Science of the Total Environment 505 (2015) 1166–1173
Contents lists available at ScienceDirect
Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv
Impact of direct greenhouse gas emissions on the carbon footprint of water reclamation processes employing nitrification–denitrification Andrew G. Schneider a,⁎, Amy Townsend-Small a,b, Diego Rosso c a b c
University of Cincinnati, Department of Geology, Cincinnati, OH 45221, United States University of Cincinnati, Department of Geography, Cincinnati, OH 45221, United States Department of Civil and Environmental Engineering, University of California, Irvine, CA 92697-2175, United States
H I G H L I G H T S • Direct greenhouse gas emissions were measured at a wastewater reclamation plant. • These greenhouse gas emissions amounted to 3.9 (±0.5) g-CO2-eq m− 3 of wastewater. • 14C analysis of the CO2 emissions was conducted to determine the fossil component. • 11.4% to 15.1% of the emitted CO2 was derived from fossil sources.
a r t i c l e
i n f o
Article history: Received 20 April 2014 Received in revised form 5 October 2014 Accepted 19 October 2014 Available online 14 November 2014 Editor: Simon Pollard Keywords: Water reclamation Greenhouse gases Direct emissions Radiocarbon Carbon footprint Fossil carbon
a b s t r a c t Water reclamation has the potential to reduce water supply demands from aquifers and more energy-intensive water production methods (e.g., seawater desalination). However, water reclamation via biological nitrification– denitrification is also associated with the direct emission of the greenhouse gases (GHGs) CO2, N2O, and CH4. We quantified these direct emissions from the nitrification–denitrification reactors of a water reclamation plant in Southern California, and measured the 14C content of the CO2 to distinguish between short- and long-lived carbon. The total emissions were 1.5 (±0.2) g-fossil CO2 m−3 of wastewater treated, 0.5 (±0.1) g-CO2-eq of CH4 m−3, and 1.8 (± 0.5) g-CO2-eq of N2O m−3, for a total of 3.9 (± 0.5) g-CO2-eq m−3. This demonstrated that water reclamation can be a source of GHGs from long lived carbon, and thus a candidate for GHG reduction credit. From the 14C measurements, we found that between 11.4% and 15.1% of the CO2 directly emitted was derived from fossil sources, which challenges past assumptions that the direct CO2 emissions from water reclamation contain only modern carbon. A comparison of our direct emission measurements with estimates of indirect emissions from several water production methods, however, showed that the direct emissions from water reclamation constitute only a small fraction of the plant's total GHG footprint. © 2014 Elsevier B.V. All rights reserved.
1. Introduction Increases in atmospheric greenhouse gas (GHG) concentrations are linked to human activities (Forster et al., 2007). Waste treatment and storage are global sources of greenhouse gases, particularly nitrous oxide (N2O) and methane (CH4) (Ahn et al., 2010; Daelman et al., 2012). In a wastewater treatment system, the microbial processes of nitrification and denitrification produce N2O (Ahn et al., 2010; Townsend-Small et al., 2011), a GHG that has a global warming potential 298 times that of carbon dioxide (CO2) over a 100-year horizon (Forster et al., 2007). CH4 is produced by the decomposition of organic matter under the anaerobic conditions found in several points of the wastewater treatment process, and has a global warming potential 25 ⁎ Corresponding author. Tel.: +1 203 785 2912. E-mail address: [email protected] (A.G. Schneider).
http://dx.doi.org/10.1016/j.scitotenv.2014.10.060 0048-9697/© 2014 Elsevier B.V. All rights reserved.
times that of CO2 during a 100-year horizon (Forster et al., 2007). CO2 is produced directly in wastewater treatment via the aerobic decomposition of organic matter, as well as indirectly through chemical manufacturing and the combustion of fossil fuels for electricity production and transportation. Annually, wastewater treatment in the United States emits approximately 7.79 × 108 kg of CH4 (2.5% of total American CH4 emissions) and 1.63 × 107 kg of N2O (1.6% of total American N2O emissions) (EPA, 2012). Akin to numerous urban and suburban conglomerates across the United States and abroad, Southern California has undergone rapid urban growth, economic progress, and population expansion over the past few decades. Groundwater alone cannot meet the water demands of the region due to the low recharge rate and threat of saltwater intrusion into the coastal aquifers. Long-distance importation is the largest source of potable water to the Los Angeles and San Diego areas, though the electricity required for importation is a major source of CO2 (Stokes
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
and Horvath, 2009). In order to reduce the impact of importation on the upstream regions, desalination and water reclamation (the extension of wastewater treatment with provisions for beneficial reuse) are also being pursued. The process of water reclamation purifies and recycles the water that is already in the collection system, reducing the demands placed on aquifers and reservoirs. Although water reclamation may relieve some of the stress on aquifers and reduce the energy required (and therefore the CO2 emitted) for long-distance importation or desalination, these changes may lead to an increase in direct N2O and CH4 emissions from biological treatment. Studies have determined that water importation and water reclamation have comparable indirect emissions of greenhouse gases, about 1 kg-CO2-eq m−3 of water, with sea-water desalination having a much greater energy demand, resulting in about 2.4 times more CO2 equivalents m−3 than water reclamation (Stokes and Horvath, 2009). It was estimated that the CH 4 and N 2 O emissions from American domestic and industrial wastewater treatment in 2010 totaled 2.13 × 1010 kg-CO2-eq (EPA, 2012). The annual GHG emissions in the Southern California air basin include 6 (±1) × 108 kg-CH4 y−1 (Wunch et al., 2009), and our previous work has confirmed that wastewater treatment emits CH4 in Los Angeles (Townsend-Small et al., 2012). Furthermore, our research in the region has shown that, although water reclamation is a minor source overall, it may be the principal point-source of N2O, emitting 7.6 × 104 kg-N2O y− 1 (2.26 × 107 kgCO 2 -eq y − 1 ) in Orange County (area = 2455 km 2 ; 2010 population = 3 million) alone (Townsend-Small et al., 2011). More broadly, wastewater treatment is not unique to the developed world, and urbanization, economic development, and population growth may result in CH4 and CO2 emissions on the order of 107 kg-CO2-eq y− 1 from wastewater systems in emerging nations by 2025 (Rosso and Stenstrom, 2008). Yet while wastewater treatment and water reclamation may be associated with some greenhouse gas emissions, it is essential to remember that these processes are designed to foster environmental and human health and that the scenario of no treatment is characterized by high emissions with no biogas energy recovery. One previous study measured both the direct CH4 and N2O emissions from a small wastewater treatment plant in New Hampshire, providing some of the first published data on the process (Czepiel et al., 1993, 1995). More recently, Ahn et al. conducted a broad study of N2O emissions from plants across the United States (Ahn et al., 2010). Wang et al. (2011) and Daelman et al. (2012) conducted measurements of CH4 from wastewater treatment in China and Europe, respectively. Direct emissions of CO2 during the treatment process are often thought to be negligible or fully biogenic (i.e., short-lived carbon) in nature (Rosso and Stenstrom, 2008). Most researchers acknowledge that CO2 is emitted as part of the water treatment and recycling process, but do not measure the fossil component of the CO2, with the exception of few cases (e.g., Griffith et al., 2009; Law et al., 2013). The goal of this study was to estimate direct greenhouse gas emissions from water reclamation in order to determine whether water reclamation has a larger carbon footprint than long-distance water importation. The three objectives of our study were: 1) to measure the emission rates of CO2, CH4, and N2O from water reclamation; 2) to identify the proportion of the emitted CO2 that was derived from fossil sources using 14C analysis; and 3) to compare the direct emissions of GHGs from water reclamation during operation to the indirect GHGs associated with the process. 2. Materials and methods 2.1. Field site description Samples were collected from a wastewater reclamation plant in Southern California in the early spring. The plant reclaims water employing a modified Ludzack–Ettinger (MLE) activated sludge system
1167
for nitrification–denitrification with supplemental methanol addition. During the period of our sampling, the flow rate ranged from 34,800 m3 d−1 to 75,000 m3 d−1, averaging 64,100 m3 d−1. In general, the mean cell retention time (MCRT) varies between 6 and 10 days through the seasons to compensate for temperature variations, and two on-site flow-equalization basins are used to abate hydraulic peaks across the diurnal cycle. During the period of testing, the average raw influent contained approximately 235 mg L−1 of total suspended solids (TSS), with a 201 mg L−1 of biochemical oxygen demand (BOD5) and 20.6 mg-NH4–N L−1. This plant received municipal wastewater with negligible industrial contribution, therefore the ratio between BOD5 and the chemical oxygen demand (COD) is expected to be aligned with municipal values for separate sewer systems in areas with food waste disposal to drain (i.e., 0.40–0.45). The average mixed liquor suspended solid (MLSS) was 2500 mg/L, with an average volatile suspended solids (VSS) fraction of 79%. Primary treatment reduced this to 102 mg-TSS L− 1 and 114 mgBOD5 L−1, while bypassing the 20.6 mg-NH4–N L−1 to the secondary reactors. Effluent water from the plant, after undergoing secondary treatment, tertiary filtration and disinfection, contained a maximum of 1 mg-TSS L−1, 2 mg-BOD5 L−1, and 0.5 mg-NH4–N L−1. Depending on demand, roughly half of the effluent enters a recycled-water distribution system and is used for irrigation of agriculture and landscaping, along with other industrial and commercial purposes (i.e., CA Title 22 water). The remaining effluent is sent to reservoirs. Our sampling was conducted in three of the secondary treatment reactors: the anoxic (denitrification) and aerobic (nitrification) reactors, and the settling basins (illustrated in Fig. 1), as previous research indicated that these sites were the sources of the majority of the on-site GHG emissions (Townsend-Small et al., 2011). There are six anoxic reactors in parallel, with a total volume of approximately 4300 m3. Each anoxic reactor is followed by aerobic tanks approximately double in volume, with a combined aerobic volume of 12,900 m3. Treated wastewater from the aerobic reactors is recombined and distributed to the settling basins (nine rectangular, one circular), with a combined volume of 11,500 m3. The dissolved oxygen (DO) set point of the aerobic tanks is 2.4 mg/L, while the DO set point for the settling basins is 2.7 mg/L.
Fig. 1. A flow diagram of the secondary treatment reactors with our approximate sampling locations (circled X): A) anoxic reactors (denitrification), B) aerated reactors (nitrification and carbon oxidation), and C) secondary clarifiers (settling basins). The single circular settling basin is physically separated from the nine parallel rectangular ones, but receives the same flow and has the same retention time. Wastewater enters this section of the system from the primary clarifiers and flow equalization basins, and exits to the tertiary filters and disinfection contact basins. The dashed lines between the anoxic and aerated reactors are baffles, used to isolate the anoxic and aerobic reactors. Dotted lines represent internal circulation.
1168
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
2.2. Sampling method
3. Results and discussion
Greenhouse gas fluxes were measured twice a day for five days in the early spring season from the surface of the reactors using three floating chambers (Beaulieu et al., 2012). The Plexiglas chambers had an interior volume of 7.73 L and a sectional area of 0.067 m2 when afloat. They were constructed following U.S. Department of Agriculture guidelines for trace gas measurements (USDA, 2010). A flexible tube (6.35 mm interior diameter) with a stopcock was attached to the chamber for sampling, with another opening on the chamber for pressure compensation. Due to the 1 m freeboard sheltering the surface of these reactors from the wind, the high off-gas flow rate, and given that the wind speed during testing was negligible, the effect of windinduced pressure gradients was deemed insignificant. In an attempt to capture the differences between tanks in each process while retaining some redundancy, two of the chambers were placed in the same well-mixed reactor (equipped with an active subsurface mixer) and one in a contiguous reactor train during each period of measurement (marked in Fig. 1). Sample gas extraction was begun immediately after placing the chamber on the surface in order to measure the rate of GHG concentration change in the headspace. Before each sample was collected, 200 mL of gas was drawn from each chamber to clear the tubing. Once the tubing was cleared, 30 mL of sample gas was extracted and stored in pre-evacuated 20 mL glass vials sealed with gray rubber butyl septa and aluminum crimps. Samples for radiocarbon analysis of CO2 were collected using evacuated 5 L stainless steel canisters. These samples were collected during the afternoon of Day 2 and during the morning of Day 3, with one sample from each of the three reactors being measured per day. The gases for radiocarbon analysis were drawn from the floating chambers at the end of the sampling interval in order to allow the greatest concentration of CO2 to accumulate in the chamber headspace.
3.1. Flux rates
2.3. Sample analysis Greenhouse gas flux samples were analyzed using a Shimadzu GC2014 Greenhouse Gas Analyzer, with flame-ionization detector for measurement of CH4 and CO2 and an electron-capture detector to measure N2O. A least-squares linear regression method comparing the sampled gas concentrations against the time of sampling was employed to calculate emission rate for the three gases in each chamber (USDA, 2010). These flux rates were then used to extrapolate the total flux of CO2, CH4, and N2O for the plant's secondary treatment basins with the formula, f V C P A M SB min −6 L 10 1440 day μL AC R T Q D
f VC PA M SB AC R T QD
flux rate (ppm min−1) volume of sampling chamber (L) atmospheric pressure (atm) molar mass of the compound (g mol−1) surface area of the basin (m2) surface area of sampling chamber (m2) gas law constant (0.08206 L atm K−1 mol−1) air temperature above basin (K) daily flow rate (m3 d−1).
The results were converted into CO2 equivalents via the IPCC method (Forster et al., 2007) using the 100-year global warming potential (GWP) for N2O and CH4. Samples for radiocarbon analysis were converted to graphite and analyzed for 14C content by accelerator mass spectrometry (Xu et al., 2007).
We measured the surface emissions of GHGs, which are presented here initially in units of mass per surface area per time. CH4 flux rates were 0.13 (±0.02) mg-CH4 m−2 min−1 in the anoxic reactor, decreasing to 0.08 (± 0.04) mg-CH4 m− 2 min− 1 in the aerated reactor, and increasing to 0.18 (± 0.03) mg-CH4 m−2 min−1 in the settling basin (all numbers are ±1 SE). The total CO2 emission rates fluctuated from 14.7 (± 1.1) mg-CO2 m−2 min− 1 up to 231.7 (± 29.8) mg-CO2 m− 2 min− 1, and back down to 17.4 (± 1.3) mg-CO2 m− 2 min−1 through the three reactors. Radiocarbon data indicated that 11.4 (± 0.2) % of the CO2 flux from the anoxic reactor, 15.1 (± 0.2) % of the CO2 from the aerated reactor, and 12.0 (±0.2) % of the CO2 flux from the settling basin originated from fossil carbon. N2O followed a similar pattern as CO2, with an emission rate of 6.1 (± 0.5) μg-N2O m−2 min−1 in the anoxic reactor (denitrification), increasing an order of magnitude to 84.2 (±19.3) μg-N2O m−2 min−1 in the aerated reactor (nitrification), and dropping back down to 0.009 (± 0.001) μg-N2O m−2 min− 1 in the settling basin. (See supplemental information for specific flux measurements.) 3.2. Methane emissions The areal CH4 flux rate was greatest in the settling basin and lowest in the aerated reactor (performing carbon oxidation and nitrification). CH4 emissions from the aerobic tank were expected to drop because of the inhibition to CH4 formation and the CH4 oxidation occurring in the aerobic conditions. In addition, the settling basins had a surface area nearly six times the size of the anoxic reactor (performing denitrification) and about twice the size of the aerated reactor. In the settling basin, this elevated emission rate and large surface area resulted in a total average emission of 859 (± 147) g-CH4 d−1 (Table 1, Fig. 2), or 74% of the total CH4 emissions measured. In comparison, the anoxic and aerated reactor emission rates were 106 (±13) g-CH4 d−1 (9% of total) and 196 (± 105) g-CH4 d− 1 (17% of total), respectively. The large variability in the emissions rates from the aerobic reactor for CH4 and other gases was potentially due to uneven aeration during the relatively short interval of sample collection. Methanogenesis from biodegradation occurs under anaerobic conditions and is controlled by the amount of organic matter in the wastewater (El-Fadel and Massoud, 2001; Cakir and Stenstrom, 2005; EPA, 2012). The sludge blanket at the bottom of the settling basin had low or no (DO) and high organic matter content, enabling CH4 production. It is unlikely that significant amounts of CH4 were produced in the nitrification reactor due to the inhibition to methanogens and the promoted methane oxidation caused by oxygenation, so we assume that the CH4 emitted from the aerated reactor was from the stripping of CH4 produced in the (anoxic) denitrification reactor. The CH4 emitted by secondary treatment cannot be captured for energy production due to its elevated dilution with other atmospheric gases in the off-gas. In other locations where off-gas is used as feed air for flares or combined heat and power generation systems, however, the oxidation of this CH4 to CO2 is achieved. The total CH4 emission rate from the entire secondary treatment system during the period of sampling was 22.0 (± 3.7) mg-CH4 m−3 (Table 2, units are per cubic meter of wastewater). Other researchers (Foley et al., 2010) have attempted to model the emission rate of water reclamation plants that employ a modified Ludzack–Ettinger activated sludge system, though their results were significantly higher (610 mg-CH4 m−3). The difference may be due to the elevated DO set points selected for this plant (2.4–2.7 mg-O2 L− 1), which promote thorough oxygenation of the wastewater during aeration and may help inhibit anaerobic activity inside bacterial flocs outside the oxic reactors. Measurements of emissions from aerated reactors in other types of wastewater processes have also
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
1169
Table 1 Average measured GHG flux rates from the three reactors. Anoxic reactor (denitrification) CH4
n Mean SEMean
28 0.11 0.01
CO2
Aerobic reactor (nitrification) N2O
Total
Fossil
27 12.1 0.9
27 1.37 0.11
26 5.0 0.4
CH4
19 0.20 0.11
CO2
Settling basin N2O
Total
Fossil
19 190.4 24.5
19 65.05 8.37
18 207.6 47.7
CH4
25 0.86 0.15
CO2
N2O
Total
Fossil
27 14.3 1.1
27 9.50 0.70
26 42.9 5.7
n = number of observations, mean = arithmetic mean, SEMean = standard error of the mean. CH4 values in kg-CH4 d−1, CO2 values in kg-CO2 d−1, N2O values in g-N2O d−1.
found annual CH4 fluxes higher (0.2 g-CH4 m−3) (Czepiel et al., 1993) than those found in the current study (0.02 g-CH4 m−3). Also, the facility studied here had no sludge processing on site; hence the return flow (i.e., supernatant and centrate or filtrate, rich in COD and ammonia) was not part of our measurements. This may imply that the separate treatment of return flows (by means, for example, of sidestream ammonification) may severely reduce the emission from the mainstream train, albeit compensating the total sum with their own emission. Future studies should focus on separate processes for mainstream and sidestream treatment, so as to compare the total emission balance with traditional, combined treatment. Most other reported values of CH4 from both physical sampling and modeling are an order of magnitude higher than those observed in the current study (Czepiel et al., 1993; Daelman et al., 2012; Wang et al., 2011). The plant influent had an average of 201 g-BOD5 m−3, which indicated that the CH4 emission rate was 0.11 (±0.02) g-CH4 kg-BOD−1 5 [1.16 (±0.21) g-CO2-eq kg-COD−1], or 0.011% of the BOD5 (the average carbon emission intensity of untreated influent COD is 0.99 kg-CO2eq kg-COD−1; Gori et al., 2011). This ratio, which can be interpreted as a relative assessment of aerobic and anaerobic digestion, was lower than those measured by others, 0.16% of influent BOD5 in New Hampshire (Czepiel et al., 1993) and 0.08% of influent COD in China (Wang et al., 2011). Comparisons with other research findings are complex because of the different systems employed and the lack of sludge processing and return flows at this plant. Another factor that may play a role in the reduction of emission fluxes here is the unusually high depth of the biological reactors (approximately 8.5 m), which may offer a chance for produced CH4 in the water phase to be recaptured and oxidized by methanotrophs before reaching the surface. 3.3. Nitrous oxide emissions Per square meter, the N2O flux rate was highest in the aerated reactor and lowest in the anoxic reactor (denitrification). When the emission rates were extrapolated across the surface areas of the three reactors, the aerated reactor was the largest source of N2O. It emitted 207.6 (±47.7) g-N2O d−1 (approximately 81% of the total), compared to 5.0 (± 0.4) g-N2O d−1 from the anoxic reactor (2%) and 42.9 (± 5.7) g-N2O d− 1 from the settling basin (17%) (Table 1). Others have observed similarly elevated emissions in the aerated reactor, and have proposed several mechanisms that may result in these high rates, two of which were possible here: increased air-stripping of N2O, and N2O production by nitrification during variations in the aeration rate linked to influent (i.e. ammonia) buildup (Ahn et al., 2010). We were unable to differentiate between these two possibilities based on our measurements. Our previous work, however, showed that N2O emitted from the aerated reactor had an N isotope signature consistent with nitrification and distinct from N2O emitted from the denitrification reactor (Townsend-Small et al., 2011), indicating that the second proposed mechanism was most likely. The average N2O emission factor from the secondary treatment was 0.03 (±0.01) % of the influent nitrogen load. This put the N2O emissions near the low end of other previously observed N2O emission fractions of 0.03 to 0.11% from similar types of water reclamation plants (Ahn et al.,
2010). In addition, the N2O emissions from the aerated reactor averaged 4.8 (±1.5) mg-N2O m−3, well below the range of emissions previously observed (31 mg-N2O m−3) at a wastewater treatment plant in New Hampshire (Czepiel et al., 1995). Previous work at this plant in Southern California (Townsend-Small et al., 2011) observed that the dissolved N2O emission factor was an order of magnitude higher than the emissions from the New Hampshire plant. These discrepancies in results from the same plant were in concordance with an analysis of the importance of sampling strategy for accurate flux estimation and comparison, reported elsewhere (Daelman et al., 2012). Since our ultimate purpose here was not to compile a detailed flux inventory of one specific gas but rather to characterize the broader greenhouse gas makeup, the high-frequency analysis recommended in Daelman et al. was not employed here. The measured N2O flux from the secondary treatment system was 6.2 (±1.6) mg-N2O m−3 (Table 2). This emission rate represents only a small fraction of the N2O that remained dissolved in the fluid. Our previous work found an average of 271.7 mg-N2O m−3 dissolved in the anoxic reactor, 108.1 mg-N2O m−3 dissolved in the aerobic reactor, and 11.5 mg-N2O m−3 dissolved in the settling basin. In this current study, the measured emission rates were much smaller, 0.08 (±0.1) mg-N2O m−3 from the anoxic reactor, 4.7 (±1.4) mg-N2O m−3 from the aerated reactor, and 0.9 (±0.2) mg-N2O m−3 from the settling basin. Our measured emissions represented 0.03%, 4.4%, and 7.6% of the dissolved N2O in each reactor, respectively. These two sets of data were collected about two years apart, though the processing conditions were not significantly modified during the interim period. Previous work assessed dissolved N2O concentrations in each reactor and assumed that the bulk of this was emitted to the atmosphere eventually (Townsend-Small et al., 2011). The current study suggests that the N2O present in the treatment tanks is either consumed by denitrifiers, emitted to the atmosphere later in the process, or both. Although possibly consequential, downstream emissions of N2O from the effluent were not considered in this investigation. It may also be important to note that in some wastewater treatment methods, the overwhelming majority of gaseous nitrogen release may be in the form of N2–N (Teiter and Mander, 2005). There was a significantly greater proportion of dissolved N2O emitted from the aerobic reactor than from the anoxic reactor, though this was still only a fraction of total dissolved N2O (4.4%). A higher percentage (7.6%) of the total dissolved N2O was emitted from the settling basin, yet by this point in the process the amount of total dissolved gas had dropped considerably. The transient high N2O concentration in the anoxic reactors is maintained by a combination of gentle deep mixing (i.e., no mechanical device shearing the water surface and promoting stripping) and a lack of volumetric stripping through aeration. These high-concentration conditions persist throughout the anoxic reactors but are heavily altered as soon as the liquid enters the aerobic reactor, i.e. where the volumetric air flow strips the excess N2O from the liquid phase. The anoxic reactor was the site of the highest measured N2O production, though most of the gas remained dissolved. Some N2O was also produced in the aerated reactor (during nitrification), much of which was stripped out of solution and emitted into the atmosphere. Dissolved oxygen levels can also play a role in the emission of N2O, though the increased emissions that occurred at the
1170
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
3.5
Table 2 Total direct GHG emissions across all three basins, assuming a normal flow of 69,300 m3 of wastewater per day.
3.0 2.5
g∙m−3 ±
kg-CH4 d-1
2.0 CH4 Fossil CO2 N2 O Total
1.5 1.0 0.5 0.0
kg∙d−1 ±
kg-CO2eq∙d−1
±
0.022 1.507
0.004 0.549 0.246 1.507
0.092 1.26 0.246 82.12
0.20 9.09
31.4 82.1
4.9 9.1
0.006
0.002 1.828 3.885
0.468 0.536
0.05
82.3 195.8
15.5 18.6
0.28
-1.0 -1.5
3.4. Fossil carbon dioxide emissions
140
In the aerated reactor, a constant supply of atmospheric oxygen supports the aerobic decomposition of biological waste, resulting in the production of CO2 through bacterial aerobic respiration. Our observations fit this principle well, with the poorly-mixed anoxic reactor emitting the least amount of CO2 (Table 1) and the well-oxygenated aerated reactors emitting the largest amount of CO2. The direct emissions of CO2 from wastewater are assumed to derive from modern organic matter in human excreta or food waste, and such emissions of recently-photosynthesized carbon are believed to play an insignificant role in climate change (EPA, 2012). This “modern” organic carbon is part of an active, short-term cycle in which it is processed through the environment as recent biological matter, waste, soil carbon, atmospheric gases, etc. over relatively short time scales (generally b106 years, though the scale can vary dramatically). Regardless of human intervention, this “modern” organic carbon would be present in the atmosphere. In comparison, “fossil” organic carbon is that which has been removed from the active, short-term, atmospheric cycle. This carbon forms fossil fuels (coal, petroleum, natural gas) with turnover timescales exceeding 106 years. The distinction of “anthropogenic” greenhouse gas CO2 is caused by human activity that altered this timescale, exhuming the “fossil” carbon and injecting it into the short-term cycle faster than it can be removed. Neither the U.S. Environmental Protection Agency (EPA) nor the Intergovernmental Panel on Climate Change (IPCC) considers wastewater treatment to be a significant source of CO2 for greenhouse gas accounting (Doorn et al., 2006; EPA, 2012). The fossil-fuel derived carbon in wastewater may be a relevant portion of total C processed in treatment, however, giving decision makers, operators, and utility managers an incentive for sequestration with the associated credits. One previous study conducted in the Northeastern United States found that approximately 25% of wastewater dissolved organic carbon (DOC) is fossil carbon, most likely derived from cleaning products, pharmaceuticals, and other petroleum-based products (Griffith et al., 2009). Another, conducted in Australia, concluded similarly that 4–14% of the TOC is of fossil origin (Law et al., 2013). Some water reclamation plants supplement the readily biodegradable carbon in the primary effluent with methanol to provide a labile substrate for denitrification (Townsend-Small et al., 2011). The plant under study added about 1.6 m3 of methanol a day to the anoxic reactors, roughly 24 μL of methanol per liter of wastewater. Although we do not know the source of this methanol or its fossil carbon content, most commercial methanol is produced from natural gas. Radiocarbon data are generally presented in a few ways: 1) as a “fraction modern”, or the percent of modern carbon in each sample, which is related to the “radiocarbon age”; or 2) as Δ14C, or the ratio of 14 C to 12C in a given sample relative to a standard and corrected for 13 C content (Stuiver and Polach, 1977). The radiocarbon content of background air is slightly elevated above “modern” (~0‰) due to atmospheric nuclear weapons testing in the mid-20th century (Nydal and
120 100
kg-CO2,fossil d-1
±
(for mass flux quantification) and from dissolved N2O, which could discriminate the contributions of nitrifiers and denitrifiers to N2O production.
-0.5
80 60 40 20 0 800 700 600 500
g-N2O d-1
g-CO2eq∙m−3
400 300 200 100 0 -100 Anoxic Reactor (Denitr.)
Aerobic Reactor (Nitr.)
Settling Basin
Fig. 2. The daily GHG flux rates of CH4, fossil CO2, and N2O. The shaded bars represent the average flux ± one standard error, while the vertical lines represent the range of data collected. See Table 1 for the individual data and Table 2 for the summed emissions for the plant.
beginning of the aerobic reactor were likely caused by ammonia buildup, rather than elevated DO (Ahn et al., 2010). Greater emissions of the gas in the early stages of treatment may be somewhat independent of DO levels, though there is some debate (Kampschreur et al., 2008). Yet the quantitative gas flux measurements gathered with a flux chamber cannot discern the contribution to N2O mass flux from nitrification versus denitrification. Future studies will benefit from a combination of stable isotope measurements from samples from the flux chamber
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
Lövseth, 1983). Background air during the study period had a Δ14C of 54‰ (Djuricin et al., 2010). Fossil fuels are generally considered “radiocarbon dead” and have a Δ14C of about − 1000‰. The CO2 in the flux chambers after incubation represents a mixture of background air and CO2 emitted from the surface of the treatment tank. Since the 14C samples were collected after the chambers had been sitting on the surface for several minutes, and the CO2 emission rates were so large, the overwhelming majority of the CO2 in the headspace was assumed to be emitted from wastewater. Radiocarbon content of emitted CO2 is presented in Table 3. The CO2 sampled from the anoxic reactors (denitrification) had an average Δ14C value of −120.5 (±2.1) ‰, which corresponded to an average of 11.4 (±0.2) % from fossil sources. In the aerated reactor (nitrification), the Δ14C of the emitted CO2 was −157.5 (±2.1) ‰ or 15.1 (±0.2) % fossil C. The average Δ14C value in the settling basin was − 126.4 (±2.3) ‰ or 12.0 (±0.2) % fossil. Total CO2 (both modern and fossil) fluxes were highest in the aerated reactor (nitrification) with a rate of 231.7 (±29.8) mg m−2 min−1. Fluxes from the other two reactors were an order of magnitude lower: 14.7 (±1.1) mg m−2 min−1 from the anoxic reactor (denitrification) and 17.4 (± 1.3) mg m− 2 min− 1 from the settling basin. Multiplied out over the course of 24 h, the total CO2 emissions for the anoxic reactor, aerated reactor, and settling basin were 12.1 (±0.9), 190.4 (±24.5), and 14.3 (±1.1) kg-CO2 d−1, respectively (Table 1). Similar to the total CO2 rates, the largest fossil CO2 fluxes per square meter were in the aerated reactors, followed by the settling basin and anoxic reactors (Fig. 2). When extrapolated across the entire surface area of the reactor, the total emission rate in the aerated reactor was 65.1 (±8.4) kg-CO2-fossil d−1. Fossil CO2 emissions from the settling basin and from the anoxic reactor were 9.5 (±0.7) kg-CO2-fossil d−1, and 1.4 (± 0.1) kg-CO2-fossil d− 1. The total emissions of fossil CO2 were 82.1 (±9.1) kg-CO2-fossil d−1, with the aerated reactor accounting for approximately 86% of the flux (Table 2). The total (fossil and modern) average CO2 emissions from the aerated reactor in New Hampshire were approximately 327 g-CO2fossil m−3, two orders of magnitude greater than our observations of 3.9 (±0.4) g-CO2-fossil m−3 (Czepiel et al., 1995). This difference, however, is not as significant as it may seem, as the actual measured CO2 fluctuation rates varied widely across three orders of magnitude in both the previous research and in our work (Fig. 2). Models of CO2 production by wastewater treatment plants have also suggested that in facilities with low BOD5 level (b 300 mg/L), such as at the study site, aerobic processes may emit less GHGs than anaerobic methods (Cakir and Stenstrom, 2005). Thus the direct comparison of flux rates may be complicated by numerous factors. There are few other studies of the carbon isotopic signature of wastewater, and they have focused on either characterizing the presence of anthropogenic carbon throughout the entire wastewater treatment cycle (Law et al., 2013), or the radiocarbon content of DOC (Griffith et al., 2009). In addition, there have been no radiocarbon analyses of the CH4 emitted from wastewater, and it may prove worthwhile to include them in future research. Petroleum-based products are widespread in our society, and many of them find their way into the wastewater collection system. Substances such as cosmetics, detergents, and food additives enter the Table 3 Radiocarbon analysis of the emitted CO2. Sample location
Sampling Δ14C (‰) day
Anoxic reactor (denitrification) Aerobic reactor (nitrification) Settling basin
2 3 2 3 2 3
−109.5 −131.5 −159.3 −155.7 −129.9 −122.9
±
Δ14C Fraction (‰) avr fossil (%)
1.5 −120.5 10.28 1.5 12.50 1.4 −157.5 15.30 1.6 14.94 1.5 −126.4 12.34 1.8 11.63
±
Fraction fossil avr (%)
0.15 11.39 0.15 0.14 15.12 0.16 0.15 11.99 0.18
1171
system through the sewer, and compounds such as methanol are added to wastewater in order to promote bacterial activity. Our measurements indicated that between 10 and 15% of the CO2 emissions from this wastewater treatment plant were derived from fossil sources. As the CO2 emissions from wastewater treatment are currently assumed to be entirely from modern biological origins, this represents an unaccounted GHG source. Our measurements, however, were limited to six samples over two days — future work should attempt to expand these data. Furthermore, this plant employs a somewhat unconventional configuration, using methanol to promote denitrification. In order to extrapolate these results broadly across the industry, similar measurements should be taken from other water reclamation systems. A number of wastewater emissions and gas exchange models could also be employed in future work to attempt to understand the relatively low observed emissions as they correspond to the entire system. As this system is further studied, it would be worthwhile to investigate the source of the differences in 14C emissions between the three basins, and the connections between emitted 14C as CO2 and the dissolved 14C in effluent DOC. 3.5. Net direct GHG emissions Emission flux rates were extrapolated across a 24-hour period using the average flow rate during the study period (64,100 m3 d−1) to calculate the daily and volumetric GHG emissions (Table 2). CH4 and N2O emission rates were converted to CO2 equivalents in accordance with IPCC protocol in order to facilitate comparison and to calculate a total GHG footprint (Forster et al., 2007). Each cubic meter of water processed by the treatment plant emitted approximately 0.5 (±0.1) g-CO2-eq of CH4, 1.5 (±0.2) g-CO2-fossil, and 1.8 (±0.5) g-CO2-eq of N2O (Table 2). Of the emissions measured, 14% of the total CO2-equivalencies were due to CH4, 39% to fossil-CO2, and 47% to N2O. The total direct greenhouse gas footprint of this plant was 3.9 (± 0.5) g-CO2-eq m−3. Multiplied by the average daily flow rate, this plant emits approximately 195.8 (±18.6) kg-CO2-eq d−1 from its secondary treatment system. This is much lower than the total potential downstream GHG emissions if the wastewater is not treated (totaling 0.99 kg-CO2-eq kg-COD− 1; Gori et al., 2011), which for this plant can be estimated at 3.0 × 104 kg-CO2-eq d− 1 for no treatment and 1.7 × 104 kg-CO2-eq d− 1 for primary treatment only. A previous study modeled the indirect greenhouse gas emissions for water supply pathways in Southern California (Stokes and Horvath, 2009). The results of this study (Fig. 3) included CO2 emissions from construction, distribution, power production, and operation. Their data indicate that, employing the current electricity-production methods in California, water importation and wastewater treatment have similar carbon footprints, while seawater desalination has a much greater impact. The carbon footprint for water reclamation amounted to 1023 gCO2-eq m−3 of treated water. Our measurements indicate that operational emissions add an additional 3.9 g-CO2-eq m−3 of treated water (increasing the carbon footprint by 0.38%). This suggests that the direct emissions of GHGs do not significantly increase the carbon footprint of wastewater treatment as a water production method. The flux measurement reported here may be missing a significant portion of N2O emissions, though, as our previous work involving dissolved GHGs indicated that N2O emissions may be up to 141 (±63) g-CO2-eq m− 3 (Townsend-Small et al., 2011). If this were the case, these N2O emissions alone could bring the carbon footprint of recycled water to approximately 1164 g-CO2-eq m−3, a 14% increase above the value calculated. Effluent GHG emissions associated with water recycling may be important considerations and need to be investigated further in order to understand the entire impact of the process. Furthermore, the difference between the measured direct GHG emissions from the plant and the modeled indirect emissions was so great that, even if the emissions had been an order of magnitude greater, they would still not change our assessment significantly. That is, if we were to use some
1172
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
2500
Measured Operational Emissions
1030
LCA Carbon Footprint 1025
g-CO2,eq m-3
g-CO2,eq m-3
2000
1500
1020 1015 1010
1000
1005
500 1000 Recycled Water
0 A
B
C
D
E
Fig. 3. The net carbon footprint of the water recycling process including non-electrical GHG emissions (our data, light gray) and comparison to other water sources (Stokes and Horvath, 2009). Indirect emissions are still the dominant source of GHGs associated with water reclamation, with electricity being the largest single contributor.
of the data collected by others as indicative of this plant, our conclusions would not change. The data suggest that direct greenhouse gas emissions from water reclamation are much smaller than indirect GHG emissions. 4. Conclusions The combined GWP of the direct GHG emissions from this plant, including fossil-CO2, CH4, and N2O amounted to 3.9 g-CO2-eq m−3 of treated water. This was several orders of magnitude less than the calculated indirect GHG emissions from other forms of water production in Southern California. Furthermore, our observations indicate that power generation remains the overwhelming source of fossil GHG emission for all forms of water production. Efforts to control GHG emissions from power sources will reduce the environmental costs of human water use, and should remain a priority. Yet the direct emissions from wastewater treatment plants, albeit small, are not negligible and need to be included in future assessments. In addition, emissions of CO2 derived from fossil sources in wastewater are important to understanding the nature of anthropogenic GHGs. Global population growth, coupled with our expanding knowledge of these direct emissions, means that wastewater treatment will become an ever more critical component of our GHG budget in the future. As wastewater treatment plants are seen increasingly more as resource recovery plants, carbon sequestration will add to their environmental benefits. Acknowledgments Our thanks go out to Mr. Matt Jeung and Dr. Xiaomei Xu of the University of California, Irvine, and Mr. Dave Hayden of the Irvine Ranch Water District. Appendix A. Supplemental data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.10.060. References Ahn JH, Kim S, Park H, Rahm B, Pagilla K, Chandran K. N2O emissions from activated sludge processes, 2008–2009: results of a national monitoring survey in the United States. Environ Sci Technol 2010;44:4505–11.
Beaulieu JJ, Shuster WD, Rebholz JA. Controls on gas transfer velocities in a large river. J Geophys Res Biogeosci 2012;117. Cakir FY, Stenstrom MK. Greenhouse gas production: a comparison between aerobic and anaerobic wastewater treatment technology. Water Res 2005;39:4197–203. Czepiel PM, Crill PM, Harriss RC. Methane emissions from municipal waste-water treatment processes. Environ Sci Technol 1993;27:2472–7. Czepiel P, Crill P, Harriss R. Nitrous-oxide emissions from municipal waste-water treatment. Environ Sci Technol 1995;29:2352–6. Daelman MRJ, van Voorthuizen EM, van Dongen UGJM, Volcke EIP, van Loosdrecht MCM. Methane emission during municipal wastewater treatment. Water Res 2012;46: 3657–70. Djuricin S, Pataki DE, Xu X. A comparison of tracer methods for quantifying CO2 sources in an urban region. J Geophys Res - Atmos 2010;115. Doorn MRJ, Towprayoon S, Vieira SMM, Irving W, Palmer C, Pipatti R, et al. Wastewater treatment and discharge. In: Eggleston S, Buendia L, Miwa K, Ngara T, Tanabe K, editors. 2006 IPCC guidelines for national greenhouse gas inventories. Hayama, Japan: Institute for Global Environmental Studies; 2006. p. 61. El-Fadel M, Massoud M. Methane emissions from wastewater management. Environ Pollut 2001;114:177–85. EPA. Inventory of U.S. greenhouse gas emissions and sinks: 1990–2010. Washington, D.C.: U.S. Environmental Protection Agency; 2012 Foley J, de Haas D, Hartley K, Lant P. Comprehensive life cycle inventories of alternative wastewater treatment systems. Water Res 2010;44:1654–66. Forster P, Ramaswamy V, Artaxo P, Berntsen T, Betts R, Fahey DW, et al. Changes in atmospheric constituents and in radiative forcing. In: Solomon S, Qin D, Manning M, Chen Z, Marquis M, Averyt KB, et al, editors. Climate change 2007: the physical science basis. Contribution of working group I to the fourth assessment report of the intergovernmental panel on climate change. New York: Cambridge University Press; 2007. Gori R, Jiang L-M, Sobhani R, Rosso D. Effects of soluble and particulate substrate on the carbon and energy footprint of wastewater treatment processes. Water Res 2011; 45:5858–72. Griffith DR, Barnes RT, Raymond PA. Inputs of fossil carbon from wastewater treatment plants to US rivers and oceans. Environ Sci Technol 2009;43:5647–51. Kampschreur MJ, van der Star WRL, Wielders HA, Mulder JW, Jetten MSM, van Loosdrecht MCM. Dynamics of nitric oxide and nitrous oxide emission during full-scale reject water treatment. Water Res 2008;42:812–26. Law Y, Jacobsen GE, Smith AM, Yuan Z, Lant P. Fossil organic carbon in wastewater and its fate in treatment plants. Water Res 2013;47:5270–81. Nydal R, Lövseth K. Tracing bomb 14C in the atmosphere 1962–1980. J Geophys Res 1983; 88. Rosso D, Stenstrom MK. The carbon-sequestration potential of municipal wastewater treatment. Chemosphere 2008;70:1468–75. Stokes JR, Horvath A. Energy and air emission effects of water supply. Environ Sci Technol 2009;43:2680–7. Stuiver M, Polach HA. Reporting of C-14 data — discussion. Radiocarbon 1977;19:355–63. Teiter S, Mander Ü. Emission of N2O, N2, CH4, and CO2 from constructed wetlands for wastewater treatment and from riparian buffer zones. Ecol Eng 2005;25:528–41. Townsend-Small A, Pataki DE, Tseng LY, Tsai C-Y, Rosso D. Nitrous oxide emissions from wastewater treatment and water reclamation plants in Southern California. J Environ Qual 2011;40:1542–50. Townsend-Small A, Tyler SC, Pataki DE, Xu X, Christensen LE. Isotopic measurements of atmospheric methane in Los Angeles, California, USA: influence of “fugitive” fossil fuel emissions. J Geophys Res - Atmos 2012;117.
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173 USDA. Chamber-based trace gas flux measurement; USDA-ARS GRACEnet protocols; 2010 [Washington, D.C.]. Wang J, Zhang J, Xie H, Qi P, Ren Y, Hu Z. Methane emissions from a full-scale A/A/O wastewater treatment plant. Bioresour Technol 2011;102:5479–85. Wunch D, Wennberg PO, Toon GC, Keppel-Aleks G, Yavin YG. Emissions of greenhouse gases from a North American megacity. Geophys Res Lett 2009;36.
1173
Xu X, Trumbore SE, Zheng S, Southon JR, McDuffee KE, Luttgen M, et al. Modifying a sealed tube zinc reduction method for preparation of AMS graphite targets: reducing background and attaining high precision. Nucl Inst Methods Phys Res B-Beam Interact Mater Atoms 2007;259:320–9.
Contents lists available at ScienceDirect
Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv
Impact of direct greenhouse gas emissions on the carbon footprint of water reclamation processes employing nitrification–denitrification Andrew G. Schneider a,⁎, Amy Townsend-Small a,b, Diego Rosso c a b c
University of Cincinnati, Department of Geology, Cincinnati, OH 45221, United States University of Cincinnati, Department of Geography, Cincinnati, OH 45221, United States Department of Civil and Environmental Engineering, University of California, Irvine, CA 92697-2175, United States
H I G H L I G H T S • Direct greenhouse gas emissions were measured at a wastewater reclamation plant. • These greenhouse gas emissions amounted to 3.9 (±0.5) g-CO2-eq m− 3 of wastewater. • 14C analysis of the CO2 emissions was conducted to determine the fossil component. • 11.4% to 15.1% of the emitted CO2 was derived from fossil sources.
a r t i c l e
i n f o
Article history: Received 20 April 2014 Received in revised form 5 October 2014 Accepted 19 October 2014 Available online 14 November 2014 Editor: Simon Pollard Keywords: Water reclamation Greenhouse gases Direct emissions Radiocarbon Carbon footprint Fossil carbon
a b s t r a c t Water reclamation has the potential to reduce water supply demands from aquifers and more energy-intensive water production methods (e.g., seawater desalination). However, water reclamation via biological nitrification– denitrification is also associated with the direct emission of the greenhouse gases (GHGs) CO2, N2O, and CH4. We quantified these direct emissions from the nitrification–denitrification reactors of a water reclamation plant in Southern California, and measured the 14C content of the CO2 to distinguish between short- and long-lived carbon. The total emissions were 1.5 (±0.2) g-fossil CO2 m−3 of wastewater treated, 0.5 (±0.1) g-CO2-eq of CH4 m−3, and 1.8 (± 0.5) g-CO2-eq of N2O m−3, for a total of 3.9 (± 0.5) g-CO2-eq m−3. This demonstrated that water reclamation can be a source of GHGs from long lived carbon, and thus a candidate for GHG reduction credit. From the 14C measurements, we found that between 11.4% and 15.1% of the CO2 directly emitted was derived from fossil sources, which challenges past assumptions that the direct CO2 emissions from water reclamation contain only modern carbon. A comparison of our direct emission measurements with estimates of indirect emissions from several water production methods, however, showed that the direct emissions from water reclamation constitute only a small fraction of the plant's total GHG footprint. © 2014 Elsevier B.V. All rights reserved.
1. Introduction Increases in atmospheric greenhouse gas (GHG) concentrations are linked to human activities (Forster et al., 2007). Waste treatment and storage are global sources of greenhouse gases, particularly nitrous oxide (N2O) and methane (CH4) (Ahn et al., 2010; Daelman et al., 2012). In a wastewater treatment system, the microbial processes of nitrification and denitrification produce N2O (Ahn et al., 2010; Townsend-Small et al., 2011), a GHG that has a global warming potential 298 times that of carbon dioxide (CO2) over a 100-year horizon (Forster et al., 2007). CH4 is produced by the decomposition of organic matter under the anaerobic conditions found in several points of the wastewater treatment process, and has a global warming potential 25 ⁎ Corresponding author. Tel.: +1 203 785 2912. E-mail address: [email protected] (A.G. Schneider).
http://dx.doi.org/10.1016/j.scitotenv.2014.10.060 0048-9697/© 2014 Elsevier B.V. All rights reserved.
times that of CO2 during a 100-year horizon (Forster et al., 2007). CO2 is produced directly in wastewater treatment via the aerobic decomposition of organic matter, as well as indirectly through chemical manufacturing and the combustion of fossil fuels for electricity production and transportation. Annually, wastewater treatment in the United States emits approximately 7.79 × 108 kg of CH4 (2.5% of total American CH4 emissions) and 1.63 × 107 kg of N2O (1.6% of total American N2O emissions) (EPA, 2012). Akin to numerous urban and suburban conglomerates across the United States and abroad, Southern California has undergone rapid urban growth, economic progress, and population expansion over the past few decades. Groundwater alone cannot meet the water demands of the region due to the low recharge rate and threat of saltwater intrusion into the coastal aquifers. Long-distance importation is the largest source of potable water to the Los Angeles and San Diego areas, though the electricity required for importation is a major source of CO2 (Stokes
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
and Horvath, 2009). In order to reduce the impact of importation on the upstream regions, desalination and water reclamation (the extension of wastewater treatment with provisions for beneficial reuse) are also being pursued. The process of water reclamation purifies and recycles the water that is already in the collection system, reducing the demands placed on aquifers and reservoirs. Although water reclamation may relieve some of the stress on aquifers and reduce the energy required (and therefore the CO2 emitted) for long-distance importation or desalination, these changes may lead to an increase in direct N2O and CH4 emissions from biological treatment. Studies have determined that water importation and water reclamation have comparable indirect emissions of greenhouse gases, about 1 kg-CO2-eq m−3 of water, with sea-water desalination having a much greater energy demand, resulting in about 2.4 times more CO2 equivalents m−3 than water reclamation (Stokes and Horvath, 2009). It was estimated that the CH 4 and N 2 O emissions from American domestic and industrial wastewater treatment in 2010 totaled 2.13 × 1010 kg-CO2-eq (EPA, 2012). The annual GHG emissions in the Southern California air basin include 6 (±1) × 108 kg-CH4 y−1 (Wunch et al., 2009), and our previous work has confirmed that wastewater treatment emits CH4 in Los Angeles (Townsend-Small et al., 2012). Furthermore, our research in the region has shown that, although water reclamation is a minor source overall, it may be the principal point-source of N2O, emitting 7.6 × 104 kg-N2O y− 1 (2.26 × 107 kgCO 2 -eq y − 1 ) in Orange County (area = 2455 km 2 ; 2010 population = 3 million) alone (Townsend-Small et al., 2011). More broadly, wastewater treatment is not unique to the developed world, and urbanization, economic development, and population growth may result in CH4 and CO2 emissions on the order of 107 kg-CO2-eq y− 1 from wastewater systems in emerging nations by 2025 (Rosso and Stenstrom, 2008). Yet while wastewater treatment and water reclamation may be associated with some greenhouse gas emissions, it is essential to remember that these processes are designed to foster environmental and human health and that the scenario of no treatment is characterized by high emissions with no biogas energy recovery. One previous study measured both the direct CH4 and N2O emissions from a small wastewater treatment plant in New Hampshire, providing some of the first published data on the process (Czepiel et al., 1993, 1995). More recently, Ahn et al. conducted a broad study of N2O emissions from plants across the United States (Ahn et al., 2010). Wang et al. (2011) and Daelman et al. (2012) conducted measurements of CH4 from wastewater treatment in China and Europe, respectively. Direct emissions of CO2 during the treatment process are often thought to be negligible or fully biogenic (i.e., short-lived carbon) in nature (Rosso and Stenstrom, 2008). Most researchers acknowledge that CO2 is emitted as part of the water treatment and recycling process, but do not measure the fossil component of the CO2, with the exception of few cases (e.g., Griffith et al., 2009; Law et al., 2013). The goal of this study was to estimate direct greenhouse gas emissions from water reclamation in order to determine whether water reclamation has a larger carbon footprint than long-distance water importation. The three objectives of our study were: 1) to measure the emission rates of CO2, CH4, and N2O from water reclamation; 2) to identify the proportion of the emitted CO2 that was derived from fossil sources using 14C analysis; and 3) to compare the direct emissions of GHGs from water reclamation during operation to the indirect GHGs associated with the process. 2. Materials and methods 2.1. Field site description Samples were collected from a wastewater reclamation plant in Southern California in the early spring. The plant reclaims water employing a modified Ludzack–Ettinger (MLE) activated sludge system
1167
for nitrification–denitrification with supplemental methanol addition. During the period of our sampling, the flow rate ranged from 34,800 m3 d−1 to 75,000 m3 d−1, averaging 64,100 m3 d−1. In general, the mean cell retention time (MCRT) varies between 6 and 10 days through the seasons to compensate for temperature variations, and two on-site flow-equalization basins are used to abate hydraulic peaks across the diurnal cycle. During the period of testing, the average raw influent contained approximately 235 mg L−1 of total suspended solids (TSS), with a 201 mg L−1 of biochemical oxygen demand (BOD5) and 20.6 mg-NH4–N L−1. This plant received municipal wastewater with negligible industrial contribution, therefore the ratio between BOD5 and the chemical oxygen demand (COD) is expected to be aligned with municipal values for separate sewer systems in areas with food waste disposal to drain (i.e., 0.40–0.45). The average mixed liquor suspended solid (MLSS) was 2500 mg/L, with an average volatile suspended solids (VSS) fraction of 79%. Primary treatment reduced this to 102 mg-TSS L− 1 and 114 mgBOD5 L−1, while bypassing the 20.6 mg-NH4–N L−1 to the secondary reactors. Effluent water from the plant, after undergoing secondary treatment, tertiary filtration and disinfection, contained a maximum of 1 mg-TSS L−1, 2 mg-BOD5 L−1, and 0.5 mg-NH4–N L−1. Depending on demand, roughly half of the effluent enters a recycled-water distribution system and is used for irrigation of agriculture and landscaping, along with other industrial and commercial purposes (i.e., CA Title 22 water). The remaining effluent is sent to reservoirs. Our sampling was conducted in three of the secondary treatment reactors: the anoxic (denitrification) and aerobic (nitrification) reactors, and the settling basins (illustrated in Fig. 1), as previous research indicated that these sites were the sources of the majority of the on-site GHG emissions (Townsend-Small et al., 2011). There are six anoxic reactors in parallel, with a total volume of approximately 4300 m3. Each anoxic reactor is followed by aerobic tanks approximately double in volume, with a combined aerobic volume of 12,900 m3. Treated wastewater from the aerobic reactors is recombined and distributed to the settling basins (nine rectangular, one circular), with a combined volume of 11,500 m3. The dissolved oxygen (DO) set point of the aerobic tanks is 2.4 mg/L, while the DO set point for the settling basins is 2.7 mg/L.
Fig. 1. A flow diagram of the secondary treatment reactors with our approximate sampling locations (circled X): A) anoxic reactors (denitrification), B) aerated reactors (nitrification and carbon oxidation), and C) secondary clarifiers (settling basins). The single circular settling basin is physically separated from the nine parallel rectangular ones, but receives the same flow and has the same retention time. Wastewater enters this section of the system from the primary clarifiers and flow equalization basins, and exits to the tertiary filters and disinfection contact basins. The dashed lines between the anoxic and aerated reactors are baffles, used to isolate the anoxic and aerobic reactors. Dotted lines represent internal circulation.
1168
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
2.2. Sampling method
3. Results and discussion
Greenhouse gas fluxes were measured twice a day for five days in the early spring season from the surface of the reactors using three floating chambers (Beaulieu et al., 2012). The Plexiglas chambers had an interior volume of 7.73 L and a sectional area of 0.067 m2 when afloat. They were constructed following U.S. Department of Agriculture guidelines for trace gas measurements (USDA, 2010). A flexible tube (6.35 mm interior diameter) with a stopcock was attached to the chamber for sampling, with another opening on the chamber for pressure compensation. Due to the 1 m freeboard sheltering the surface of these reactors from the wind, the high off-gas flow rate, and given that the wind speed during testing was negligible, the effect of windinduced pressure gradients was deemed insignificant. In an attempt to capture the differences between tanks in each process while retaining some redundancy, two of the chambers were placed in the same well-mixed reactor (equipped with an active subsurface mixer) and one in a contiguous reactor train during each period of measurement (marked in Fig. 1). Sample gas extraction was begun immediately after placing the chamber on the surface in order to measure the rate of GHG concentration change in the headspace. Before each sample was collected, 200 mL of gas was drawn from each chamber to clear the tubing. Once the tubing was cleared, 30 mL of sample gas was extracted and stored in pre-evacuated 20 mL glass vials sealed with gray rubber butyl septa and aluminum crimps. Samples for radiocarbon analysis of CO2 were collected using evacuated 5 L stainless steel canisters. These samples were collected during the afternoon of Day 2 and during the morning of Day 3, with one sample from each of the three reactors being measured per day. The gases for radiocarbon analysis were drawn from the floating chambers at the end of the sampling interval in order to allow the greatest concentration of CO2 to accumulate in the chamber headspace.
3.1. Flux rates
2.3. Sample analysis Greenhouse gas flux samples were analyzed using a Shimadzu GC2014 Greenhouse Gas Analyzer, with flame-ionization detector for measurement of CH4 and CO2 and an electron-capture detector to measure N2O. A least-squares linear regression method comparing the sampled gas concentrations against the time of sampling was employed to calculate emission rate for the three gases in each chamber (USDA, 2010). These flux rates were then used to extrapolate the total flux of CO2, CH4, and N2O for the plant's secondary treatment basins with the formula, f V C P A M SB min −6 L 10 1440 day μL AC R T Q D
f VC PA M SB AC R T QD
flux rate (ppm min−1) volume of sampling chamber (L) atmospheric pressure (atm) molar mass of the compound (g mol−1) surface area of the basin (m2) surface area of sampling chamber (m2) gas law constant (0.08206 L atm K−1 mol−1) air temperature above basin (K) daily flow rate (m3 d−1).
The results were converted into CO2 equivalents via the IPCC method (Forster et al., 2007) using the 100-year global warming potential (GWP) for N2O and CH4. Samples for radiocarbon analysis were converted to graphite and analyzed for 14C content by accelerator mass spectrometry (Xu et al., 2007).
We measured the surface emissions of GHGs, which are presented here initially in units of mass per surface area per time. CH4 flux rates were 0.13 (±0.02) mg-CH4 m−2 min−1 in the anoxic reactor, decreasing to 0.08 (± 0.04) mg-CH4 m− 2 min− 1 in the aerated reactor, and increasing to 0.18 (± 0.03) mg-CH4 m−2 min−1 in the settling basin (all numbers are ±1 SE). The total CO2 emission rates fluctuated from 14.7 (± 1.1) mg-CO2 m−2 min− 1 up to 231.7 (± 29.8) mg-CO2 m− 2 min− 1, and back down to 17.4 (± 1.3) mg-CO2 m− 2 min−1 through the three reactors. Radiocarbon data indicated that 11.4 (± 0.2) % of the CO2 flux from the anoxic reactor, 15.1 (± 0.2) % of the CO2 from the aerated reactor, and 12.0 (±0.2) % of the CO2 flux from the settling basin originated from fossil carbon. N2O followed a similar pattern as CO2, with an emission rate of 6.1 (± 0.5) μg-N2O m−2 min−1 in the anoxic reactor (denitrification), increasing an order of magnitude to 84.2 (±19.3) μg-N2O m−2 min−1 in the aerated reactor (nitrification), and dropping back down to 0.009 (± 0.001) μg-N2O m−2 min− 1 in the settling basin. (See supplemental information for specific flux measurements.) 3.2. Methane emissions The areal CH4 flux rate was greatest in the settling basin and lowest in the aerated reactor (performing carbon oxidation and nitrification). CH4 emissions from the aerobic tank were expected to drop because of the inhibition to CH4 formation and the CH4 oxidation occurring in the aerobic conditions. In addition, the settling basins had a surface area nearly six times the size of the anoxic reactor (performing denitrification) and about twice the size of the aerated reactor. In the settling basin, this elevated emission rate and large surface area resulted in a total average emission of 859 (± 147) g-CH4 d−1 (Table 1, Fig. 2), or 74% of the total CH4 emissions measured. In comparison, the anoxic and aerated reactor emission rates were 106 (±13) g-CH4 d−1 (9% of total) and 196 (± 105) g-CH4 d− 1 (17% of total), respectively. The large variability in the emissions rates from the aerobic reactor for CH4 and other gases was potentially due to uneven aeration during the relatively short interval of sample collection. Methanogenesis from biodegradation occurs under anaerobic conditions and is controlled by the amount of organic matter in the wastewater (El-Fadel and Massoud, 2001; Cakir and Stenstrom, 2005; EPA, 2012). The sludge blanket at the bottom of the settling basin had low or no (DO) and high organic matter content, enabling CH4 production. It is unlikely that significant amounts of CH4 were produced in the nitrification reactor due to the inhibition to methanogens and the promoted methane oxidation caused by oxygenation, so we assume that the CH4 emitted from the aerated reactor was from the stripping of CH4 produced in the (anoxic) denitrification reactor. The CH4 emitted by secondary treatment cannot be captured for energy production due to its elevated dilution with other atmospheric gases in the off-gas. In other locations where off-gas is used as feed air for flares or combined heat and power generation systems, however, the oxidation of this CH4 to CO2 is achieved. The total CH4 emission rate from the entire secondary treatment system during the period of sampling was 22.0 (± 3.7) mg-CH4 m−3 (Table 2, units are per cubic meter of wastewater). Other researchers (Foley et al., 2010) have attempted to model the emission rate of water reclamation plants that employ a modified Ludzack–Ettinger activated sludge system, though their results were significantly higher (610 mg-CH4 m−3). The difference may be due to the elevated DO set points selected for this plant (2.4–2.7 mg-O2 L− 1), which promote thorough oxygenation of the wastewater during aeration and may help inhibit anaerobic activity inside bacterial flocs outside the oxic reactors. Measurements of emissions from aerated reactors in other types of wastewater processes have also
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
1169
Table 1 Average measured GHG flux rates from the three reactors. Anoxic reactor (denitrification) CH4
n Mean SEMean
28 0.11 0.01
CO2
Aerobic reactor (nitrification) N2O
Total
Fossil
27 12.1 0.9
27 1.37 0.11
26 5.0 0.4
CH4
19 0.20 0.11
CO2
Settling basin N2O
Total
Fossil
19 190.4 24.5
19 65.05 8.37
18 207.6 47.7
CH4
25 0.86 0.15
CO2
N2O
Total
Fossil
27 14.3 1.1
27 9.50 0.70
26 42.9 5.7
n = number of observations, mean = arithmetic mean, SEMean = standard error of the mean. CH4 values in kg-CH4 d−1, CO2 values in kg-CO2 d−1, N2O values in g-N2O d−1.
found annual CH4 fluxes higher (0.2 g-CH4 m−3) (Czepiel et al., 1993) than those found in the current study (0.02 g-CH4 m−3). Also, the facility studied here had no sludge processing on site; hence the return flow (i.e., supernatant and centrate or filtrate, rich in COD and ammonia) was not part of our measurements. This may imply that the separate treatment of return flows (by means, for example, of sidestream ammonification) may severely reduce the emission from the mainstream train, albeit compensating the total sum with their own emission. Future studies should focus on separate processes for mainstream and sidestream treatment, so as to compare the total emission balance with traditional, combined treatment. Most other reported values of CH4 from both physical sampling and modeling are an order of magnitude higher than those observed in the current study (Czepiel et al., 1993; Daelman et al., 2012; Wang et al., 2011). The plant influent had an average of 201 g-BOD5 m−3, which indicated that the CH4 emission rate was 0.11 (±0.02) g-CH4 kg-BOD−1 5 [1.16 (±0.21) g-CO2-eq kg-COD−1], or 0.011% of the BOD5 (the average carbon emission intensity of untreated influent COD is 0.99 kg-CO2eq kg-COD−1; Gori et al., 2011). This ratio, which can be interpreted as a relative assessment of aerobic and anaerobic digestion, was lower than those measured by others, 0.16% of influent BOD5 in New Hampshire (Czepiel et al., 1993) and 0.08% of influent COD in China (Wang et al., 2011). Comparisons with other research findings are complex because of the different systems employed and the lack of sludge processing and return flows at this plant. Another factor that may play a role in the reduction of emission fluxes here is the unusually high depth of the biological reactors (approximately 8.5 m), which may offer a chance for produced CH4 in the water phase to be recaptured and oxidized by methanotrophs before reaching the surface. 3.3. Nitrous oxide emissions Per square meter, the N2O flux rate was highest in the aerated reactor and lowest in the anoxic reactor (denitrification). When the emission rates were extrapolated across the surface areas of the three reactors, the aerated reactor was the largest source of N2O. It emitted 207.6 (±47.7) g-N2O d−1 (approximately 81% of the total), compared to 5.0 (± 0.4) g-N2O d−1 from the anoxic reactor (2%) and 42.9 (± 5.7) g-N2O d− 1 from the settling basin (17%) (Table 1). Others have observed similarly elevated emissions in the aerated reactor, and have proposed several mechanisms that may result in these high rates, two of which were possible here: increased air-stripping of N2O, and N2O production by nitrification during variations in the aeration rate linked to influent (i.e. ammonia) buildup (Ahn et al., 2010). We were unable to differentiate between these two possibilities based on our measurements. Our previous work, however, showed that N2O emitted from the aerated reactor had an N isotope signature consistent with nitrification and distinct from N2O emitted from the denitrification reactor (Townsend-Small et al., 2011), indicating that the second proposed mechanism was most likely. The average N2O emission factor from the secondary treatment was 0.03 (±0.01) % of the influent nitrogen load. This put the N2O emissions near the low end of other previously observed N2O emission fractions of 0.03 to 0.11% from similar types of water reclamation plants (Ahn et al.,
2010). In addition, the N2O emissions from the aerated reactor averaged 4.8 (±1.5) mg-N2O m−3, well below the range of emissions previously observed (31 mg-N2O m−3) at a wastewater treatment plant in New Hampshire (Czepiel et al., 1995). Previous work at this plant in Southern California (Townsend-Small et al., 2011) observed that the dissolved N2O emission factor was an order of magnitude higher than the emissions from the New Hampshire plant. These discrepancies in results from the same plant were in concordance with an analysis of the importance of sampling strategy for accurate flux estimation and comparison, reported elsewhere (Daelman et al., 2012). Since our ultimate purpose here was not to compile a detailed flux inventory of one specific gas but rather to characterize the broader greenhouse gas makeup, the high-frequency analysis recommended in Daelman et al. was not employed here. The measured N2O flux from the secondary treatment system was 6.2 (±1.6) mg-N2O m−3 (Table 2). This emission rate represents only a small fraction of the N2O that remained dissolved in the fluid. Our previous work found an average of 271.7 mg-N2O m−3 dissolved in the anoxic reactor, 108.1 mg-N2O m−3 dissolved in the aerobic reactor, and 11.5 mg-N2O m−3 dissolved in the settling basin. In this current study, the measured emission rates were much smaller, 0.08 (±0.1) mg-N2O m−3 from the anoxic reactor, 4.7 (±1.4) mg-N2O m−3 from the aerated reactor, and 0.9 (±0.2) mg-N2O m−3 from the settling basin. Our measured emissions represented 0.03%, 4.4%, and 7.6% of the dissolved N2O in each reactor, respectively. These two sets of data were collected about two years apart, though the processing conditions were not significantly modified during the interim period. Previous work assessed dissolved N2O concentrations in each reactor and assumed that the bulk of this was emitted to the atmosphere eventually (Townsend-Small et al., 2011). The current study suggests that the N2O present in the treatment tanks is either consumed by denitrifiers, emitted to the atmosphere later in the process, or both. Although possibly consequential, downstream emissions of N2O from the effluent were not considered in this investigation. It may also be important to note that in some wastewater treatment methods, the overwhelming majority of gaseous nitrogen release may be in the form of N2–N (Teiter and Mander, 2005). There was a significantly greater proportion of dissolved N2O emitted from the aerobic reactor than from the anoxic reactor, though this was still only a fraction of total dissolved N2O (4.4%). A higher percentage (7.6%) of the total dissolved N2O was emitted from the settling basin, yet by this point in the process the amount of total dissolved gas had dropped considerably. The transient high N2O concentration in the anoxic reactors is maintained by a combination of gentle deep mixing (i.e., no mechanical device shearing the water surface and promoting stripping) and a lack of volumetric stripping through aeration. These high-concentration conditions persist throughout the anoxic reactors but are heavily altered as soon as the liquid enters the aerobic reactor, i.e. where the volumetric air flow strips the excess N2O from the liquid phase. The anoxic reactor was the site of the highest measured N2O production, though most of the gas remained dissolved. Some N2O was also produced in the aerated reactor (during nitrification), much of which was stripped out of solution and emitted into the atmosphere. Dissolved oxygen levels can also play a role in the emission of N2O, though the increased emissions that occurred at the
1170
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
3.5
Table 2 Total direct GHG emissions across all three basins, assuming a normal flow of 69,300 m3 of wastewater per day.
3.0 2.5
g∙m−3 ±
kg-CH4 d-1
2.0 CH4 Fossil CO2 N2 O Total
1.5 1.0 0.5 0.0
kg∙d−1 ±
kg-CO2eq∙d−1
±
0.022 1.507
0.004 0.549 0.246 1.507
0.092 1.26 0.246 82.12
0.20 9.09
31.4 82.1
4.9 9.1
0.006
0.002 1.828 3.885
0.468 0.536
0.05
82.3 195.8
15.5 18.6
0.28
-1.0 -1.5
3.4. Fossil carbon dioxide emissions
140
In the aerated reactor, a constant supply of atmospheric oxygen supports the aerobic decomposition of biological waste, resulting in the production of CO2 through bacterial aerobic respiration. Our observations fit this principle well, with the poorly-mixed anoxic reactor emitting the least amount of CO2 (Table 1) and the well-oxygenated aerated reactors emitting the largest amount of CO2. The direct emissions of CO2 from wastewater are assumed to derive from modern organic matter in human excreta or food waste, and such emissions of recently-photosynthesized carbon are believed to play an insignificant role in climate change (EPA, 2012). This “modern” organic carbon is part of an active, short-term cycle in which it is processed through the environment as recent biological matter, waste, soil carbon, atmospheric gases, etc. over relatively short time scales (generally b106 years, though the scale can vary dramatically). Regardless of human intervention, this “modern” organic carbon would be present in the atmosphere. In comparison, “fossil” organic carbon is that which has been removed from the active, short-term, atmospheric cycle. This carbon forms fossil fuels (coal, petroleum, natural gas) with turnover timescales exceeding 106 years. The distinction of “anthropogenic” greenhouse gas CO2 is caused by human activity that altered this timescale, exhuming the “fossil” carbon and injecting it into the short-term cycle faster than it can be removed. Neither the U.S. Environmental Protection Agency (EPA) nor the Intergovernmental Panel on Climate Change (IPCC) considers wastewater treatment to be a significant source of CO2 for greenhouse gas accounting (Doorn et al., 2006; EPA, 2012). The fossil-fuel derived carbon in wastewater may be a relevant portion of total C processed in treatment, however, giving decision makers, operators, and utility managers an incentive for sequestration with the associated credits. One previous study conducted in the Northeastern United States found that approximately 25% of wastewater dissolved organic carbon (DOC) is fossil carbon, most likely derived from cleaning products, pharmaceuticals, and other petroleum-based products (Griffith et al., 2009). Another, conducted in Australia, concluded similarly that 4–14% of the TOC is of fossil origin (Law et al., 2013). Some water reclamation plants supplement the readily biodegradable carbon in the primary effluent with methanol to provide a labile substrate for denitrification (Townsend-Small et al., 2011). The plant under study added about 1.6 m3 of methanol a day to the anoxic reactors, roughly 24 μL of methanol per liter of wastewater. Although we do not know the source of this methanol or its fossil carbon content, most commercial methanol is produced from natural gas. Radiocarbon data are generally presented in a few ways: 1) as a “fraction modern”, or the percent of modern carbon in each sample, which is related to the “radiocarbon age”; or 2) as Δ14C, or the ratio of 14 C to 12C in a given sample relative to a standard and corrected for 13 C content (Stuiver and Polach, 1977). The radiocarbon content of background air is slightly elevated above “modern” (~0‰) due to atmospheric nuclear weapons testing in the mid-20th century (Nydal and
120 100
kg-CO2,fossil d-1
±
(for mass flux quantification) and from dissolved N2O, which could discriminate the contributions of nitrifiers and denitrifiers to N2O production.
-0.5
80 60 40 20 0 800 700 600 500
g-N2O d-1
g-CO2eq∙m−3
400 300 200 100 0 -100 Anoxic Reactor (Denitr.)
Aerobic Reactor (Nitr.)
Settling Basin
Fig. 2. The daily GHG flux rates of CH4, fossil CO2, and N2O. The shaded bars represent the average flux ± one standard error, while the vertical lines represent the range of data collected. See Table 1 for the individual data and Table 2 for the summed emissions for the plant.
beginning of the aerobic reactor were likely caused by ammonia buildup, rather than elevated DO (Ahn et al., 2010). Greater emissions of the gas in the early stages of treatment may be somewhat independent of DO levels, though there is some debate (Kampschreur et al., 2008). Yet the quantitative gas flux measurements gathered with a flux chamber cannot discern the contribution to N2O mass flux from nitrification versus denitrification. Future studies will benefit from a combination of stable isotope measurements from samples from the flux chamber
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
Lövseth, 1983). Background air during the study period had a Δ14C of 54‰ (Djuricin et al., 2010). Fossil fuels are generally considered “radiocarbon dead” and have a Δ14C of about − 1000‰. The CO2 in the flux chambers after incubation represents a mixture of background air and CO2 emitted from the surface of the treatment tank. Since the 14C samples were collected after the chambers had been sitting on the surface for several minutes, and the CO2 emission rates were so large, the overwhelming majority of the CO2 in the headspace was assumed to be emitted from wastewater. Radiocarbon content of emitted CO2 is presented in Table 3. The CO2 sampled from the anoxic reactors (denitrification) had an average Δ14C value of −120.5 (±2.1) ‰, which corresponded to an average of 11.4 (±0.2) % from fossil sources. In the aerated reactor (nitrification), the Δ14C of the emitted CO2 was −157.5 (±2.1) ‰ or 15.1 (±0.2) % fossil C. The average Δ14C value in the settling basin was − 126.4 (±2.3) ‰ or 12.0 (±0.2) % fossil. Total CO2 (both modern and fossil) fluxes were highest in the aerated reactor (nitrification) with a rate of 231.7 (±29.8) mg m−2 min−1. Fluxes from the other two reactors were an order of magnitude lower: 14.7 (±1.1) mg m−2 min−1 from the anoxic reactor (denitrification) and 17.4 (± 1.3) mg m− 2 min− 1 from the settling basin. Multiplied out over the course of 24 h, the total CO2 emissions for the anoxic reactor, aerated reactor, and settling basin were 12.1 (±0.9), 190.4 (±24.5), and 14.3 (±1.1) kg-CO2 d−1, respectively (Table 1). Similar to the total CO2 rates, the largest fossil CO2 fluxes per square meter were in the aerated reactors, followed by the settling basin and anoxic reactors (Fig. 2). When extrapolated across the entire surface area of the reactor, the total emission rate in the aerated reactor was 65.1 (±8.4) kg-CO2-fossil d−1. Fossil CO2 emissions from the settling basin and from the anoxic reactor were 9.5 (±0.7) kg-CO2-fossil d−1, and 1.4 (± 0.1) kg-CO2-fossil d− 1. The total emissions of fossil CO2 were 82.1 (±9.1) kg-CO2-fossil d−1, with the aerated reactor accounting for approximately 86% of the flux (Table 2). The total (fossil and modern) average CO2 emissions from the aerated reactor in New Hampshire were approximately 327 g-CO2fossil m−3, two orders of magnitude greater than our observations of 3.9 (±0.4) g-CO2-fossil m−3 (Czepiel et al., 1995). This difference, however, is not as significant as it may seem, as the actual measured CO2 fluctuation rates varied widely across three orders of magnitude in both the previous research and in our work (Fig. 2). Models of CO2 production by wastewater treatment plants have also suggested that in facilities with low BOD5 level (b 300 mg/L), such as at the study site, aerobic processes may emit less GHGs than anaerobic methods (Cakir and Stenstrom, 2005). Thus the direct comparison of flux rates may be complicated by numerous factors. There are few other studies of the carbon isotopic signature of wastewater, and they have focused on either characterizing the presence of anthropogenic carbon throughout the entire wastewater treatment cycle (Law et al., 2013), or the radiocarbon content of DOC (Griffith et al., 2009). In addition, there have been no radiocarbon analyses of the CH4 emitted from wastewater, and it may prove worthwhile to include them in future research. Petroleum-based products are widespread in our society, and many of them find their way into the wastewater collection system. Substances such as cosmetics, detergents, and food additives enter the Table 3 Radiocarbon analysis of the emitted CO2. Sample location
Sampling Δ14C (‰) day
Anoxic reactor (denitrification) Aerobic reactor (nitrification) Settling basin
2 3 2 3 2 3
−109.5 −131.5 −159.3 −155.7 −129.9 −122.9
±
Δ14C Fraction (‰) avr fossil (%)
1.5 −120.5 10.28 1.5 12.50 1.4 −157.5 15.30 1.6 14.94 1.5 −126.4 12.34 1.8 11.63
±
Fraction fossil avr (%)
0.15 11.39 0.15 0.14 15.12 0.16 0.15 11.99 0.18
1171
system through the sewer, and compounds such as methanol are added to wastewater in order to promote bacterial activity. Our measurements indicated that between 10 and 15% of the CO2 emissions from this wastewater treatment plant were derived from fossil sources. As the CO2 emissions from wastewater treatment are currently assumed to be entirely from modern biological origins, this represents an unaccounted GHG source. Our measurements, however, were limited to six samples over two days — future work should attempt to expand these data. Furthermore, this plant employs a somewhat unconventional configuration, using methanol to promote denitrification. In order to extrapolate these results broadly across the industry, similar measurements should be taken from other water reclamation systems. A number of wastewater emissions and gas exchange models could also be employed in future work to attempt to understand the relatively low observed emissions as they correspond to the entire system. As this system is further studied, it would be worthwhile to investigate the source of the differences in 14C emissions between the three basins, and the connections between emitted 14C as CO2 and the dissolved 14C in effluent DOC. 3.5. Net direct GHG emissions Emission flux rates were extrapolated across a 24-hour period using the average flow rate during the study period (64,100 m3 d−1) to calculate the daily and volumetric GHG emissions (Table 2). CH4 and N2O emission rates were converted to CO2 equivalents in accordance with IPCC protocol in order to facilitate comparison and to calculate a total GHG footprint (Forster et al., 2007). Each cubic meter of water processed by the treatment plant emitted approximately 0.5 (±0.1) g-CO2-eq of CH4, 1.5 (±0.2) g-CO2-fossil, and 1.8 (±0.5) g-CO2-eq of N2O (Table 2). Of the emissions measured, 14% of the total CO2-equivalencies were due to CH4, 39% to fossil-CO2, and 47% to N2O. The total direct greenhouse gas footprint of this plant was 3.9 (± 0.5) g-CO2-eq m−3. Multiplied by the average daily flow rate, this plant emits approximately 195.8 (±18.6) kg-CO2-eq d−1 from its secondary treatment system. This is much lower than the total potential downstream GHG emissions if the wastewater is not treated (totaling 0.99 kg-CO2-eq kg-COD− 1; Gori et al., 2011), which for this plant can be estimated at 3.0 × 104 kg-CO2-eq d− 1 for no treatment and 1.7 × 104 kg-CO2-eq d− 1 for primary treatment only. A previous study modeled the indirect greenhouse gas emissions for water supply pathways in Southern California (Stokes and Horvath, 2009). The results of this study (Fig. 3) included CO2 emissions from construction, distribution, power production, and operation. Their data indicate that, employing the current electricity-production methods in California, water importation and wastewater treatment have similar carbon footprints, while seawater desalination has a much greater impact. The carbon footprint for water reclamation amounted to 1023 gCO2-eq m−3 of treated water. Our measurements indicate that operational emissions add an additional 3.9 g-CO2-eq m−3 of treated water (increasing the carbon footprint by 0.38%). This suggests that the direct emissions of GHGs do not significantly increase the carbon footprint of wastewater treatment as a water production method. The flux measurement reported here may be missing a significant portion of N2O emissions, though, as our previous work involving dissolved GHGs indicated that N2O emissions may be up to 141 (±63) g-CO2-eq m− 3 (Townsend-Small et al., 2011). If this were the case, these N2O emissions alone could bring the carbon footprint of recycled water to approximately 1164 g-CO2-eq m−3, a 14% increase above the value calculated. Effluent GHG emissions associated with water recycling may be important considerations and need to be investigated further in order to understand the entire impact of the process. Furthermore, the difference between the measured direct GHG emissions from the plant and the modeled indirect emissions was so great that, even if the emissions had been an order of magnitude greater, they would still not change our assessment significantly. That is, if we were to use some
1172
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173
2500
Measured Operational Emissions
1030
LCA Carbon Footprint 1025
g-CO2,eq m-3
g-CO2,eq m-3
2000
1500
1020 1015 1010
1000
1005
500 1000 Recycled Water
0 A
B
C
D
E
Fig. 3. The net carbon footprint of the water recycling process including non-electrical GHG emissions (our data, light gray) and comparison to other water sources (Stokes and Horvath, 2009). Indirect emissions are still the dominant source of GHGs associated with water reclamation, with electricity being the largest single contributor.
of the data collected by others as indicative of this plant, our conclusions would not change. The data suggest that direct greenhouse gas emissions from water reclamation are much smaller than indirect GHG emissions. 4. Conclusions The combined GWP of the direct GHG emissions from this plant, including fossil-CO2, CH4, and N2O amounted to 3.9 g-CO2-eq m−3 of treated water. This was several orders of magnitude less than the calculated indirect GHG emissions from other forms of water production in Southern California. Furthermore, our observations indicate that power generation remains the overwhelming source of fossil GHG emission for all forms of water production. Efforts to control GHG emissions from power sources will reduce the environmental costs of human water use, and should remain a priority. Yet the direct emissions from wastewater treatment plants, albeit small, are not negligible and need to be included in future assessments. In addition, emissions of CO2 derived from fossil sources in wastewater are important to understanding the nature of anthropogenic GHGs. Global population growth, coupled with our expanding knowledge of these direct emissions, means that wastewater treatment will become an ever more critical component of our GHG budget in the future. As wastewater treatment plants are seen increasingly more as resource recovery plants, carbon sequestration will add to their environmental benefits. Acknowledgments Our thanks go out to Mr. Matt Jeung and Dr. Xiaomei Xu of the University of California, Irvine, and Mr. Dave Hayden of the Irvine Ranch Water District. Appendix A. Supplemental data Supplementary data to this article can be found online at http://dx. doi.org/10.1016/j.scitotenv.2014.10.060. References Ahn JH, Kim S, Park H, Rahm B, Pagilla K, Chandran K. N2O emissions from activated sludge processes, 2008–2009: results of a national monitoring survey in the United States. Environ Sci Technol 2010;44:4505–11.
Beaulieu JJ, Shuster WD, Rebholz JA. Controls on gas transfer velocities in a large river. J Geophys Res Biogeosci 2012;117. Cakir FY, Stenstrom MK. Greenhouse gas production: a comparison between aerobic and anaerobic wastewater treatment technology. Water Res 2005;39:4197–203. Czepiel PM, Crill PM, Harriss RC. Methane emissions from municipal waste-water treatment processes. Environ Sci Technol 1993;27:2472–7. Czepiel P, Crill P, Harriss R. Nitrous-oxide emissions from municipal waste-water treatment. Environ Sci Technol 1995;29:2352–6. Daelman MRJ, van Voorthuizen EM, van Dongen UGJM, Volcke EIP, van Loosdrecht MCM. Methane emission during municipal wastewater treatment. Water Res 2012;46: 3657–70. Djuricin S, Pataki DE, Xu X. A comparison of tracer methods for quantifying CO2 sources in an urban region. J Geophys Res - Atmos 2010;115. Doorn MRJ, Towprayoon S, Vieira SMM, Irving W, Palmer C, Pipatti R, et al. Wastewater treatment and discharge. In: Eggleston S, Buendia L, Miwa K, Ngara T, Tanabe K, editors. 2006 IPCC guidelines for national greenhouse gas inventories. Hayama, Japan: Institute for Global Environmental Studies; 2006. p. 61. El-Fadel M, Massoud M. Methane emissions from wastewater management. Environ Pollut 2001;114:177–85. EPA. Inventory of U.S. greenhouse gas emissions and sinks: 1990–2010. Washington, D.C.: U.S. Environmental Protection Agency; 2012 Foley J, de Haas D, Hartley K, Lant P. Comprehensive life cycle inventories of alternative wastewater treatment systems. Water Res 2010;44:1654–66. Forster P, Ramaswamy V, Artaxo P, Berntsen T, Betts R, Fahey DW, et al. Changes in atmospheric constituents and in radiative forcing. In: Solomon S, Qin D, Manning M, Chen Z, Marquis M, Averyt KB, et al, editors. Climate change 2007: the physical science basis. Contribution of working group I to the fourth assessment report of the intergovernmental panel on climate change. New York: Cambridge University Press; 2007. Gori R, Jiang L-M, Sobhani R, Rosso D. Effects of soluble and particulate substrate on the carbon and energy footprint of wastewater treatment processes. Water Res 2011; 45:5858–72. Griffith DR, Barnes RT, Raymond PA. Inputs of fossil carbon from wastewater treatment plants to US rivers and oceans. Environ Sci Technol 2009;43:5647–51. Kampschreur MJ, van der Star WRL, Wielders HA, Mulder JW, Jetten MSM, van Loosdrecht MCM. Dynamics of nitric oxide and nitrous oxide emission during full-scale reject water treatment. Water Res 2008;42:812–26. Law Y, Jacobsen GE, Smith AM, Yuan Z, Lant P. Fossil organic carbon in wastewater and its fate in treatment plants. Water Res 2013;47:5270–81. Nydal R, Lövseth K. Tracing bomb 14C in the atmosphere 1962–1980. J Geophys Res 1983; 88. Rosso D, Stenstrom MK. The carbon-sequestration potential of municipal wastewater treatment. Chemosphere 2008;70:1468–75. Stokes JR, Horvath A. Energy and air emission effects of water supply. Environ Sci Technol 2009;43:2680–7. Stuiver M, Polach HA. Reporting of C-14 data — discussion. Radiocarbon 1977;19:355–63. Teiter S, Mander Ü. Emission of N2O, N2, CH4, and CO2 from constructed wetlands for wastewater treatment and from riparian buffer zones. Ecol Eng 2005;25:528–41. Townsend-Small A, Pataki DE, Tseng LY, Tsai C-Y, Rosso D. Nitrous oxide emissions from wastewater treatment and water reclamation plants in Southern California. J Environ Qual 2011;40:1542–50. Townsend-Small A, Tyler SC, Pataki DE, Xu X, Christensen LE. Isotopic measurements of atmospheric methane in Los Angeles, California, USA: influence of “fugitive” fossil fuel emissions. J Geophys Res - Atmos 2012;117.
A.G. Schneider et al. / Science of the Total Environment 505 (2015) 1166–1173 USDA. Chamber-based trace gas flux measurement; USDA-ARS GRACEnet protocols; 2010 [Washington, D.C.]. Wang J, Zhang J, Xie H, Qi P, Ren Y, Hu Z. Methane emissions from a full-scale A/A/O wastewater treatment plant. Bioresour Technol 2011;102:5479–85. Wunch D, Wennberg PO, Toon GC, Keppel-Aleks G, Yavin YG. Emissions of greenhouse gases from a North American megacity. Geophys Res Lett 2009;36.
1173
Xu X, Trumbore SE, Zheng S, Southon JR, McDuffee KE, Luttgen M, et al. Modifying a sealed tube zinc reduction method for preparation of AMS graphite targets: reducing background and attaining high precision. Nucl Inst Methods Phys Res B-Beam Interact Mater Atoms 2007;259:320–9.
