberg, 1997; Vol. 3/3H. [Google Scholar]
  48. Zabetakis M. G.Flammability Characteristics of Combustible Gases and Vapors, 1924.
  49. Lee S.; Esfahani I. J.; Ifaei P.; Moya W.; Yoo C. Thermo-Environ-Economic Modeling and Optimization of an Integrated Wastewater Treatment Plant with a Combined Heat and Power Generation System. Energy Convers. Manage. 2017, 142, 385–401. 10.1016/j.enconman.2017.03.060. [DOI] [Google Scholar]
  50. Pierce J. L.Siloxane Sampling, Analysis and Data Reporting Recommendations on Standardization for the Biogas Utilization Industry. Proceedings of the 14th Annual EPA LMOP Conference and Project Expo, Baltimore, MD, January 18–20 2011; United States Environmental Protection Agency: Washington DC, USA, 2011.
  51. Shen M.; Zhang Y.; Hu D.; Fan J.; Zeng G. A Review on Removal of Siloxanes from Biogas: With a Special Focus on Volatile Methylsiloxanes. Environ. Sci. Pollut. Res. 2018, 25, 30847–30862. 10.1007/s11356-018-3000-4. [DOI] [PubMed] [Google Scholar]
  52. Gaj K. Adsorptive Biogas Purification from Siloxanes-A Critical Review. Energies 2020, 13, 2605. 10.3390/en13102605. [DOI] [Google Scholar]

Associated Data

This section collects any data citations, data availability statements, or supplementary materials included in this article.

Supplementary Materials

ee2c00256_si_001.pdf (1.1MB, pdf)

Articles from ACS Es&t Engineering are provided here courtesy of American Chemical Society

RESOURCES