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Life cycle assessment and costing of wastewater treatment systems coupled to constructed wetlands
Resources, Conservation & Recycling 148 (2019) 170–177
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Life cycle assessment and costing of wastewater treatment systems coupled to constructed wetlands Juliana Dalia Resende, Marcelo Antunes Nolasco, Sérgio Almeida Pacca
T
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Sustainability Program, School of Arts, Sciences and Humanities, University of São Paulo, Rua Arlindo Bettio, 1000, Sao Paulo, 03828-000, Brazil
A R T I C LE I N FO
A B S T R A C T
Keywords: Life cycle assessment Constructed wetlands Wastewater treatment Decentralized small-scale wastewater treatment Life cycle costs
This study evaluates the environmental and economic performance (ecoefficiency) of two decentralized, smallscale, wastewater treatment systems coupled to constructed wetlands. System One comprises a vertical and a horizontal flow wetland. System Two comprises a vertical subsurface flow wetland with artificial aeration. A life cycle assessment based on data from two actual pilot structures was carried out. The functional unit was 1 m3 of treated water over a 20-year long lifetime. The systems were modeled in open LCA software, with the aid of Ecoinvent 3.3 data, and the impact assessment was based on the ReCIPe method. Results have demonstrated that foreground emissions such as direct greenhouse gas released from the septic tank and nutrients released in the effluent have driven potential impacts related to Climate Change, Photochemical Oxidants, and Freshwater Eutrophication. Artificial aeration reduces the area required for the installation of the system and electricity consumption was responsible for only 7% of total Climate Change related potential impact. Technologies that reduce direct (foreground) liquid and airborne emissions will improve the environmental performance of the systems. The analysis of the aerated wetland has shown that the operation stage had the greatest environmental impact potential for all analyzed impact categories, with results varying between 64% for Human Toxicity and 100% for Freshwater Eutrophication. The life cycle cost per cubic meter of treated sewage for the aerated system was 1.8 times smaller than that of the system without aeration. Thus, aeration is cost-effective for small-scale wastewater treatment systems coupled to constructed wetlands.
1. Introduction Constructed wetlands (CW) and stabilization ponds are considered nature-based technologies for wastewater treatment because they are based on, processes that are observed in the natural environment (Garfí et al., 2017). CW has been recognized as a promising technology for residential wastewater treatment due to its easiness of management and maintenance (WU et al., 2015). Such systems are designed to mimic natural wetlands and contain vegetation typical from moist areas, combined with substrate and microorganisms that together enable nutrient and toxic substance removal from residual water (Vymazal, 2011; Saeed and Sun, 2012). Hybrid CW systems combine different technologies along the treated flow. One example is the use of a horizontal flow CW (HFCW) after a vertical flow CW (VFCW) or vice-versa (Vymazal, 2013). This hybrid system combines the nitrification ability of the VFCW to the denitrification ability of the HFCW, with low oxygen concentration in its substrate (Vymazal, 2009). VFCW are more efficient than HFCW in the conversion of ammonia into nitrate (Fuchs, 2009). Due to greater oxygen transport efficiency, VFCW demand a
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smaller installation area than HFCW to attain the same outflow water quality (Brix and Arias, 2005). Artificial aeration has been coupled to CW to increase system’s nutrient removal efficiency and improve its residual water treatment performance (Wu et al., 2014; Maltais-Landry et al., 2009). However, tradeoffs between intensifying water treatment efficiency and increasing operation and maintenance costs due to electricity consumption and maintenance to operate the pumps must be considered (Fan et al., 2013; Chang et al., 2014). Local climate affects CW performance, especially related to its biological oxygen demand (BOD) and nitrogen removal (Cerezo et al., 2001 e Taylor et al., 2011). In addition, qualitative and quantitative properties of the residual water intake, such as flow, dissolved solids, and pollution content also affect the performance and the lifetime of CW (Garfí et al., 2012). Different CW system setups have been evaluated, including simple systems and improved systems such as artificially aerated systems (Wu et al., 2014). Environmental, technical, and economic aspects must be considered for the selection of the most suitable wastewater treatment
Corresponding author. E-mail addresses: [email protected] (J.D. Resende), [email protected] (M.A. Nolasco), [email protected] (S.A. Pacca).
https://doi.org/10.1016/j.resconrec.2019.04.034 Received 8 August 2018; Received in revised form 29 April 2019; Accepted 30 April 2019 Available online 16 May 2019 0921-3449/ © 2019 Elsevier B.V. All rights reserved.
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2017; McNamara, 2018), however these do not usually involve the analysis of improved systems, such as, aerated CW. Although CW with forced aeration potentially intensify the removal of pollutants and reduce the area required for installation, they introduce new potential impacts due to the consumption of electricity for aeration. Thus, understanding how these choices affect the environmental performance of these systems and comparing them with other systems allows a better understanding of the tradeoffs. The goal of the present study was to assess the potential environmental impacts and the costs of two small scale WWTS coupled with CW. Results provide a benchmark and facilitate evaluating opportunities for improvements over the lifecycle of the systems.
technology (Molinos-Senante et al., 2014; Zeng et al., 2017). The analysis of potential environmental impacts related to wastewater treatment systems (WWTS) can help decision-makers in choosing the best treatment alternative or in making changes to existing systems that reduce the potential environmental impacts of their activities (Gallego et al., 2008). Singh and Kansal (2018) calculate total energy and GHG footprints of wastewater infrastructure, including energy consumption and GHG emissions from transport and treatment of wastewater, taking Delhi (India) as a case study. Zeng et al. (2017), based on the distance function approach, has assessed the efficiency of China’s urban WWTPs by considering the reduction of pollution load, as well as its related costs, energy consumption and GHG emissions. Both studies argue that the performance of WWTS is subject to increasing returns to scale. However, such studies have considered a limited scope regarding the selected category of impacts, and have not considered CW as part of the systems. Life cycle assessment (LCA) is effective in the quantitative evaluation of the overall potential environmental impacts from WWTS (Corominas et al., 2013; Hu et al., 2019). According to Roeleveld et al. (1997), the first WWTS LCA was published in 1997. After that, LCA has been applied to several WWTS, including centralized and decentralized options, pilot and actual systems, and considering construction, operation, and decommissioning stages or only part of the system. However, the evaluation is not straightforward due to intrinsic spatial and temporal variability of the WWTS (Niero et al., 2014). Both technological aspects and management options affect the environmental performance of WWTS (Machado et al., 2006). According to Hospido et al. (2012) and Foley et al. (2010), the environmental performance of a wastewater treatment plant is driven by the final disposition of the outflows and the sludge, and yet, the performance of the plant may be affected by the inflow composition, station size and local climate (Lorenzo-Toja et al., 2015). The goals for WWTS need to go beyond the protection of surface water and human health, including minimizing the loss of resources, reducing waste generation and enabling nutrients recycling (Bodík and Kubaská, 2013; Masi et al., 2018; Sena and Hicks, 2018). Diverse system boundaries are considered in wastewater treatment systems LCA. Some studies are limited to the operation of the wastewater treatment equipment whereas other studies enclose the full wastewater treatment system and pipelines for transport, including feeders, collectors and structures for water diversion from the watershed (Zang et al., 2015). Most studies consider only the operation of the WWTS ignoring construction and decommissioning phases. However, studies including the construction phase have concluded that its impact was sizeable when compared to the operation phase (Corominas et al., 2013). In the case of CW systems, due to its low energy demand and the amount of construction materials applied, depending on the impact category, the construction phase might be responsible for 80% of the potential environmental impacts (Dixon; Simon; Burkitt, 2003; Machado et al., 2006; Lopsik, 2013; Lutterbeck et al., 2017). Such results are affected by the bill of materials and the lifetime considered for the infrastructure and no generalization is possible. The lifetime of the infrastructure drives the lifecycle energy due to materials use and correspondingly the final energy footprint. (Singh; Kansal, 2018). In comparison to traditional WWTS, larger areas are required for the installation of CW (Corbella et al., 2017). Evaluating the environmental performance of WWTS is not enough, the analysis of economic aspects must be considered as well because high costs can make a project unfeasible. A method that can be used for such an analysis is Life Cycle Costing (LCC). LCC allows for determining the total cost of a project over its entire life cycle. It considers the costs incurred throughout the life cycle of the system, including costs related to electricity consumption, chemicals, maintenance, repairs, equipment replacement, and waste disposal (Rawal and Duggal, 2016). Several LCA and LCC have been applied to CW (Dixon et al., 2003; Machado et al., 2006; Lopsik, 2013; Brown, 2016; Lutterbeck et al.,
2. Material and methods In this session, we introduce the analyzed systems and the underlying information for the LCA and LCC. Detailed information is provided in the supplementary materials file. 2.1. Brief description of the assessed systems Two pilot WWTS, installed at the University of São Paulo, in São Paulo, Brazil, were assessed. The systems receive wastewater from the student housing building and the cafeteria. The intake capacities of System One (S1) and System Two (S2) were respectively 640 and 1500 liters of wastewater per day. Both systems comprise a septic tank (ST) that receives the wastewater inflow (Fig. 1). The discharge of the ST in S1 goes to a hybrid CW. A hybrid CW is a combination of a vertical and a horizontal flow CW and is an appropriate solution for wastewater treatment and reuse in small communities (Garfí et al., 2017). In S2, the discharge of the ST goes to a CW with artificial aeration followed by a secondary settler and a vertical subsurface flow (VSSF) CW. Because the two systems were constructed with different materials, a sensitivity analysis substituting fiberglass for masonry in S1 was carried out. System S1A is like S1 but was constructed with fiberglass instead of masonry. Environmental regulations in Brazil do not set specific thresholds for nitrite, nitrate, and chemical oxygen demand (COD) released from WWTS. Regarding the biological oxygen demand (BOD) the regulations set a minimum removal efficiency of 60% and for N-NH4 the maximum release concentration is 20 mg/liter. However, the environmental authority may disregard this obligation based on local conditions. Phosphorous concentration thresholds may be set by the environmental authority as well (CONAMA 2005 e 2011). In China, the regulation of pollution load reduction for WWTS is site specific as well (Zeng et al., 2017). Inlet and outlet flows were analyzed to determine chemical oxygen demand (COD), total nitrogen (Kjeldahl), ammoniacal nitrogen, nitrate, nitrite, total phosphorous, and phosphate. Table 1 presents inflow and outflow environmental parameters for S1 and S2. The performance of S2 is superior to the performance of S1 regarding the removal efficiency for most of the parameters. 2.2. Wastewater treatment system’s LCA The attributional LCA was based on ISO 14,040 and ISO 14,044 (ISO, 2006a, b). The systems were modeled in OpenLCA software (version 1.6.3). Among the impact categories considered were Terrestrial Acidification, Climate Change, Freshwater Eutrophication, Formation of Photochemical Oxidants, Human Toxicity, and Freshwater Ecotoxicity, which were evaluated based on the ReCiPe method. The system boundaries comprise unit processes related to construction, operation, and decommissioning. Material and energy input flows, emissions, and discharge output flows were assessed (Fig. 2). The functional unit was one cubic meter of treated wastewater and the lifetime of the facilities was 20 years, based on the revised literature 171
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Fig. 1. Schematic representation of S1 and S2.
reports (IPCC, 2006, 2013), which have been applied in previous wastewater treatment LCAs such as: Chen et al. (2011); Pan et al. (2011); Yoshida et al. (2014); Slagstad and Brattebø (2014); Faria et al. (2015); Guo et al. (2016). Biogenic CO2 direct emissions from the systems were not considered. The decommissioning of the systems was based on the mass of transported materials to the landfill and a 50 km long trip was considered. The LCC has included property costs, materials acquisition costs, construction costs, operation and maintenance costs, labor costs, and decommissioning costs. All costs are 2015 values and were obtained from spreadsheets provided by the professionals who build and assembled the facilities, equipment suppliers, and service companies. The installation area of S1 and S2 is respectively 7.2 m2 and 6.8 m2. The present value of the systems’ costs was determined based on a 9.6% discount rate, which was the standard for this type of assessment in Brazil.
Table 1 WWTS environmental performance. Parameters (mg/l)
inflow
outflow (S1)
outflow (S2)
S1 efficiency
S2 efficiency
COD Total N N-NH4 N-NO3 N-NO2 P-total P-PO4
967.0 80.9 52.3 1.2 1.5 10.1 7.3
129.6 74.6 45.9 0.5 1.0 7.2 5.4
76.0 38.5 30.2 1.0 1.4 5.4 4.8
87% 8% 12% 58% 33% 28% 26%
92% 52% 42% 17% 7% 47% 34%
(Shutes, 2001; Machado et al., 2006; Liu et al., 2012; Mburu et al., 2013). 2.3. Life cycle inventory data
3. Results and discussion Data was collected during the installation and the operation of the two pilot systems. Electricity consumption by the pump in S2 was determined based on the power of the equipment and its operation period. Direct atmospheric CH4 and N2O emissions were calculated based on algorithms from the International Panel on Climate Change (IPCC)
Table 2 summarizes the inventory data for the construction and the operation of the systems. In addition, transportation energy was evaluated based on the location of material supply stores, warehouses, and the installation site (S1: 0.188 t.km by truck plus 0.049 t.km by train;
Fig. 2. System boundary for the LCA of the systems. 172
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Table 2 Summary of inventory for S1, S2, and alternative construction materials for S1A based on 1 m3 of treated wastewater.
inputs Fiberglass – ST Portland cement Hydrated lime Sand Water Clay bricks Hollow bricks (21 holes) Steel rebars Fiberglass secondary treatment tanks Small gravel Large gravel PVC pipes Silicone hose PEAD hose Rubber hose Electricity (for aeration only) outputs COD Phosphorus, total Phosphate Nitrate Nitrite Ammonium Total N Methane (direct emissions) Nitrous oxide (direct emissions)
unit
S1
S2
S1A
Ecoinvent Process
kg kg kg kg kg kg kg kg kg kg kg kg kg kg kg kWh
0.00679 0.019 0.054 0.191 0.063 0.462 0.038 0.0118 – 2.63 0.454 0.00296 0.000334 – –
0.00441 – – 0.329 – – –
0.00679 – – – – – –
0.0133 – 0.746 0.000809 0.0000481 0.00054 0.0000365 1.6
0.0184 2.63 0.454 0.00296 0.000334 – –
kg kg kg kg kg kg kg kg.yr−1 kg.yr−1
0.12959 0.00723 0.0054 0.00,048 0.00095 0.04585 0.07461 0.193 0.00114
0.076 0.0054 0.0048 0.001 0.0015 0.0302 0.0385 0.179 0.0000486
0.12959 0.00723 0.0054 0.00,048 0.00095 0.04585 0.07461 0.193 0.00114
glass fibre reinforced plastic production, polyester resin, hand lay-up – GLO cement production, blast furnace slag 5-25%-RoW lime production, hydrated, packed-RoW Brazilian inventory* tap water production, conventional treatment-RoW clay brick production-GLO clay brick production-GLO reinforcing steel production glass fibre reinforced plastic production, polyester resin, hand lay-up–GLO gravel production, crushed-RoW gravel production, crushed-RoW polyvinylchloride production, suspension –RoW*polymerization** silicone product production-RoW** polyethylene production, high density, granulate –RoW** synthetic rubber production –RoW** Determined based on the 2015 Brazilian energy mix Ecoinvent Flow Chemical Oxygen Demand Phosphorus, total Phosphate Nitrate Nitrite Ammonia Nitrogen, total Methane Nitrous oxide
* Sources: Souza (2012); Castro et al. (2015) e Moraga (2017). ** Manufacturing process inventory data was not considered.
(Corbella et al., 2017). Fig. 4 shows the relative potential life cycle impact of the materials used in the construction of S1 and S2 For S1, the material that presented the greatest potential environmental impact in all categories except Freshwater Eutrophication was bricks, which presented percentages between 32% (for the Human Toxicity category) and 46% (for the category of Climate Change). The second most impacting material in almost all impact categories was steel bars (except for the categories of Climate Change, for which lime was the second most important, and Freshwater Eutrophication for which the steel bars were the material with the greatest impact potential). Although steel was used in much smaller quantities than bricks (2335.5 kg of bricks versus 55.17 kg of steel) its impact is sizeable. For S2, the tanks that are made of fiberglass are responsible for the greatest share of potential environmental impacts in all categories. Summing up the impact of the septic tank, the VSSF CW, and the settler it would reach 69% for the category of Climate Change, 87% for the Freshwater Eutrophication category, 81% for the Freshwater Ecotoxicity category, 90% for the category of Human Toxicity, 70% for the category of Terrestrial acidification, and 52% for the category of Formation of Photochemical Oxidants. Comparable results were found by Lutterbeck et al. (2017), who reported that the construction of the system represented 36% of the potential total impacts, mainly due to the use of fiberglass tanks. However, according to Lutterbeck et al. (2017), the use of fiberglass also has some advantages. Especially considering that, they are lightweight and compact materials, easy to handle and have higher chemical resistance than concrete. To compare the effect of materials choice for the installation of CW, an alternative scenario for S1 was conceived. In S1A fiberglass substitutes for masonry (cement, lime, sand, gravel, bricks, and steel rebars). Material requirements for S1 and S1A are presented in Table 1. Most potential environmental impacts associated to S1A were lower than the ones related to S1 for all analyzed impact categories. For the Climate Change category, S1 LCIA results were 4.7 greater than S1A results. For Freshwater Eutrophication, Freshwater Ecotoxicity, Human Toxicity, Terrestrial Acidification, Photochemical Oxidants the differences were 1.92, 3.04, 1.02, 2.05, and 3.22, respectively. Thus, the only
S2: 0.055 t.km by truck plus 0.084 t.km by train; S1A: 0.097 t.km by truck). The decommissioning was based on hauling the mass of demolished materials to the landfill. Decommissioning for S1, S2, and S1A was evaluated at 0.116 t.km, 0.036 t.km, and 0.078 t.km, respectively. Background data were retrieved from Ecoinvent 3.3 database (Weidema et al., 2013; Moreno Ruiz et al., 2016) and the electricity generation unit process was replaced by the Brazilian mix, which comprises 64% hydro, 12.9% natural gas, 8% biomass, 4.8% petroleum products, 4.5% coal, 3.5% wind, 2.4% nuclear, 0.01% solar photovoltaic (Empresa de Pesquisa Energética, 2016). Results from the lifecycle impact assessment (LCIA), using characterization factors from ReCiPe, show that the construction phase for S1 is responsible for more than 80% of the total impact when Freshwater Ecotoxicity, Human Toxicity and Acidification impact categories are at stake. However, when Climate Change and Eutrophication impact categories are at stake, the operation phase is responsible for more than 90% of the total impact. Such result is affected by emissions of N2O and CH4 from ST and CW. Finally, when the Photochemical Oxidant Formation impact category is considered, 66% of the impacts are due to the operation phase and 30% due to the construction phase. All S2 impact categories are dominated by the effect of the operation phase. In contrast, decommissioning LCIA are below 6% of the total for all impact categories (Fig. 3). More detailed information, including the absolute LCIA values, is available in the supplementary materials. The end-of-life stage does not significantly affect the overall systems’ impact (Corbella et al., 2017; Dixon et al., 2003; Machado et al., 2006; Foley et al., 2007; Lopsik, 2013; De Feo and Ferrara, 2017). Therefore, details of this phase are ignored in the assessment because our results agree with the literature. 3.1. Construction phase results Usually, the construction phase dominates the LCIA of CW, which are considered extensive low-tech and low energy technologies 173
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Fig. 3. Relative contribution of construction, operation, and decommissioning for S1 and S2 based on the ReCiPe impact assessment method.
Fig. 4. Relative potential lifecycle impact of construction materials for S1 and S2 based on the ReCiPe impact assessment method.
impacts for Climate Change and photochemical oxidants formation categories. Similar results were found by Fuchs et al. (2011). In their opinion greenhouse gas (GHG) emissions (CH4 and N2O) from the vertical and horizontal CWs have caused the greatest impact for the Climate Change category. The system’s discharge was responsible for most of the impact in the Freshwater Eutrophication category (99.95%), and a similar outcome was observed in S2. Roux et al. (2010) attribute the occurrence of the potential impacts related to the category of eutrophication, mainly to the incomplete removal of nitrogen and phosphorus in these systems. The replacement of gravel was responsible for 90% of the potential impacts related to ecotoxicity of fresh water, Human Toxicity and Terrestrial acidification during the operation phase of S1. Replacing gravel by an alternative material could be an
category with similar results was Human Toxicity. It should be mentioned that the amount of fiberglass used in the WC tanks of S1A (145.52 kg) was much lower than the amount of building materials used in S1 (89.76 kg Portland cement, 254.32 kg of hydrated lime, 892.80 kg of sand, 295.10 kg of water, 2160.00 kg of clay solid bricks, 175.50 kg of hollow clay bricks (21 holes), and 55.17 kg of steel rebars).
3.2. Operation phase results Fig. 5 presents the relative life cycle impacts caused by different activities during the operation of S1 and S2. For S1, methane (CH4) and nitrous oxide (N2O) directly emitted due to biogenic activity in the CW was responsible for most of the life cycle 174
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Fig. 5. Relative potential lifecycle impact of operation for S1 and S2 based on the ReCiPe impact assessment method.
3.3. Life cycle costing results
alternative to minimize impacts; however, Lopsik (2013) has found that the substitution of expanded clay for gravel has caused 10–42% more impacts. For S2, the GHG directly emitted from the system was responsible for most of the potential impacts related to the categories of Climate Change (92%) and Photochemical oxidant (62%). In contrast, electricity accounted for 31% of potential environmental impacts related to Photochemical Oxidant and its effect on the Climate Change category was below 8%. However, electricity was responsible for most of the potential impacts related to Freshwater Ecotoxicity (95%), Human Toxicity (95%), and Terrestrial acidification (91%). Considering the categories Freshwater Eutrophication, Climate Change, and the Photochemical Oxidant, the operation stage was responsible for more than 70% of the total impact for S1 and S2. Potential Freshwater Eutrophication impacts are related to Phosphorous and Nitrogen removal efficiency from the system discharge. Climate Change and the Formation of Photochemical Oxidants were driven by direct GHG emissions from the septic tank. An assessment of WWTS carried out in Delhi, which also included the transportation infrastructure, concluded that 53% of the GHG emissions were due to direct emissions (Singh and Kansal, 2018). Therefore, technologies that reduce direct (foreground) liquid and airborne emissions will improve the environmental performance of the systems. Considering the categories of human and Freshwater Ecotoxicity, and Terrestrial acidification, in S1 the impacts were driven by gravel replacement. In this case alternative filtering materials might be considered; however, the substitution of expanded clay for gravel is not recommended, as was highlighted by Lopsik (2013). For S2, electricity consumption during the operation phase has driven potential impacts related to Human and Freshwater Ecotoxicity, and Terrestrial acidification. In this case, investing in energy efficiency and powering the system with renewable energy will improve the environmental performance of S2. However, in terms of Climate Change, the reduction of biogenic direct emissions from the tanks is key to minimize potential impacts. Results from the scenario S1A demonstrate that the potential to cause environmental impacts of fiberglass was lower than the potential to cause environmental impacts of masonry throughout all analyzed impact categories. Thus, the use of fiberglass instead of masonry with bricks improves the environmental performance of the system. Masonry performance was driven by background emissions from bricks and steel rebars manufacturing. Therefore, the substitution of low impact blocks for bricks may reduce potential impacts.
Although the environmental performance of the fiberglass design was superior to the environmental performance of the masonry structure, it was more expensive. The economic assessment of the systems is presented on Table 3. More detailed information is available in the supplementary material. The total LCC of S1 was the lowest one. However, the LCC costs normalized by the capacity of the facility demonstrate that S2 is the lowest cost alternative. The design capacity of S2 was 10,950 m3 of wastewater treated over 20 years whereas the design capacity of S1 and S1A was 4672 m3 over 20 years. Thus, the flow normalized cost of S1 and S2 was respectively US$1.55/m3 and US$0.84/m3. The use of the aerated system in S2 has not compromised its cost competitiveness. Electricity expenses are responsible for only 14% of the O&M costs of the system or 4.6% of the total LCC. The use of fiberglass in the construction of S1A instead of masonry has increased the LCC of S1A by 12%, and the materials acquisition costs by 27%. This difference is because the cost of fiberglass tanks is higher than the cost for masonry construction. If the property cost is ignored the LCC per m3 of wastewater treated decreases by 48% for S1, by 36% for S2, and by 43% for S1A. We have carried out a sensitivity analysis using discount rates of 5% and 15%. The 5% discount rate implied in 2015 values 15% higher than the baseline whereas the 10% discount rate implied in 2015 values 10% lower than baseline values for both systems. Such results highlight the relevance of O&M, replacement, and decommissioning costs with respect to total present values, which is around to one third for both systems. Thus, the sensitivity analysis with respect to the discount rate did not affect the relative comparison of the systems. Garfí et al. (2017) have assessed the costs of a CW with a treatment capacity of 292.5 m3 per day and a 20-year lifetime in Spain, including a septic tank, two vertical flow CWs operating alternately, and one Table 3 Systems’ upfront cost.
175
Cost Type
S1 (US$)
S2 (US$)
Acquisition Construction and installation Operation and maintenance (present value) Replacement (present value) Decommissioning (present value) LCC in 2015 values
4,426.25 396.55 2365.69 35.97 22.10 7,595.28
6,079.86 31.43 2,745.75 323.46 20.32 9,605.56
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horizontal subsurface flow planted with Phragmites australis. The acquisition cost of the system (excluding the cost of the land) was U $2.55/m3 of treated wastewater and the operation and maintenance costs was U$6.66/m3. In the present study, acquisition (excluding cost of land), operation, and maintenance costs of S1 were, respectively, US $ 0.85/m3 and US$ 0.51/m3. For S2, these costs were, respectively, US$ 0.57/m3 and US$ 0.25/m3, being below the values found by these authors.
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ISO, International, Organization for Standardization, 2006a. ISO 14040.ENvironmental Management - Life Cycle Assessment - Principles and Framework. ISO, International, Organization for Standardization, 2006b. ISO 14044. Environmental Management - Life Cycle Assessment - Requirements and Guidelines. Liu, D., Wu, X., Chang, J., Gu, B., Min, Y., Ge, Y., Shi, Y., Xue, H., Peng, C., Wu, J., 2012. Constructed wetlands as biofuel production systems. Nat. Clim. Chang. 3, 190–194. Lopsik, K., 2013. Life cycle assessment of small-scale constructed wetland and extended aeration activated sludge wastewater treatment system. Int. J. Environ. Sci. Technol. (Tehran) 10 (6), 1295–1308. https://doi.org/10.1007/s13762-012-0159-y. Lorenzo-Toja, Y., Amores, M.J., Alfonsín, C., Aldea, X., Marínnavarro, D., Moreira, M.T., Feijoo, G., 2015. Beyond the Conventional Life Cycle Inventory in Wastewater Treatment Plants, vol.1 EcoSTP, Verona. 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4. Conclusion The present study has identified which life cycle stages (construction, operation or end-of-life) and inflows or outflows are the most significant potential impact triggers. Most potential impacts are related to the operation phase of the systems. The mitigation of foreground flows such as GHG direct emissions (CH4 and N2O) and nutrients released in the discharge of the systems are recommended strategies to improve the environmental performance of the system because they are the major drivers for Climate Change, Freshwater Eutrophication, and Photochemical Oxidant impact categories. For S2, which consumes electricity, investing in energy efficiency and decarbonization of the electricity slightly improves the life cycle performance of the system. Potential impacts of the construction phase of S1 might be reduced if fiberglass substitutes for masonry. However, the fiberglass design is more expensive than masonry. LCC results show that S2 is the least expensive alternative even if property costs and electricity expenses are considered. Finally, nutrient removal efficiency of S2 (aerated system) is superior to the nutrient removal efficiency of S1. Acknowledgements Juliana Dalia Resende is grateful to Coordenação de Aperfeiçoamento de Pessoal de Nível Superior - Brasil (CAPES) for finance in part this study - Finance Code 001. Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.resconrec.2019.04. 034. References Bodík, I., Kubaská, M., 2013. Energy and sustainability of operation of a wastewater treatment plant. Process Saf. Environ. Prot. 39 (2), 15–24. https://doi.org/10.5277/ EPE130202. Brix, H., Arias, C.A., 2005. The use of vertical flow constructed wetlands for on-site treatment of domestic wastewater: new Danish guidelines. Ecol. Eng. 25 (5), 491–500. Brown, C., 2016. Life-cycle Cost Analysis of Nutrient Reduction Technologies Employed in Municipal Wastewater Treatment. Ph.D. Dissertation. Environmental Studies Department, Oberlin College, Oberlin, pp. 59 pp.. Castro, A.L., Silva, F.B., Arduin, R.H., Oliveira, L.A., Becere, O.H., 2015. Análise Da Viabilidade Técnica Da Adaptação De Dados Internacionais De Inventário De Ciclo De Vida Para O Contexto Brasileiro: Um Estudo De Caso Do Concreto Para Paredes Moldadas No Local. Congresso Brasileiro do Concreto, Bonito, Brazil. Cerezo, R.G., Suárez, M.L., Vidal-Abarca, M.R., 2001. The performance of a multi-stage system of constructed wetlands for urban wastewater treatment in a semiarid region of SE Spain. Ecol. Eng. 16 (4), 501–517. Chang, Y., Wu, S., Zhang, T., Mazur, R., Pang, C., Dong, R., 2014. Dynamics of nitrogen transformation depending on different operational strategies in laboratory-scale tidal flow constructed wetlands. Sci. Total Environ. 487, 49–56. Chen, G.Q., Shao, L., Chen, Z.M., Li, Z., Zhang, B., Chen, H., Wu, Z., 2011. Low-carbon assessment for ecological wastewater treatment by a constructed wetland in Beijing. J. Ecol. Eng. 37 (4), 622–628. https://doi.org/10.1016/j.ecoleng.2010.12.027. Corbella, C., Puigagut, J., Garfí, M., 2017. Life cycle assessment of constructed wetland systems for wastewater treatment coupled with microbial fuel cells. Sci. Total Environ. 584–585, 355–362. https://doi.org/10.1016/j.scitotenv.2016.12.186. Corominas, L., Foley, J., Guest, J.S., Hospido, A., Larsen, H.F., Morera, S., Shaw, A., 2013. Life cycle assessment applied to wastewater treatment: state of the art. Water Res. n. 47 (15), 5480–5492. https://doi.org/10.1016/j.watres.2013.06.049. De Feo, G., Ferrara, C., 2017. A procedure for evaluating the most environmentally sound
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Full length article
Life cycle assessment and costing of wastewater treatment systems coupled to constructed wetlands Juliana Dalia Resende, Marcelo Antunes Nolasco, Sérgio Almeida Pacca
T
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Sustainability Program, School of Arts, Sciences and Humanities, University of São Paulo, Rua Arlindo Bettio, 1000, Sao Paulo, 03828-000, Brazil
A R T I C LE I N FO
A B S T R A C T
Keywords: Life cycle assessment Constructed wetlands Wastewater treatment Decentralized small-scale wastewater treatment Life cycle costs
This study evaluates the environmental and economic performance (ecoefficiency) of two decentralized, smallscale, wastewater treatment systems coupled to constructed wetlands. System One comprises a vertical and a horizontal flow wetland. System Two comprises a vertical subsurface flow wetland with artificial aeration. A life cycle assessment based on data from two actual pilot structures was carried out. The functional unit was 1 m3 of treated water over a 20-year long lifetime. The systems were modeled in open LCA software, with the aid of Ecoinvent 3.3 data, and the impact assessment was based on the ReCIPe method. Results have demonstrated that foreground emissions such as direct greenhouse gas released from the septic tank and nutrients released in the effluent have driven potential impacts related to Climate Change, Photochemical Oxidants, and Freshwater Eutrophication. Artificial aeration reduces the area required for the installation of the system and electricity consumption was responsible for only 7% of total Climate Change related potential impact. Technologies that reduce direct (foreground) liquid and airborne emissions will improve the environmental performance of the systems. The analysis of the aerated wetland has shown that the operation stage had the greatest environmental impact potential for all analyzed impact categories, with results varying between 64% for Human Toxicity and 100% for Freshwater Eutrophication. The life cycle cost per cubic meter of treated sewage for the aerated system was 1.8 times smaller than that of the system without aeration. Thus, aeration is cost-effective for small-scale wastewater treatment systems coupled to constructed wetlands.
1. Introduction Constructed wetlands (CW) and stabilization ponds are considered nature-based technologies for wastewater treatment because they are based on, processes that are observed in the natural environment (Garfí et al., 2017). CW has been recognized as a promising technology for residential wastewater treatment due to its easiness of management and maintenance (WU et al., 2015). Such systems are designed to mimic natural wetlands and contain vegetation typical from moist areas, combined with substrate and microorganisms that together enable nutrient and toxic substance removal from residual water (Vymazal, 2011; Saeed and Sun, 2012). Hybrid CW systems combine different technologies along the treated flow. One example is the use of a horizontal flow CW (HFCW) after a vertical flow CW (VFCW) or vice-versa (Vymazal, 2013). This hybrid system combines the nitrification ability of the VFCW to the denitrification ability of the HFCW, with low oxygen concentration in its substrate (Vymazal, 2009). VFCW are more efficient than HFCW in the conversion of ammonia into nitrate (Fuchs, 2009). Due to greater oxygen transport efficiency, VFCW demand a
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smaller installation area than HFCW to attain the same outflow water quality (Brix and Arias, 2005). Artificial aeration has been coupled to CW to increase system’s nutrient removal efficiency and improve its residual water treatment performance (Wu et al., 2014; Maltais-Landry et al., 2009). However, tradeoffs between intensifying water treatment efficiency and increasing operation and maintenance costs due to electricity consumption and maintenance to operate the pumps must be considered (Fan et al., 2013; Chang et al., 2014). Local climate affects CW performance, especially related to its biological oxygen demand (BOD) and nitrogen removal (Cerezo et al., 2001 e Taylor et al., 2011). In addition, qualitative and quantitative properties of the residual water intake, such as flow, dissolved solids, and pollution content also affect the performance and the lifetime of CW (Garfí et al., 2012). Different CW system setups have been evaluated, including simple systems and improved systems such as artificially aerated systems (Wu et al., 2014). Environmental, technical, and economic aspects must be considered for the selection of the most suitable wastewater treatment
Corresponding author. E-mail addresses: [email protected] (J.D. Resende), [email protected] (M.A. Nolasco), [email protected] (S.A. Pacca).
https://doi.org/10.1016/j.resconrec.2019.04.034 Received 8 August 2018; Received in revised form 29 April 2019; Accepted 30 April 2019 Available online 16 May 2019 0921-3449/ © 2019 Elsevier B.V. All rights reserved.
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2017; McNamara, 2018), however these do not usually involve the analysis of improved systems, such as, aerated CW. Although CW with forced aeration potentially intensify the removal of pollutants and reduce the area required for installation, they introduce new potential impacts due to the consumption of electricity for aeration. Thus, understanding how these choices affect the environmental performance of these systems and comparing them with other systems allows a better understanding of the tradeoffs. The goal of the present study was to assess the potential environmental impacts and the costs of two small scale WWTS coupled with CW. Results provide a benchmark and facilitate evaluating opportunities for improvements over the lifecycle of the systems.
technology (Molinos-Senante et al., 2014; Zeng et al., 2017). The analysis of potential environmental impacts related to wastewater treatment systems (WWTS) can help decision-makers in choosing the best treatment alternative or in making changes to existing systems that reduce the potential environmental impacts of their activities (Gallego et al., 2008). Singh and Kansal (2018) calculate total energy and GHG footprints of wastewater infrastructure, including energy consumption and GHG emissions from transport and treatment of wastewater, taking Delhi (India) as a case study. Zeng et al. (2017), based on the distance function approach, has assessed the efficiency of China’s urban WWTPs by considering the reduction of pollution load, as well as its related costs, energy consumption and GHG emissions. Both studies argue that the performance of WWTS is subject to increasing returns to scale. However, such studies have considered a limited scope regarding the selected category of impacts, and have not considered CW as part of the systems. Life cycle assessment (LCA) is effective in the quantitative evaluation of the overall potential environmental impacts from WWTS (Corominas et al., 2013; Hu et al., 2019). According to Roeleveld et al. (1997), the first WWTS LCA was published in 1997. After that, LCA has been applied to several WWTS, including centralized and decentralized options, pilot and actual systems, and considering construction, operation, and decommissioning stages or only part of the system. However, the evaluation is not straightforward due to intrinsic spatial and temporal variability of the WWTS (Niero et al., 2014). Both technological aspects and management options affect the environmental performance of WWTS (Machado et al., 2006). According to Hospido et al. (2012) and Foley et al. (2010), the environmental performance of a wastewater treatment plant is driven by the final disposition of the outflows and the sludge, and yet, the performance of the plant may be affected by the inflow composition, station size and local climate (Lorenzo-Toja et al., 2015). The goals for WWTS need to go beyond the protection of surface water and human health, including minimizing the loss of resources, reducing waste generation and enabling nutrients recycling (Bodík and Kubaská, 2013; Masi et al., 2018; Sena and Hicks, 2018). Diverse system boundaries are considered in wastewater treatment systems LCA. Some studies are limited to the operation of the wastewater treatment equipment whereas other studies enclose the full wastewater treatment system and pipelines for transport, including feeders, collectors and structures for water diversion from the watershed (Zang et al., 2015). Most studies consider only the operation of the WWTS ignoring construction and decommissioning phases. However, studies including the construction phase have concluded that its impact was sizeable when compared to the operation phase (Corominas et al., 2013). In the case of CW systems, due to its low energy demand and the amount of construction materials applied, depending on the impact category, the construction phase might be responsible for 80% of the potential environmental impacts (Dixon; Simon; Burkitt, 2003; Machado et al., 2006; Lopsik, 2013; Lutterbeck et al., 2017). Such results are affected by the bill of materials and the lifetime considered for the infrastructure and no generalization is possible. The lifetime of the infrastructure drives the lifecycle energy due to materials use and correspondingly the final energy footprint. (Singh; Kansal, 2018). In comparison to traditional WWTS, larger areas are required for the installation of CW (Corbella et al., 2017). Evaluating the environmental performance of WWTS is not enough, the analysis of economic aspects must be considered as well because high costs can make a project unfeasible. A method that can be used for such an analysis is Life Cycle Costing (LCC). LCC allows for determining the total cost of a project over its entire life cycle. It considers the costs incurred throughout the life cycle of the system, including costs related to electricity consumption, chemicals, maintenance, repairs, equipment replacement, and waste disposal (Rawal and Duggal, 2016). Several LCA and LCC have been applied to CW (Dixon et al., 2003; Machado et al., 2006; Lopsik, 2013; Brown, 2016; Lutterbeck et al.,
2. Material and methods In this session, we introduce the analyzed systems and the underlying information for the LCA and LCC. Detailed information is provided in the supplementary materials file. 2.1. Brief description of the assessed systems Two pilot WWTS, installed at the University of São Paulo, in São Paulo, Brazil, were assessed. The systems receive wastewater from the student housing building and the cafeteria. The intake capacities of System One (S1) and System Two (S2) were respectively 640 and 1500 liters of wastewater per day. Both systems comprise a septic tank (ST) that receives the wastewater inflow (Fig. 1). The discharge of the ST in S1 goes to a hybrid CW. A hybrid CW is a combination of a vertical and a horizontal flow CW and is an appropriate solution for wastewater treatment and reuse in small communities (Garfí et al., 2017). In S2, the discharge of the ST goes to a CW with artificial aeration followed by a secondary settler and a vertical subsurface flow (VSSF) CW. Because the two systems were constructed with different materials, a sensitivity analysis substituting fiberglass for masonry in S1 was carried out. System S1A is like S1 but was constructed with fiberglass instead of masonry. Environmental regulations in Brazil do not set specific thresholds for nitrite, nitrate, and chemical oxygen demand (COD) released from WWTS. Regarding the biological oxygen demand (BOD) the regulations set a minimum removal efficiency of 60% and for N-NH4 the maximum release concentration is 20 mg/liter. However, the environmental authority may disregard this obligation based on local conditions. Phosphorous concentration thresholds may be set by the environmental authority as well (CONAMA 2005 e 2011). In China, the regulation of pollution load reduction for WWTS is site specific as well (Zeng et al., 2017). Inlet and outlet flows were analyzed to determine chemical oxygen demand (COD), total nitrogen (Kjeldahl), ammoniacal nitrogen, nitrate, nitrite, total phosphorous, and phosphate. Table 1 presents inflow and outflow environmental parameters for S1 and S2. The performance of S2 is superior to the performance of S1 regarding the removal efficiency for most of the parameters. 2.2. Wastewater treatment system’s LCA The attributional LCA was based on ISO 14,040 and ISO 14,044 (ISO, 2006a, b). The systems were modeled in OpenLCA software (version 1.6.3). Among the impact categories considered were Terrestrial Acidification, Climate Change, Freshwater Eutrophication, Formation of Photochemical Oxidants, Human Toxicity, and Freshwater Ecotoxicity, which were evaluated based on the ReCiPe method. The system boundaries comprise unit processes related to construction, operation, and decommissioning. Material and energy input flows, emissions, and discharge output flows were assessed (Fig. 2). The functional unit was one cubic meter of treated wastewater and the lifetime of the facilities was 20 years, based on the revised literature 171
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Fig. 1. Schematic representation of S1 and S2.
reports (IPCC, 2006, 2013), which have been applied in previous wastewater treatment LCAs such as: Chen et al. (2011); Pan et al. (2011); Yoshida et al. (2014); Slagstad and Brattebø (2014); Faria et al. (2015); Guo et al. (2016). Biogenic CO2 direct emissions from the systems were not considered. The decommissioning of the systems was based on the mass of transported materials to the landfill and a 50 km long trip was considered. The LCC has included property costs, materials acquisition costs, construction costs, operation and maintenance costs, labor costs, and decommissioning costs. All costs are 2015 values and were obtained from spreadsheets provided by the professionals who build and assembled the facilities, equipment suppliers, and service companies. The installation area of S1 and S2 is respectively 7.2 m2 and 6.8 m2. The present value of the systems’ costs was determined based on a 9.6% discount rate, which was the standard for this type of assessment in Brazil.
Table 1 WWTS environmental performance. Parameters (mg/l)
inflow
outflow (S1)
outflow (S2)
S1 efficiency
S2 efficiency
COD Total N N-NH4 N-NO3 N-NO2 P-total P-PO4
967.0 80.9 52.3 1.2 1.5 10.1 7.3
129.6 74.6 45.9 0.5 1.0 7.2 5.4
76.0 38.5 30.2 1.0 1.4 5.4 4.8
87% 8% 12% 58% 33% 28% 26%
92% 52% 42% 17% 7% 47% 34%
(Shutes, 2001; Machado et al., 2006; Liu et al., 2012; Mburu et al., 2013). 2.3. Life cycle inventory data
3. Results and discussion Data was collected during the installation and the operation of the two pilot systems. Electricity consumption by the pump in S2 was determined based on the power of the equipment and its operation period. Direct atmospheric CH4 and N2O emissions were calculated based on algorithms from the International Panel on Climate Change (IPCC)
Table 2 summarizes the inventory data for the construction and the operation of the systems. In addition, transportation energy was evaluated based on the location of material supply stores, warehouses, and the installation site (S1: 0.188 t.km by truck plus 0.049 t.km by train;
Fig. 2. System boundary for the LCA of the systems. 172
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Table 2 Summary of inventory for S1, S2, and alternative construction materials for S1A based on 1 m3 of treated wastewater.
inputs Fiberglass – ST Portland cement Hydrated lime Sand Water Clay bricks Hollow bricks (21 holes) Steel rebars Fiberglass secondary treatment tanks Small gravel Large gravel PVC pipes Silicone hose PEAD hose Rubber hose Electricity (for aeration only) outputs COD Phosphorus, total Phosphate Nitrate Nitrite Ammonium Total N Methane (direct emissions) Nitrous oxide (direct emissions)
unit
S1
S2
S1A
Ecoinvent Process
kg kg kg kg kg kg kg kg kg kg kg kg kg kg kg kWh
0.00679 0.019 0.054 0.191 0.063 0.462 0.038 0.0118 – 2.63 0.454 0.00296 0.000334 – –
0.00441 – – 0.329 – – –
0.00679 – – – – – –
0.0133 – 0.746 0.000809 0.0000481 0.00054 0.0000365 1.6
0.0184 2.63 0.454 0.00296 0.000334 – –
kg kg kg kg kg kg kg kg.yr−1 kg.yr−1
0.12959 0.00723 0.0054 0.00,048 0.00095 0.04585 0.07461 0.193 0.00114
0.076 0.0054 0.0048 0.001 0.0015 0.0302 0.0385 0.179 0.0000486
0.12959 0.00723 0.0054 0.00,048 0.00095 0.04585 0.07461 0.193 0.00114
glass fibre reinforced plastic production, polyester resin, hand lay-up – GLO cement production, blast furnace slag 5-25%-RoW lime production, hydrated, packed-RoW Brazilian inventory* tap water production, conventional treatment-RoW clay brick production-GLO clay brick production-GLO reinforcing steel production glass fibre reinforced plastic production, polyester resin, hand lay-up–GLO gravel production, crushed-RoW gravel production, crushed-RoW polyvinylchloride production, suspension –RoW*polymerization** silicone product production-RoW** polyethylene production, high density, granulate –RoW** synthetic rubber production –RoW** Determined based on the 2015 Brazilian energy mix Ecoinvent Flow Chemical Oxygen Demand Phosphorus, total Phosphate Nitrate Nitrite Ammonia Nitrogen, total Methane Nitrous oxide
* Sources: Souza (2012); Castro et al. (2015) e Moraga (2017). ** Manufacturing process inventory data was not considered.
(Corbella et al., 2017). Fig. 4 shows the relative potential life cycle impact of the materials used in the construction of S1 and S2 For S1, the material that presented the greatest potential environmental impact in all categories except Freshwater Eutrophication was bricks, which presented percentages between 32% (for the Human Toxicity category) and 46% (for the category of Climate Change). The second most impacting material in almost all impact categories was steel bars (except for the categories of Climate Change, for which lime was the second most important, and Freshwater Eutrophication for which the steel bars were the material with the greatest impact potential). Although steel was used in much smaller quantities than bricks (2335.5 kg of bricks versus 55.17 kg of steel) its impact is sizeable. For S2, the tanks that are made of fiberglass are responsible for the greatest share of potential environmental impacts in all categories. Summing up the impact of the septic tank, the VSSF CW, and the settler it would reach 69% for the category of Climate Change, 87% for the Freshwater Eutrophication category, 81% for the Freshwater Ecotoxicity category, 90% for the category of Human Toxicity, 70% for the category of Terrestrial acidification, and 52% for the category of Formation of Photochemical Oxidants. Comparable results were found by Lutterbeck et al. (2017), who reported that the construction of the system represented 36% of the potential total impacts, mainly due to the use of fiberglass tanks. However, according to Lutterbeck et al. (2017), the use of fiberglass also has some advantages. Especially considering that, they are lightweight and compact materials, easy to handle and have higher chemical resistance than concrete. To compare the effect of materials choice for the installation of CW, an alternative scenario for S1 was conceived. In S1A fiberglass substitutes for masonry (cement, lime, sand, gravel, bricks, and steel rebars). Material requirements for S1 and S1A are presented in Table 1. Most potential environmental impacts associated to S1A were lower than the ones related to S1 for all analyzed impact categories. For the Climate Change category, S1 LCIA results were 4.7 greater than S1A results. For Freshwater Eutrophication, Freshwater Ecotoxicity, Human Toxicity, Terrestrial Acidification, Photochemical Oxidants the differences were 1.92, 3.04, 1.02, 2.05, and 3.22, respectively. Thus, the only
S2: 0.055 t.km by truck plus 0.084 t.km by train; S1A: 0.097 t.km by truck). The decommissioning was based on hauling the mass of demolished materials to the landfill. Decommissioning for S1, S2, and S1A was evaluated at 0.116 t.km, 0.036 t.km, and 0.078 t.km, respectively. Background data were retrieved from Ecoinvent 3.3 database (Weidema et al., 2013; Moreno Ruiz et al., 2016) and the electricity generation unit process was replaced by the Brazilian mix, which comprises 64% hydro, 12.9% natural gas, 8% biomass, 4.8% petroleum products, 4.5% coal, 3.5% wind, 2.4% nuclear, 0.01% solar photovoltaic (Empresa de Pesquisa Energética, 2016). Results from the lifecycle impact assessment (LCIA), using characterization factors from ReCiPe, show that the construction phase for S1 is responsible for more than 80% of the total impact when Freshwater Ecotoxicity, Human Toxicity and Acidification impact categories are at stake. However, when Climate Change and Eutrophication impact categories are at stake, the operation phase is responsible for more than 90% of the total impact. Such result is affected by emissions of N2O and CH4 from ST and CW. Finally, when the Photochemical Oxidant Formation impact category is considered, 66% of the impacts are due to the operation phase and 30% due to the construction phase. All S2 impact categories are dominated by the effect of the operation phase. In contrast, decommissioning LCIA are below 6% of the total for all impact categories (Fig. 3). More detailed information, including the absolute LCIA values, is available in the supplementary materials. The end-of-life stage does not significantly affect the overall systems’ impact (Corbella et al., 2017; Dixon et al., 2003; Machado et al., 2006; Foley et al., 2007; Lopsik, 2013; De Feo and Ferrara, 2017). Therefore, details of this phase are ignored in the assessment because our results agree with the literature. 3.1. Construction phase results Usually, the construction phase dominates the LCIA of CW, which are considered extensive low-tech and low energy technologies 173
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Fig. 3. Relative contribution of construction, operation, and decommissioning for S1 and S2 based on the ReCiPe impact assessment method.
Fig. 4. Relative potential lifecycle impact of construction materials for S1 and S2 based on the ReCiPe impact assessment method.
impacts for Climate Change and photochemical oxidants formation categories. Similar results were found by Fuchs et al. (2011). In their opinion greenhouse gas (GHG) emissions (CH4 and N2O) from the vertical and horizontal CWs have caused the greatest impact for the Climate Change category. The system’s discharge was responsible for most of the impact in the Freshwater Eutrophication category (99.95%), and a similar outcome was observed in S2. Roux et al. (2010) attribute the occurrence of the potential impacts related to the category of eutrophication, mainly to the incomplete removal of nitrogen and phosphorus in these systems. The replacement of gravel was responsible for 90% of the potential impacts related to ecotoxicity of fresh water, Human Toxicity and Terrestrial acidification during the operation phase of S1. Replacing gravel by an alternative material could be an
category with similar results was Human Toxicity. It should be mentioned that the amount of fiberglass used in the WC tanks of S1A (145.52 kg) was much lower than the amount of building materials used in S1 (89.76 kg Portland cement, 254.32 kg of hydrated lime, 892.80 kg of sand, 295.10 kg of water, 2160.00 kg of clay solid bricks, 175.50 kg of hollow clay bricks (21 holes), and 55.17 kg of steel rebars).
3.2. Operation phase results Fig. 5 presents the relative life cycle impacts caused by different activities during the operation of S1 and S2. For S1, methane (CH4) and nitrous oxide (N2O) directly emitted due to biogenic activity in the CW was responsible for most of the life cycle 174
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Fig. 5. Relative potential lifecycle impact of operation for S1 and S2 based on the ReCiPe impact assessment method.
3.3. Life cycle costing results
alternative to minimize impacts; however, Lopsik (2013) has found that the substitution of expanded clay for gravel has caused 10–42% more impacts. For S2, the GHG directly emitted from the system was responsible for most of the potential impacts related to the categories of Climate Change (92%) and Photochemical oxidant (62%). In contrast, electricity accounted for 31% of potential environmental impacts related to Photochemical Oxidant and its effect on the Climate Change category was below 8%. However, electricity was responsible for most of the potential impacts related to Freshwater Ecotoxicity (95%), Human Toxicity (95%), and Terrestrial acidification (91%). Considering the categories Freshwater Eutrophication, Climate Change, and the Photochemical Oxidant, the operation stage was responsible for more than 70% of the total impact for S1 and S2. Potential Freshwater Eutrophication impacts are related to Phosphorous and Nitrogen removal efficiency from the system discharge. Climate Change and the Formation of Photochemical Oxidants were driven by direct GHG emissions from the septic tank. An assessment of WWTS carried out in Delhi, which also included the transportation infrastructure, concluded that 53% of the GHG emissions were due to direct emissions (Singh and Kansal, 2018). Therefore, technologies that reduce direct (foreground) liquid and airborne emissions will improve the environmental performance of the systems. Considering the categories of human and Freshwater Ecotoxicity, and Terrestrial acidification, in S1 the impacts were driven by gravel replacement. In this case alternative filtering materials might be considered; however, the substitution of expanded clay for gravel is not recommended, as was highlighted by Lopsik (2013). For S2, electricity consumption during the operation phase has driven potential impacts related to Human and Freshwater Ecotoxicity, and Terrestrial acidification. In this case, investing in energy efficiency and powering the system with renewable energy will improve the environmental performance of S2. However, in terms of Climate Change, the reduction of biogenic direct emissions from the tanks is key to minimize potential impacts. Results from the scenario S1A demonstrate that the potential to cause environmental impacts of fiberglass was lower than the potential to cause environmental impacts of masonry throughout all analyzed impact categories. Thus, the use of fiberglass instead of masonry with bricks improves the environmental performance of the system. Masonry performance was driven by background emissions from bricks and steel rebars manufacturing. Therefore, the substitution of low impact blocks for bricks may reduce potential impacts.
Although the environmental performance of the fiberglass design was superior to the environmental performance of the masonry structure, it was more expensive. The economic assessment of the systems is presented on Table 3. More detailed information is available in the supplementary material. The total LCC of S1 was the lowest one. However, the LCC costs normalized by the capacity of the facility demonstrate that S2 is the lowest cost alternative. The design capacity of S2 was 10,950 m3 of wastewater treated over 20 years whereas the design capacity of S1 and S1A was 4672 m3 over 20 years. Thus, the flow normalized cost of S1 and S2 was respectively US$1.55/m3 and US$0.84/m3. The use of the aerated system in S2 has not compromised its cost competitiveness. Electricity expenses are responsible for only 14% of the O&M costs of the system or 4.6% of the total LCC. The use of fiberglass in the construction of S1A instead of masonry has increased the LCC of S1A by 12%, and the materials acquisition costs by 27%. This difference is because the cost of fiberglass tanks is higher than the cost for masonry construction. If the property cost is ignored the LCC per m3 of wastewater treated decreases by 48% for S1, by 36% for S2, and by 43% for S1A. We have carried out a sensitivity analysis using discount rates of 5% and 15%. The 5% discount rate implied in 2015 values 15% higher than the baseline whereas the 10% discount rate implied in 2015 values 10% lower than baseline values for both systems. Such results highlight the relevance of O&M, replacement, and decommissioning costs with respect to total present values, which is around to one third for both systems. Thus, the sensitivity analysis with respect to the discount rate did not affect the relative comparison of the systems. Garfí et al. (2017) have assessed the costs of a CW with a treatment capacity of 292.5 m3 per day and a 20-year lifetime in Spain, including a septic tank, two vertical flow CWs operating alternately, and one Table 3 Systems’ upfront cost.
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Cost Type
S1 (US$)
S2 (US$)
Acquisition Construction and installation Operation and maintenance (present value) Replacement (present value) Decommissioning (present value) LCC in 2015 values
4,426.25 396.55 2365.69 35.97 22.10 7,595.28
6,079.86 31.43 2,745.75 323.46 20.32 9,605.56
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horizontal subsurface flow planted with Phragmites australis. The acquisition cost of the system (excluding the cost of the land) was U $2.55/m3 of treated wastewater and the operation and maintenance costs was U$6.66/m3. In the present study, acquisition (excluding cost of land), operation, and maintenance costs of S1 were, respectively, US $ 0.85/m3 and US$ 0.51/m3. For S2, these costs were, respectively, US$ 0.57/m3 and US$ 0.25/m3, being below the values found by these authors.
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4. Conclusion The present study has identified which life cycle stages (construction, operation or end-of-life) and inflows or outflows are the most significant potential impact triggers. Most potential impacts are related to the operation phase of the systems. The mitigation of foreground flows such as GHG direct emissions (CH4 and N2O) and nutrients released in the discharge of the systems are recommended strategies to improve the environmental performance of the system because they are the major drivers for Climate Change, Freshwater Eutrophication, and Photochemical Oxidant impact categories. For S2, which consumes electricity, investing in energy efficiency and decarbonization of the electricity slightly improves the life cycle performance of the system. Potential impacts of the construction phase of S1 might be reduced if fiberglass substitutes for masonry. However, the fiberglass design is more expensive than masonry. LCC results show that S2 is the least expensive alternative even if property costs and electricity expenses are considered. Finally, nutrient removal efficiency of S2 (aerated system) is superior to the nutrient removal efficiency of S1. Acknowledgements Juliana Dalia Resende is grateful to Coordenação de Aperfeiçoamento de Pessoal de Nível Superior - Brasil (CAPES) for finance in part this study - Finance Code 001. Appendix A. Supplementary data Supplementary material related to this article can be found, in the online version, at doi:https://doi.org/10.1016/j.resconrec.2019.04. 034. References Bodík, I., Kubaská, M., 2013. Energy and sustainability of operation of a wastewater treatment plant. Process Saf. Environ. Prot. 39 (2), 15–24. https://doi.org/10.5277/ EPE130202. Brix, H., Arias, C.A., 2005. The use of vertical flow constructed wetlands for on-site treatment of domestic wastewater: new Danish guidelines. Ecol. Eng. 25 (5), 491–500. Brown, C., 2016. Life-cycle Cost Analysis of Nutrient Reduction Technologies Employed in Municipal Wastewater Treatment. Ph.D. Dissertation. Environmental Studies Department, Oberlin College, Oberlin, pp. 59 pp.. Castro, A.L., Silva, F.B., Arduin, R.H., Oliveira, L.A., Becere, O.H., 2015. Análise Da Viabilidade Técnica Da Adaptação De Dados Internacionais De Inventário De Ciclo De Vida Para O Contexto Brasileiro: Um Estudo De Caso Do Concreto Para Paredes Moldadas No Local. Congresso Brasileiro do Concreto, Bonito, Brazil. Cerezo, R.G., Suárez, M.L., Vidal-Abarca, M.R., 2001. The performance of a multi-stage system of constructed wetlands for urban wastewater treatment in a semiarid region of SE Spain. Ecol. Eng. 16 (4), 501–517. Chang, Y., Wu, S., Zhang, T., Mazur, R., Pang, C., Dong, R., 2014. Dynamics of nitrogen transformation depending on different operational strategies in laboratory-scale tidal flow constructed wetlands. Sci. Total Environ. 487, 49–56. Chen, G.Q., Shao, L., Chen, Z.M., Li, Z., Zhang, B., Chen, H., Wu, Z., 2011. Low-carbon assessment for ecological wastewater treatment by a constructed wetland in Beijing. J. Ecol. Eng. 37 (4), 622–628. https://doi.org/10.1016/j.ecoleng.2010.12.027. Corbella, C., Puigagut, J., Garfí, M., 2017. Life cycle assessment of constructed wetland systems for wastewater treatment coupled with microbial fuel cells. Sci. Total Environ. 584–585, 355–362. https://doi.org/10.1016/j.scitotenv.2016.12.186. Corominas, L., Foley, J., Guest, J.S., Hospido, A., Larsen, H.F., Morera, S., Shaw, A., 2013. Life cycle assessment applied to wastewater treatment: state of the art. Water Res. n. 47 (15), 5480–5492. https://doi.org/10.1016/j.watres.2013.06.049. De Feo, G., Ferrara, C., 2017. A procedure for evaluating the most environmentally sound
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