Comparative Life Cycle and Technoeconomic Assessment for Energy Recovery from Dilute Wastewater

Abstract

Anaerobic treatment offers a sustainable alternative to aerobic treatment by recovering energy from wastewater. Anaerobic treatment, however, is challenged by reduced performance at lower temperatures and dissolved methane in the effluent, which represents a loss in recoverable energy and a potent greenhouse gas emission. Long-term operation of a bench-scale anaerobic baffled reactor (ABR) at 15–20°C, with domestic wastewater, provided data to evaluate life cycle environmental and economic performances of mainstream anaerobic treatment. Performance of the ABR was compared to a conceptual design of treatment with trickling filter+anaerobic digestion (TF+AD). The ABR and TF were modeled as example low-cost anaerobic and aerobic treatments, respectively, because these technologies may be more easily implemented (compared to membrane systems and activated sludge) in developing countries. This type of comprehensive study, not conducted previously for the ABR, is imperative for selecting appropriate technologies as we determine how to bring sanitation to billions who are unserved. The ABR recovered approximately six times more energy from low-strength wastewater than TF+AD and resulted in significantly more beneficial life cycle impacts for ecosystem quality and human health. Furthermore, technoeconomic analysis showed that the ABR has a life cycle cost that is about 40% lower than TF+AD. Dissolved methane in the effluent of the ABR, however, resulted in a harmful impact on climate change, whereas TF+AD resulted in approximately neutral impacts on climate change. A combined ABR+TF assembly, which was assumed to oxidize residual organics and dissolved methane, resulted in beneficial environmental impacts for four environmental categories. Advancements in synergistic technologies that remove or recover dissolved methane in anaerobic effluent would allow implementation of the ABR without increasing climate change impacts.

Introduction

In 2010, ∼80% of the world's wastewater was untreated (Tauseef et al., 2013), and this poses a significant threat to human health and the environment. A majority of populations that lack proper sanitation reside in the developing world where implementation of conventional wastewater treatment (e.g., activated sludge) is hindered by high costs and the lack of reliable electricity. Approximately 1.5 billion people living in urban centers discharge wastewater through collection systems with no treatment (Baum et al., 2013). Furthermore, although >50% of households in the upper middle-income countries (e.g., India) are connected to sewers, over 85% of this sewage is not treated (Baum et al., 2013). Central India and Pakistan represent regions that are served by sewers, but lack wastewater treatment with average wastewater temperatures as low as 15°C (Sato et al., 2006; Shahzad et al., 2015). In addition, developing countries that implement conventional wastewater treatment allocate almost half of their municipal budgets to water management (James et al., 2002). In response to these challenges, the U.N. Sustainable Development Goals 6 and 7 aim “to achieve adequate and equitable access to sanitation; to improve water quality by reducing pollution, halving the proportion of untreated wastewater; and to provide access to affordable and sustainable energy” (United Nations General Assembly, 2015). Low-cost wastewater treatment technologies that recover energy at relatively low temperatures and do not rely on grid electricity may help reach these sustainable development targets.

Electricity requirements for aeration in conventional activated sludge treatment of domestic wastewater contributes to high costs and subsequent greenhouse gas emissions. Considering the significant electricity demand, carbon footprint, and biosolids production (Smith et al., 2014), the export of activated sludge technology to developing countries should be carefully evaluated. By contrast, anaerobic treatment requires no aeration and produces methane that can be converted to electricity and heat. In addition, anaerobic systems may achieve effluent requirements typical of developed countries (Agrawal et al., 1997; Shoener et al., 2014), when treating dilute wastewater at relatively low temperatures (e.g., 15–20°C). Given the significant lack of domestic wastewater treatment in developing countries, emphasis on removal of the majority of pollutants with processes that offer low capital and operating costs, minimal energy input, and low biosolids production seems more critical in underserved regions than achievement of stringent biochemical oxygen demand (BOD) and total suspended solids (TSS) effluent quality (Mara, 2004).

The anaerobic baffled reactor (ABR)—which converts organic waste to methane—has been proposed because of its stable performance, simple design, and low capital and operating costs (Barber and Stuckey, 1999; Gutterer et al., 2010; Tilley et al., 2014; Reynaud and Buckley, 2015). The ABR consists of multiple compartments separated by vertical baffles and it behaves like a series of upflow anaerobic sludge blankets (UASBs). This compartmentalized design decouples solids retention time and hydraulic retention time (HRT), promoting the retention of slow growing anaerobes, minimizes reactor volume, promotes resistance to hydraulic and organic shocks, and prevents sludge expansion—a common problem with UASBs (Nachaiyasit and Stuckey, 1997; Manariotis and Grigoropoulos, 2002; Gopala Krishna et al., 2008). In addition, the ABR may release H₂ in the first chamber, whereas the UASB would consume this H₂ as wastewater moved upward. Drawbacks of the ABR (as well as other anaerobic technologies) include reduced performance as temperature and feed concentration decrease, as well as high concentrations of dissolved methane trapped in the effluent (Liu et al., 2014; Smith et al., 2014; Shoener et al., 2016). Aerobic polishing processes have been proposed to reduce organics to levels below 30 mg/L as BOD₅ and to biologically oxidize dissolved methane (Hatamoto et al., 2010, 2011; van der Ha et al., 2011; Kalbar et al., 2013).

Although a number of studies examine treatment of dilute wastewater with the ABR, no studies have reported the effect of variable wastewater characteristics on laboratory-scale reactor performance (Reynaud and Buckley, 2015). In addition, the concentrations of dissolved methane in ABR effluent have not been quantified. Methane, a potent greenhouse gas, remaining in wastewater effluents is ultimately released to the atmosphere and should, therefore, be accounted for when assessing the environmental sustainability of treatment technologies.

Low-energy aerobic treatment with fixed film processes, such as a trickling filter (TF), combined with energy recovery through anaerobic digestion (AD) of biosolids provides an alternative option for wastewater treatment in underserved communities (von Sperling, 1996; Elmitwalli et al., 2004; Naz et al., 2015). A comparative analysis of the environmental impacts and economics for low-energy anaerobic and aerobic treatment technologies—which are appropriate for the developing world—will help stakeholders choose among appropriate treatment options.

Life cycle assessment (LCA) is a method used to estimate the environmental performance of a product or process over its entire lifetime, including extraction of raw materials, production, operation, and end of life (Guinee, 2002). LCA can assist in designing environmentally beneficial processes (e.g., wastewater treatment) by preventing burden shifting through the evaluation of trade-offs among different impact categories (e.g., climate change vs. ecosystem quality). Although numerous studies have used LCA to assess the environmental performance of wastewater treatment, about 60% of 45 published studies did not follow ISO standards (ISO, 2006a, 2006b; Corominas et al., 2013), making it difficult to compare among published results. In addition, most studies did not couple LCA with an economic assessment (Corominas et al., 2013), and few LCA studies have analyzed smaller scale treatment technologies (<5 million gallons per day [MGD]) that would be appropriate for the developing world (Cornejo et al., 2013). Similarly, dissolved methane trapped in anaerobic effluent was not considered in a number of LCA studies on anaerobic treatment technologies (Hospido et al., 2008; Foley et al., 2010; Kalbar et al., 2013), yet it represents a significant environmental impact (Smith et al., 2014; Shoener et al., 2016).

In this study, we compare the environmental and economic performances for treatment of dilute domestic wastewater at low temperatures (15–20°C) with ABR and TF+AD technologies. The ABR and TF were chosen as examples of anaerobic and aerobic treatment that may be implemented more easily (compared to membrane processes and activated sludge) in developing countries. Long-term laboratory data (for the ABR, using domestic wastewater with large variations in concentration of organics) was incorporated into a modeling framework that includes wastewater treatment process modeling, LCA coupled with uncertainty analysis, and technoeconomic analysis (TEA). Since the performance of the ABR may not produce high-quality effluent (i.e., with acceptable levels of organic matter), we evaluated two aerobic polishing processes—TF and constructed wetland (CW)—which have been used to treat ABR effluent or effluent from mainstream anaerobic treatment (Chernicharo and Nascimento, 2001; Singh et al., 2009). The use of long-term laboratory data for the ABR combined with the results from the modeling framework elucidates trade-offs for these four treatment assemblies. Although a few thorough systems modeling studies have been presented on the sustainability of the anaerobic membrane reactor (AnMBR) (Smith et al., 2014; Shoener et al., 2016), this is the first study that couples LCA, TEA, and uncertainty analysis for an ABR and compares it to an aerobic fixed film process.

Experimental Protocols

ABR description and protocols

A laboratory-scale ABR consisting of three chambers with a total empty bed volume of 17.8 L was used for data collection (Supplementary Fig. S1). The empty bed volumes for Chambers 1, 2, and 3 were 8.8, 5.0, and 4.0 L, respectively. Liquid flowed into each chamber through a vertical down-flow baffle with a 45° bend along the bottom edge and trickled out of the first and second chambers over a vertically oriented up-flow baffle. Effluent exited the reactor through polyvinyl chloride piping attached to an opening on the side of the third chamber.

The ABR was equipped with biosolids sampling ports on the bottom panel of each chamber and liquid sampling ports on the front and back panels of each chamber. A liquid sampling port was also located on effluent tubing between the reactor and a siphon break to obtain ABR effluent before atmospheric exposure. Tubing with gas sampling ports, equipped with natural rubber septa (Sigma-Aldrich), was attached to the center of the top panel of each chamber, and gas flow was measured with a tipping meter. The ABR temperature was controlled by a MM7 water chiller (VWR), which directed a flow of cooled water into six tubes oriented horizontally from the entrance to exit across the ABR.

The ABR was inoculated in 2008 with anaerobic granules from a full-scale UASB reactor that operated at 35°C (J.M. Smucker) and was operated continuously until May 2012. From May 2013 to April 2014, the reactor was operated continuously and fed with a synthetic wastewater feed. Starting May 2014, the reactor was fed raw domestic wastewater from a local wastewater treatment facility. Data collection for this study began after 14 days of operation with real wastewater. Steady-state reactor performance was assessed at 15°C for 56 days and 20°C for 43 days. Because performance of the reactor was not significantly different (p-value >0.1) at the two temperature regimens, performance data were averaged over the two temperature regimens.

Sodium bicarbonate (EMD Chemicals, Inc.) was added to the feed every other day to achieve alkalinity of 800 mg/L as CaCO₃ at 15°C to ensure that the pH within the reactor remained in the range of 6.5–7.5. The feed was continuously mixed with a paddle mixer (REX Engineering Co.) and fed into a side port of the ABR using a peristaltic Masterflex^® L/S^® pump (Model #7518-00; Cole-Parmer Instrument Company). A timer (XT Timer ChronTrol Corporation) controlled operating cycles for the pump and achieved a 0.55 day HRT. The organic loading rate varied due to daily differences in feed concentration (Supplementary Fig. S2).

Solids (5–10 g) were wasted from Chamber 1 twice weekly to reduce the buildup of nonbiodegradable solids in Chamber 1, which obstructed the liquid sampling port. Butyl rubber septum on sampling ports was replaced twice per week, and influent tubing was replaced as needed due to buildup of solids within the tubing.

Analytical methods

Total chemical oxygen demand (TCOD) and soluble COD (SCOD) of the influent and effluent were measured according to the standard closed reflux colorimetric method (APHA et al., 2012).

A 5-day BOD test was conducted on influent and effluent samples according to standard procedure 5210 (APHA et al., 2012), and BOD₅/TCOD ratios were calculated. A Hach HQ40d dual input portable multiparameter meter (Hach Company) was used to measure dissolved oxygen concentrations.

TSS and volatile suspended solids (VSS) of the influent and effluent were measured three times per week using standard methods (APHA et al., 2012). Total solids and volatile solids (VS) testing was conducted on solids wasted from Chamber 1 to estimate the composition of the wasted biosolids. In addition, an aliquot of the wasted biosolids was diluted and blended (with a kitchen blender) for COD analysis. A ratio of COD/VS was calculated.

A calibrated precision wet tip gas meter (Rebel Point Wet Tip Gas Meter Company) was used to quantify gas production by the ABR. Biogas composition was analyzed using a HP6890 GC coupled with a 15 ft × 1/8 in × 2.1 mm 60/80 Carboxen 1000 packed column (Supelco) and a thermal conductivity detector. Nitrogen served as the carrier gas, and the oven temperature was maintained at 150°C throughout a 7-min run. A 1 mL Hamilton^® SampleLock syringe (Hamilton Company) was used to take a 0.3 mL headspace sample from each chamber of the ABR. A standard curve generated using 99.99% methane gas (Associated Gas Products) and a 60% methane gas mixture (Praxair) was used to convert peak area to percent methane. Daily gaseous methane production was quantified by multiplying the total biogas production by the average percent methane in the headspace of the three ABR chambers.

Dissolved methane in ABR effluent was measured daily. The dissolved methane test was conducted by injecting 3 mL of ABR effluent into an 11.5 mL serum bottle sealed with a butyl rubber stopper and crimped aluminum cap. Samples were drawn directly from a sampling port attached to an extension on the third chamber, which minimized loss of methane to the atmosphere. The sealed serum bottle was shaken for 90 s to allow the system to equilibrate. Then, a 100 μL headspace sample was taken with a 100 μL Hamilton SampleLock syringe (Hamilton Company) and injected into a Hewlett Packard (HP) 5890A GC System coupled with a 2 m × 2 mm Rt-sulfur micropacked Silcosteel^® column (Restek Corporation) and an FID. Nitrogen carrier gas was used, and the oven temperature increased from 140°C to 230°C at a rate of 15°C per minute and then remained at 230°C for 2.33 min.

A standard curve created using 99.99% methane gas (Associated Gas Products) was used to convert peak area to moles of methane in the injected headspace sample. The partial pressure of methane in the headspace was used in conjunction with the ideal gas law, Henry's law, and a mass balance on the contents of the serum bottle to determine the total moles of methane in the serum bottle. Dimensionless Henry's constants (C_aq/C_g, where C is in units of mol/L) of 0.0356 and 0.0384 were calculated for 20°C and 15°C based on Equations (S-1) and (S-2), provided in the Supplementary Data (Tchobanoglous et al., 2003).

COD mass balance

A mass balance on COD was conducted over a 98-day period, during which the performance of the system was stable, as follows: COD was measured for the influent, effluent, and wasted solids. In addition, masses of gaseous methane produced and methane dissolved in the effluent were converted to units of mass of COD. To calculate the mass balance closure, the sum of all values (except for COD in the influent) was subtracted from the mass of COD in the influent. This difference was divided by the mass of COD in the influent.

Modeling Framework

System boundary and functional unit

LCA, conducted within the ISO 14040/14044 framework (Guinee, 2002; ISO, 2006a, 2006b), was used to compare the environmental impacts of an ABR and TF+AD. Since the effluent concentration of BOD₅ for the ABR was higher than 30 mg/L, we also evaluated two aerobic polishing processes—TF and CW—to treat the ABR effluent and reduce the BOD₅ to 30 mg/L. We assumed that the TF (when used as a post-ABR process) oxidized all of the dissolved methane present in the ABR effluent and that the CW removed no dissolved methane, all of which was eventually released to the atmosphere. Neither of these assumptions is absolutely correct, but they were modeled to illustrate the hypothetical difference between a process that removes BOD and dissolved methane (e.g., TF) and a process that removes BOD only (e.g., CW). In total, four treatment assemblies were evaluated: ABR, TF+AD, ABR+TF+AD, and ABR+CW. Process flow diagrams for the four assemblies are presented in Supplementary Figures S3–S6. Average operating parameters for the TF, AD, ABR, TF (as a post-ABR polishing process), and CW are presented in Supplementary Tables S1–S5.

The system was designed to treat wastewater with influent BOD₅ concentrations based on average ±95% confidence interval (CI) values determined during the experimental section of this article (334 ± 36 mg/L). Although this study focuses on technologies that may be appropriate for the developing world, the range of influent BOD₅ concentrations is more representative of the United States, due to the location of our laboratory study. In addition, this model only accounted for removal of COD and did not account for impacts related to the presence or removal of nutrients. All secondary treatment processes (ABR, TF, and CW) were assumed to be operated at 15–20°C. Disposal of biosolids was modeled using an aggregate of three processes: 13% landfilling, 62% land application, and 25% incineration (EPA, 1999). AD operation was assumed at 35°C with an HRT of 15 days, while biogas (from the ABR and AD) would be converted to usable energy using an on-site combined heat and power (CHP) system (EPA, 2007). Heat generated by the CHP would be used for digester heating, with excess heat wasted. We verified that environmental impacts associated with the construction phase of the treatment plant were negligible, as shown previously (Renou et al., 2008), and thus only accounted for the operation-phase impacts.

The functional unit was defined as the treatment of 2 MGD of domestic wastewater, assuming a plant lifetime of 30 years. The unit processes included within the system boundary are presented in Fig. 1.

FIG. 1.

System boundary for ABR and TF assemblies. ABR, anaerobic baffled reactor; TF, trickling filter.

System design

Life cycle inventories (LCIs), constructed within the Simapro 8.1 software package, were created for the ABR by scaling up laboratory data and were created for the TF, anaerobic digester, CHP system, and CW using engineering design equations. LCI data that account for off-site emissions (e.g., grid electricity) were obtained from the ecoinvent 3.2 database (Weidema et al., 2013), except for acrylonitrile and quicklime, which were obtained from the U.S. LCI database (Norris, 2004). To represent impacts in a developing country, the electricity grid was assumed to represent the average electricity grid in India: 65% hard coal, 10% natural gas, 9% petroleum, 14% hydropower, 2% nuclear, and 1% imported (Itten and Frischknecht, 2012; Weidema et al., 2013). Biogenic methane emissions were assumed to have a global warming potential (GWP) of 24 CO₂ equivalents (eq.) integrated over a 100-year time frame, and biogenic CO₂ emissions were assumed to have a GWP of zero (IPCC, 2013; Myhre et al., 2013). The inventories for the four treatment assemblies—ABR, TF, ABR+TF, and ABR+CW—are presented in Supplementary Table S6.

Impact assessment

IMPACT2002+, used for life cycle impact assessment, converts elementary flows to impacts on four damage categories: climate change (kg CO₂-eq.), human health (disability adjusted life years [DALYs]), resource depletion (megajoule [MJ] primary energy), and ecosystem quality (potentially disappeared fraction of species·m²·years) (Jolliet et al., 2003). In addition, normalization factors from IMPACT2002+ (Jolliet et al., 2003) were applied to convert the damage categories to units of LCA “points,” which allow a direct comparison of the four impact categories. One point represents one person—year of impacts in Europe (Jolliet et al., 2003). Although the context of the present study is not Europe, we used the normalization factors to allow readers to compare impacts. Future work should update these results when global normalization factors are published, such as those being developed as part of the IMPACT World method (Roy et al., 2013).

Uncertainty analysis

Uncertainty analysis was conducted as described in Sills et al. (2013). Briefly, model parameters presented in Supplementary Table S7 were input as probability distribution functions to represent parameter variability and uncertainty. Triangular and uniform distributions were fit to parameters with scarce data. Normal or lognormal distributions were fit to parameters derived from our laboratory study using the Distribution Fitting Toolbox in MATLAB (2013; Mathworks). One thousand Monte Carlo simulations were used to propagate the uncertainty of the parameters, as well as the uncertainty of the ecoinvent processes used to calculate off-site emissions (e.g., grid electricity).

Technoeconomic assessment

TEA was conducted for ABR, TF+AD, ABR+TF, and ABR+CW using the same system process design as for the LCA. Design criteria used to estimate capital and operating costs are presented in Supplementary Table S8. CAPDETWorks (Hydromantis, Inc.) was used to estimate capital and operating costs (data presented in Supplementary Tables S11 and S12, respectively), except for the CHP system and CW, which were calculated separately. Life cycle costs, presented as net present values (NPVs), were calculated with two discount rates (3.375% and 8%), assuming a plant lifetime of 30 years. Estimates of capital and operating costs are reported as 2015 U.S. dollars, and life cycle costs (as NPV) were used to compare among the four treatment assemblies, and not to predict costs in a developing country.

Subtotal capital costs for a CHP microturbine system were determined by assuming a value of $2,689 per kW (EPA, 2015). Additional fees listed in Supplementary Table S9 were added to calculate a total capital cost for the microturbine. Annual maintenance costs for cogeneration were calculated assuming a typical value of $0.01/kWh of electricity generated by the system (EPA, 2015). Annual electricity costs for gas cleanup to remove H₂S from biogas were also accounted for by assuming 0.25 kWh/kg VS processed (0.41 kWh/m³ CH₄ produced) for AD (Sills et al., 2013). The ABR, TF, ABR+TF, and ABR+CW assemblies produced 56, 43, 71, and 56 kW, respectively, which are comparable to a reported estimate of 26 kW of electricity per MGD (EPA, 2015). Annual electricity generation for each system in kWh/y was used to calculate the potential electricity savings, assuming an electricity cost of $0.11 per kWh. Costs for the CW, presented in Supplementary Table S10, were calculated based on EPA (2000) and adjusted to 2015 dollars.

Results and Discussion

Laboratory study—reactor performance

Performance of the ABR, with respect to solids and organic removal, was consistent, with percent removals of COD that ranged from 60% to 90% (Fig. 2) over 98 days. Furthermore, effluent COD concentrations ranged from 60 to 180 mg/L, although the influent concentrations varied widely from about 250 to 1,150 mg/L (Fig. 2), demonstrating the ability of this reactor to accommodate shock loads—a well-documented advantage of the ABR (Barber and Stuckey, 1999; Manariotis and Grigoropoulos, 2002; Foxon et al., 2004). Based on a measured BOD₅ to COD ratio of 0.6, we calculated an effluent BOD₅ concentration (average ± 95% CI) of 72.3 ± 4.7 mg/L. The ABR was operated with a constant HRT of ∼13 h, but increasing HRT can enhance removal of organics in anaerobic mainstream systems (Agrawal et al., 1997), suggesting that lower organic concentrations in the effluent may be achievable at the temperature range studied here. Solids concentrations (TSS) in the effluent never rose above 20 mg/L (Supplementary Fig. S4), demonstrating that the ABR alone can satisfy regulations of 30 mg/L solids. Additional performance data for the ABR are provided in Supplementary Figure S7.

FIG. 2.

Reactor performance: (a) Removal of organics in a laboratory-scale anaerobic baffled reactor (ABR) with a hydraulic retention time (HRT) of 0.55 days. Influent chemical oxygen demand (COD), effluent COD, and percent removal values are shown in black, blue, and red symbols, respectively. Lines do not represent model fits and were added to help guide the eye. (b) COD mass balances (mg/L) and distributions (%) for the ABR.

Laboratory study—COD mass balance

A COD mass balance (Fig. 2b) shows that solids produced by the ABR represent ∼9% of the influent COD. COD removed by the ABR, that is, the difference between influent and effluent COD concentrations, was 436 mg/L. In addition to stable removal of organics, the ABR converted 57% of the COD to methane, with 35% of the produced methane trapped in the dissolved phase (Fig. 2b). Therefore, on average, the ABR produced 2.5 L/day of gaseous CH₄, equivalent to a specific methane production of 0.17 ± 0.02 (average ± 95% CI) m³ per kg COD removed. The specific methane production, within the range of reported methane production values for ABRs of 0.13 to 0.23 m³ CH₄ per kg COD removed (Schoener et al., 2014), represents ∼37% of the COD in the influent (Fig. 2b). The mass balance closure of −12% for COD is similar to a previous study on anaerobic treatment of dilute wastewater (Shin et al., 2014).

Dissolved methane concentrations of 28.1 ± 3.0 (average ±95% CI) mg/L are within the range of reported values for anaerobic systems treating domestic wastewater (Shin et al., 2014; Smith et al., 2014). In addition, dissolved methane concentrations were supersaturated at 1.5 times the values expected based on Henry's Law, which is equivalent to the average of reported supersaturation values (Kim et al., 2011; Bandara et al., 2012; Smith et al., 2014). As is common in dilute anaerobic systems (Liu et al., 2014), dissolved methane in the effluent represented a significant fraction of the total methane produced. Unused dissolved methane represents a loss of bioenergy production and potential harmful impact for climate change. The performance data of the ABR, including amounts of methane trapped in the dissolved phase, were used for the design of a full-scale system used to estimate life cycle environmental impacts and costs described in the following sections of this article.

Modeling results

Environmental trade-offs for ABR and TF

Beneficial impacts (indicated by negative values) for the four modeled environmental indicators—climate change, resource depletion, ecosystem quality, and human health—for all treatment assemblies result from produced bioelectricity (Fig. 3). Bioelectricity was assumed to replace the average grid electricity in India, of which 65% is produced from coal, and 10% is produced from natural gas (Weidema et al., 2013), and the avoided impacts of grid electricity were thus credited to the system. Similarly, consumption of grid electricity contributes most to harmful impacts (as indicated by positive values) for the resource depletion, ecosystem quality, and human health indicators. Our results show that replacing grid electricity with bioelectricity recovered from wastewater may yield benefits on ecosystems and human health, in addition to improving net energy balances and reducing impacts on climate change, indicators on which most LCA studies focus. The effects of fossil fuel dominated electricity on human health and ecosystem impacts, in addition to climate change and resource depletion, have been shown previously for algae biofuel production (Beal et al., 2015; Gerber et al., 2016).

FIG. 3.

Life cycle environmental impacts (per 2 million gallons of treated wastewater) calculated with IMPACT 2002+ for four endpoint categories: (a) climate change (kg CO₂ eq.), (b) resource depletion (MJ of primary energy), (c) ecosystem quality (PDF · m²·years), and (d) human health (DALY). Contributions of individual processes are represented by colored bars, and net impacts are represented by black diamonds. Positive values represent harmful impacts, and negative values represent beneficial impacts. DALY, disability adjusted life year; MJ, megajoule; PDF, potentially disappeared fraction.

Results presented in Fig. 3a confirm that dissolved methane in anaerobic effluents may negate the beneficial impacts on climate obtained from bioelectricity production in anaerobic systems (Liu et al., 2014). Approximately 95% of the GWP for the ABR assembly was associated with dissolved methane. Smith et al. (2014) similarly showed that for an AnMBR, dissolved methane in the effluent (also assumed to be released to the atmosphere) contributed to ∼75% of the GWP. Much of the remaining GWP for the AnMBR was associated with operational electricity requirements (Smith et al., 2014), which are not required for an ABR. For the ABR and ABR+CW assemblies, dissolved methane led to a climate change impact approximately three times greater than the benefits resulting from bioelectricity production from ABR biogas. In addition, even if the ABR is followed by posttreatment with CW, which (as an independent treatment technology) sequesters CO₂ at a rate significant enough to reduce GWP (Kalbar et al., 2013), it cannot counteract the detrimental impact of dissolved methane (Fig. 3a). Although our results show that ABR treatment (at 15–20°C) is not beneficial for climate change unless a posttreatment is used, the benefits of implementing the ABR in developing countries should be weighed against these countries' smaller contributions to climate change relative to industrialized nations. For example, 36 developed or emerging countries account for ∼70% of global greenhouse gas (GHG) emissions, whereas 176 other countries account for only 20% of global GHG emissions (Baumert et al., 2005).

The ABR+TF assembly produced a net GWP that is an order of magnitude lower than the GWP associated with the ABR+CW assembly (black diamonds in Fig. 3a). Dissolved methane in the effluent for the ABR+CW assembly, which is assumed to be released to the environment, drives this large difference. The two aerobic polishing processes—TF and CW—were designed to reduce effluent BOD₅ concentrations to 30 mg/L, based on our laboratory results with average BOD₅ concentration of 72.3 mg/L in the ABR effluent. In addition to oxidizing organic matter, we assumed that the TF oxidized all of the dissolved methane in the ABR effluent. The two polishing processes represent two extreme possibilities, and, in reality, the amount of methane oxidized as part of aerobic polishing processes will likely lie in between these two examples. We present results of climate change impacts, however, for these two example aerobic polishing processes to illustrate the importance of removing dissolved methane from anaerobic effluent.

The overall GWP for each of the four assemblies (as indicated by the black diamonds in Fig. 3a and converted to kg CO₂-eq. per m³ of wastewater) ranged from −0.19 kg CO₂-eq. per m³ of treated wastewater (median value for the ABR+TF assembly) to 0.48 kg CO₂-eq. per m³ (median value for the ABR). The GWP of the ABR is comparable to the range of GWPs reported for an AnMBR system, which ranged from 0.10 to 0.85 kg CO₂-eq. per m³ of wastewater (Shoener et al., 2016). Since for both treatment technologies (i.e., the AnMBR in Schoener et al., and the ABR modeled herein) the main contributor to climate change is dissolved methane, it follows that the impacts are similar. Note that the lower values provided by Schoener include capture of dissolved methane with a degassing membrane. Smith et al. reported that for conventional activated sludge+AD, climate change impacts were ∼0.08 kg CO₂-eq. per m³ of wastewater. In contrast, the impacts of the TF and ABR+TF assemblies on climate change were beneficial (i.e., negative values), demonstrating the advantage of using a fixed-film, low-energy aerobic process in place of conventional activated sludge.

Energy recovered by the four assemblies, as shown with the Resource Depletion indicator (Fig. 3b), demonstrates the advantage of using technologies with low-energy requirements to recover energy from wastewater. All four assemblies recover more energy than they consume, as is evident by the negative values presented as black diamonds in Fig. 3b. Furthermore, when uncertainty is accounted for, the ABR, ABR+TF, and ABR+CW assemblies are likely to recover more energy than the TF assembly (Fig. 4). High energy recovery relies on the assumption that gaseous methane produced by the ABR and AD was converted to heat and power. However, it is common for mainstream anaerobic treatment in developing countries to not capture produced methane. In addition to capturing methane, the locations of energy producing sanitation systems should be chosen appropriately to enable the use of captured energy. If gaseous methane is not captured and converted to heat and power, all treatment assemblies modeled herein will produce harmful impacts for the climate change category (Fig. 3). In addition, the TF and ABR+TF assemblies will produce harmful impacts for all impact categories, whereas the ABR and ABR+CW will produce approximately neutral impacts for the resources, ecosystem quality, and human health categories.

FIG. 4.

Overall life cycle environmental impacts of each of the four treatment assemblies (TF, ABR, ABR+TF, and ABR+CW) for four endpoint categories: climate change, resource depletion, ecosystem quality, and human health in units of LCA points. Center lines represent median values, edges of boxes represent 25th and 75th percentiles, and limiting bars represent 5th and 95th percentiles of the distributions resulting from 1,000 Monte Carlo simulations. Positive values represent harmful impacts, and negative values represent beneficial impacts. CW, constructed wetland; LCA, life cycle assessment.

If dissolved methane is captured as gaseous methane and converted to heat and power, the beneficial impacts on resources, ecosystem quality, and human health will increase for the ABR, ABR+TF, and ABR+CW assemblies. For example, assuming that dissolved methane represents approximately one-third of the produced methane (as shown in Fig. 2b), beneficial impacts (for all impact categories) from bioelectricity production will increase by ∼30% for the ABR and ABR+CW assemblies and 20% for the ABR+TF assembly. Moreover, capture or oxidation of dissolved methane will cause the net climate change impact of the ABR and ABR+CW assemblies to be beneficial. A sustainability analysis of AnMBRs showed that large reductions in GHG emissions may be possible if a degassing membrane is used to capture dissolved methane (Bandara et al., 2011; Cookney et al., 2016; Shoener et al., 2016). In addition, aerobic oxidation of dissolved methane in anaerobic effluent has been demonstrated (Matsuura et al., 2015), and coupling methane oxidation with nitrogen removal may yield synergistic benefits. For example, dissolved methane may be used as an electron donor for denitrification (Kampman et al., 2012), which would extend the systems modeled here to remove nitrogen from wastewater, in addition to managing dissolved methane. The benefits of anaerobic systems are strongly dependent on the conversion of gaseous methane to useful forms of energy, such as electricity, as well as oxidation or valorization of dissolved methane. Furthermore, in some regions, coupling such energy recovery with water-reuse technologies may increase overall energy yields and enhance access to clean water (Mo and Zhang, 2012).

For all impact categories, disposal of biosolids was about five times more harmful for the TF assembly compared to the ABR (Fig. 3), due to the bioenergetics of anaerobic versus aerobic processes. This demonstrates that in addition to reducing energy consumption and costs, as has been shown previously (Smith et al., 2014), less biosolids production reduces impacts on ecosystems and human health.

One limitation of this study is that the reactors were modeled to treat low-strength wastewater, but wastewater in developing countries is more concentrated. The ABR and TF were designed based on hydraulic loading and will be able to accommodate wastewater with higher concentrations of organic matter that is typical of developing countries. ABR treatment of higher-strength wastewater will produce more methane, and the amount of methane dissolved in the effluent will most likely not change (compared to treatment of low-strength wastewater), since wastewater strength is not expected to change methane solubility. Thus, as shown in Smith et al. (2014) for an AnMBR, all of the incremental increase in methane produced can be captured as biogas, improving the beneficial environmental impacts (relative to the treatment of low-strength waste) of the ABR.

Local health conditions, such as the risk of exposure to pathogens, should be accounted for and incorporated into the design and implementation of anaerobic systems (Manser et al., 2015). The LCA conducted herein, however, did not account for pathogens present in effluent of the ABR, when assessing impacts on human health, because no life cycle impact assessment methods exist that model pathogen-induced disease. To improve LCA of wastewater treatment, researchers working at the nexus of wastewater and LCA could develop new LCA characterization factors that convert concentrations of pathogenic organisms into DALYs.

Uncertainty analysis

The range of results presented for the impact categories is presented in points of environmental damage (LCA points), and each point represents the impact of one person in 1 year in Europe. The four endpoint categories were normalized to allow a comparison among four different impacts (climate change, resource depletion, ecosystem quality, and human health). Comparing impacts, a recommended practice in LCA (Jolliet et al., 2003), may prevent shifting environmental burdens from one impact (e.g., climate change) to another (e.g., human health) when using LCA as part of the engineering design process. In addition, normalizing to LCA points provides more information than normalizing to minimum and maximum values of impact. Since there are no normalization factors for developing countries, we presented impacts normalized within the European context to allow readers to compare among the four damage categories. Results for climate change (Fig. 4), which agree with the results that were presented without uncertainty (Fig. 3a), show that the ABR and ABR+CW assemblies are detrimental to climate change, whereas the ABR+TF assembly is likely beneficial for climate change. In addition, the ranges of results presented in Fig. 4 show that the TF assembly is likely neutral for climate change.

For the ecosystem quality indicator, when uncertainty is accounted for, the four treatment assemblies are likely to produce only slightly beneficial impacts. In contrast, the three assemblies that include ABR treatment yield more beneficial impacts on human health than the TF assembly does. As discussed previously, the beneficial impacts represent avoided harmful impacts from electricity consumption. The high percentage of fossil fuel in India's electricity grid combined with power plants that produce harmful pollution is the reason for the avoided impacts, resulting from the production of bioelectricity that replaces grid electricity.

For climate change, even when accounting for wide ranges of uncertainty, anaerobic treatment coupled with removal of dissolved methane (represented by the ABR+TF assembly in Fig. 4) results in the most beneficial impact. For the three other impacts—namely, resource depletion, ecosystem quality, and human health—the ABR, ABR+TF, and ABR+CW perform similarly.

Technoeconomic analysis

The ABR is the least expensive technology with a life cycle cost, in terms of NPV, that is ∼40% lower than the TF assembly (Fig. 5). In addition, the capital costs for the TF and ABR+TF assemblies exceed the total life cycle costs for the ABR and ABR+CW assemblies. The capital costs for all assemblies were approximately three times greater than the operating costs, which reflect the low-energy requirements of the technologies assessed and illustrate the economic benefit of recovering energy from domestic wastewater. The results are in agreement with Smith et al., who reported that the present worth of capital costs for conventional activated sludge with AD and an AnMBR was ∼2 and 2.5 times greater than the present value of operating costs (Smith et al. 2014).

FIG. 5.

Life cycle costs presented as net present values for TF, ABR, ABR+TF, and ABR+CW. Bars represent a discount rate of 3.375%. Diamonds represent the sum of capital and operating costs calculated with a discount rate of 8%.

Treatment and disposal of solids were a key contributor to both capital and operating costs (Supplementary Tables S4 and S5) and contributed to 25% and 22% of the capital costs for the TF and ABR+TF assemblies, respectively—most of which were associated with AD. In contrast, treatment and disposal of solids contributed to about 7% of capital costs for the ABR and ABR+CW assemblies, demonstrating an economic advantage of anaerobic treatment, namely lower production of biosolids compared to aerobic treatment.

The TEA conducted here likely overestimates the capital and operating costs of each technology, because CAPDETWorks overestimates costs for small treatment systems. In addition, costs were determined for the United States and not for a developing country. However, since costs were likely overestimated similarly for the four treatment technologies, the comparison presented here would scale similarly, resulting in a similar comparative analysis.

Summary

Laboratory data acquired during treatment of municipal wastewater with a bench-scale ABR, showed that the ABR removed ∼75% of the COD and that about one-third of the produced methane remained in the dissolved phase. These data were used as a basis for full-scale modeling of ABR treatment. A modeling framework—which included process models, LCA coupled with uncertainty analysis, and TEA—was applied to four wastewater treatment assemblies: ABR, TF, ABR+TF, and ABR+CW. The results illustrated the trade-offs among four environmental indicators (climate change, resource depletion, ecosystem quality, and human health) and life cycle costs. The overall environmental impact—that is, the sum of the four impact categories in units of LCA points—versus the life cycle costs (presented as NPV) of the four process assemblies shows that the ABR has the lowest NPV, but it may be harmful to the environment (Fig. 6). However, all of the ABR's harmful environmental impacts result from the release of dissolved methane to the atmosphere. If technologies that oxidize or capture dissolved methane are used, the ABR will benefit the environment through energy recovery—as indicated by the large negative value for LCA points associated with the ABR+TF (Fig. 6). Developing low-cost processes to recover and valorize dissolved methane in anaerobic effluent will allow implementation of the ABR in developing countries without increasing risks of climate change.

FIG. 6.

Average values for total environmental impact (sum of LCA points from climate change, resource depletion, human health, and ecosystem impacts) versus life cycle costs (shown as net present value) for TF, ABR, ABR+TF, and ABR+CW (CW) assemblies. The shaded arrows indicate beneficial impacts and lower costs (green) and harmful impacts and higher costs (red).

Footnotes

Acknowledgments

The authors thank the Bucknell University graduate studies program, the Swanson family, and the Civil & Environmental Engineering Chiloro Endowment for funding. They thank Bob Michael and Genie Bausinger, Milton Regional Sewer Authority, for help with sample collection. They thank Steve Beightol, Huan Luong, Monica Hoover, and D.J. Wacker, Bucknell University, for their laboratory expertise. They thank Bucknell undergraduate students Dylan Cowell, Melissa Warshauer, and Christian Intriago-Velez for assistance with sample collection and reactor operation. Finally, they thank two anonymous reviewers for providing feedback to improve this article.

Author Disclosure Statement

No competing financial interests exist.

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