Comparative environmental impacts of source-separation systems for domestic wastewater management in rural China

Accepted Manuscript Comparative environmental impacts of source-separation systems for domestic wastewater management in rural China Lauho Lam, Kiyo Kurisu, Keisuke Hanaki PII:

S0959-6526(15)00501-6

DOI:

10.1016/j.jclepro.2015.04.126

Reference:

JCLP 5495

To appear in:

Journal of Cleaner Production

Received Date: 2 October 2014 Revised Date:

27 April 2015

Accepted Date: 28 April 2015

Please cite this article as: Lam L, Kurisu K, Hanaki K, Comparative environmental impacts of sourceseparation systems for domestic wastewater management in rural China, Journal of Cleaner Production (2015), doi: 10.1016/j.jclepro.2015.04.126. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.

ACCEPTED MANUSCRIPT

Comparative environmental impacts of source-separation systems for domestic wastewater management in rural China Lauho LAMa, Kiyo KURISUa,*, Keisuke HANAKIa Department of Urban Engineering, The University of Tokyo, 7-3-1, Hongo, Bunkyo, Tokyo 113-8656, Japan

*

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a

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Corresponding author. Tel.: +81-3-5841-8975, Email address: [email protected] (K. Kurisu)

ACCEPTED MANUSCRIPT Abstract To address issues of poor sanitation and water scarcity in developing countries, innovative wastewater management systems should be proposed, based on considerations of life cycle and local environmental impacts. Source-separation separates urine, feces, and gray water, enhancing the reuse of nutrients, and is considered a promising wastewater treatment method.

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This study evaluates the life cycle and local environmental impacts of source-separation systems. Each system was set to achieve the same effluent quality as the functional unit. Four scenarios were examined: offsite treatment (B1), onsite treatment (B2), source-separation (A1), and pour-flush toilet use (A2). The study area was a hypothetical village in the

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municipality of Tianjin in China, which has experienced severe water shortages and poor sanitation. Life cycle impacts such as global warming, acidification, and eutrophication were

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evaluated using a life cycle assessment (LCA) framework, and direct water use was evaluated as the local impact. The scenarios enhancing nutrient recovery (B2 and A2) showed significant benefit by avoiding mineral fertilizer use; however, the source-separation system (A1) showed the best performance in terms of life cycle environmental impacts, which supports the findings of several previous LCA studies with similar system boundaries. Within the selected framework and assumptions, the results show that the source-separation system

categories.

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had the best environmental performance in terms of direct water use and the lifecycle impact

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Highlights:

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Keywords: Source-separation, Wastewater treatment, Nutrient reuse, Life cycle assessment (LCA), Water shortage, Decentralized systems

• •

Life cycle analysis and direct water use of four wastewater treatment scenarios Offsite activated sludge, onsite johkasou, source-separation, and pour-flush toilet

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Processes were designed to obtain the same effluent quality from each scenario Source-separation performed best for global warming, acidification, eutrophication Source-separation also had lowest direct water use

ACCEPTED MANUSCRIPT Graphical abstract: Toilet

Feces

Urine

B1

Flush toilet

Conventional activated sludge

Anaerobic digestion

B2

Flush toilet

Johkasou (onsite treatment system)

Anaerobic digestion

A1

Urine separation

Dehydration and application to household garden

A2

Pour-flush toilet

Rotating biological contactor

Application to agricultural lands

100

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6 5 4 3 2

50

0 B2

A1

A2

B1 B2 A1 A2 Land application of mineral fertilizers Land application of urine Irrigation with treated effluent Land application of sludge Operation (others) Construction

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B1

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1 0

Sludge drying reed bed

25 20 15

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150

Constructed wetland

Direct freshwater use (m3/person/year)

Acidification (kgSO2eq/p/year)

200

Sludge

Sludge drying reed bed

7 System expansion Operation Construction

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Global warming (kgCO2eq/p/year)

250

Grey water

10 5 0

B1

B2

A1

A2

ACCEPTED MANUSCRIPT 1. Introduction According to the World Health Organization and the United Nations International Children’s Emergency Fund (WHO/UNICEF, 2012) 2.5 billion people still lack access to adequate sanitation. Poor sanitation causes illness and death worldwide, and establishing appropriate wastewater management systems is a vital issue, especially in developing

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countries. Moreover, the number of people living below the water stress threshold of 1700 m3/person/year is expected to reach 3 billion by 2025, due to increasing water demand in China, India, and Sub-Saharan Africa (United Nations Development Programme: UNEP, 2006). An emerging paradigm in sanitation is “ecological sanitation,” which promotes closed

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material-flow through the safe recovery of nutrients, water, and organic matter from wastewater and recycling them into the environment (Esrey et. al., 2001). Under such scenarios, wastewater is no longer a “waste,” but a “resource” (Langergraber and Muellegger,

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2005).

In this respect, one promising approach is the separation of urine, feces, and gray water (wastewater from shower, bath, and kitchen) and optimization of resource recovery via tailored technologies (Wilderer, 2001). Urine collected and aged in separate containers can be used directly as a pathogen-free fertilizer (Schönning, 2002). Separately collected feces can

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be treated by anaerobic digestion or composting and then used as a soil conditioner in agriculture. Gray water can be treated by various methods, e.g., constructed wetlands, stabilization ponds, and membranes, followed by reuse for irrigation, groundwater recharge, or direct use as flush water. Thus, there are numerous options, but few quantitative

systems.

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evaluations have been conducted of the life cycle environmental loads of source-separation

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Tillman et al. (1998) analyzed the life cycle environmental burdens resulting from changes to an existing centralized wastewater treatment plant in two different settings (suburban and rural), by comparing two alternative scenarios: local treatment of combined wastewater based on sand filter beds; and separate treatment of urine, feces, and gray water. Similarly, Lundin et al. (2000) compared traditional centralized wastewater treatment systems with source-separation systems at two different scales. Both studies used the life cycle assessment (LCA) method and accounted for the environmental benefits of reusing nutrients from wastewater rather than synthetic fertilizers in agriculture, but analyzed the environmental loads without conducting an impact assessment. Benetto et al. (2009) carried out a full LCA study of traditional centralized and source-separation systems, but only considered the operation phase. The use of a predefined model for the traditional wastewater treatment plant resulted in a lack of transparency when interpreting the results. Remy (2010) also compared

ACCEPTED MANUSCRIPT conventional and source-separation systems in an urban setting, considering both the construction and operation phases. Tidåker et al (2006) focused on sludge reuse to farming land and evaluated several impact categories such as global warming, acidification, eutrophication, and resource consumption in the context of a rural Swedish town. Thibodeau et al. (2014) compared conventional and black-water source-separation systems in the context

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of Quebec, Canada, via four safeguard categories: human health, ecosystem quality, climate change, and resource consumption. These studies reported some significantly different conclusions, depending on the study area, the choice of system boundaries, and the assumptions made. As shown above, most LCAs of the environmental loads of

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source-separation systems were carried out in a European context, and very few LCAs of wastewater management systems have been conducted in emerging or developing countries (Pan et al., 2011; Yildirim and Topkaya, 2012; Zhang et al., 2010). Therefore, evaluation of

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wastewater management in developing countries is important, especially for rural areas where sanitation conditions are usually poor. Furthermore, the systems that are compared rarely have identical functions. Indeed, to ensure reliable comparison, the investigated systems should have the same efficiency, which can be determined by identical treatment performances (Renou, 2006). Most of the reviewed LCA studies did not consider the differing treatment performances of the wastewater treatment systems while defining the functional unit; the

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amount of treated water or sludge has often been used as the functional unit (Benetto et al., 2009; Emmerson et al., 1995; Lundin et al., 2000; Remy, 2010; Tillman et al., 1998). Furthermore, the local impacts are usually not well accounted for, despite their potential

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significance. Especially in North China, where annual precipitation is low and water resources are limited, the impact of water use itself is also important for evaluating the local impacts of various wastewater management scenarios.

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Therefore, the present study evaluates the life cycle and local environmental impacts of source-separation systems and other alternatives, and compares the relative advantages of source-separation systems using the same discharged water quality for all scenarios as the functional unit and based on local data. The well-established LCA method was chosen to quantify the life cycle environmental impacts, while direct freshwater use was used as an indicator of local environmental impact. These estimations based on detailed quantitative settings of wastewater treatment processes enable reliable comparison between lifecycle and local impacts, detailed evaluation of environmental loadings from various stages, and more meaningful comparison between source-separation and other types of systems, based on realistic treatment performances and local data.

ACCEPTED MANUSCRIPT 2. Life Cycle Assessment Framework The life cycle framework, such as target area, functional unit, and system boundaries was defined with reference to a rural area of North China. 2.1. Study area The study area was defined as a hypothetical village, typical of those in rural areas of

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Tianjin, North China. Tianjin was chosen because it faces severe water shortages due to declining water resource and increasing water demand, which is a scenario common among developing countries. Rapid industrialization and urbanization in the study are associated with excessive utilization of limited water resources, resulting in reduced river flow (Bai and Imura,

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2001). Even though various countermeasures help manage the scale of abstraction, groundwater has been over-exploited to address surface water shortages, causing land

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subsidence, salt water intrusion, and soil salinization (Institute for Global Environmental Strategies, 2007). The rural population of Tianjin is spread among villages of various sizes, and was estimated at 2.6 million in 2011, representing 20% of the total population (Tianjin Municipal Bureau of Statistics: TMBS, 2012).

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The settings in the hypothetical village were based on realistic conditions (Table 1). The population was set at 2,000 inhabitants, which is a common size for villages in North China. The village was assumed to be surrounded by an agricultural area managed by the inhabitants. The houses were placed in a grid (Fig. 1), and the number of houses was estimated from the average family size of 3.27 persons in Tianjin in 2011 (TMBS, 2012). The characteristics of each component of domestic wastewater (urine, feces, and gray water) were established from an extensive literature review of rural conditions in Tianjin

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(Table 2). For urine, the excreted volume was taken from Gao et al. (2002), and the nitrogen and phosphorus contents were calculated based on Swedish data that were adapted to the Chinese nutritional diet (Jönsson et al., 2004). The values of Chemical Oxygen Demand (COD) for urine (g/person/day) have been reported in several studies: 12 (Putnam, 1971), 17 (Takahashi et al., 1989), 13 (Bennett and Linstedt, 1975), 13 (Udert et al., 2003), 15 (Udert et al., 2003), and 10 (Vinnerås et al., 2006); we used an estimate of 13 g/person/day assuming daily urine generation of 1.59 l/person/day (Gao et al., 2002). Similarly, the quantity of feces assumed in this study comes from Gao et al. (2002), and the nitrogen and phosphorus levels in feces were calculated in the same way as for urine to reflect the Chinese nutritional diet (Jönsson et al., 2004). The COD values of feces reported in the literature showed greater variability than those of urine. The fecal COD value used in this study is the average of the following values (g/person/day): 48 (Takahashi et al., 1989), 34 (Bennett and Linstedt, 1975), 68 (Nwaneri et al., 2008), and 87 (Lopez Zavala et al., 2002), assuming a dry mass of feces of 60 g/person/day (Gao et al., 2002). The daily volume of gray water is typical of rural areas in China (Haase, 2011). The COD and total nitrogen (TN) values are the median values of the

ACCEPTED MANUSCRIPT concentrations measured by Li and Zhou (2011), and total phosphorous (TP) is the maximum measured concentration from the same source. Table 1 Conditions in the study area. Value 560 12 −6 30 2000 20 625 3.2

Tap water at home

Yes

Private flush toilets

Water resources

Area of cultivated land by rural households Wheat (double rotation) Corn (double rotation) Others Water availability

194 107 107 87 113

Evapotranspiration Summer maize Winter wheat Groundwater level Estimated pump lift Distance to the closest sludge treatment plant

Source 1961–1990 average from World Meteorological Organization

Assumed (a) Assumed area of 450 m × 450 m Assumed (b) Calculated based on (a) and (b). Assumed based on National Bureau of Statistics and Department of Environmental Protection of China (2006) Assumed based on Zeng et al. (2008) Calculated based on TMBS (2012) Calculated based on sown areas in Jixian County (TMBS, 2012)

ha ha ha ha m3/person/year

2003–2011 average from National Bureau of Statistics of China

460 mm/year

Foster et al. (2003)

340 40 58 30

Foster et al. (2003) Wang et al. (2012) Wang et al. (2012) Assumed

mm/year m m km

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Wastewater

No

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Agricultural land

Unit mm/year °C °C °C inhabitants ha houses persons

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Village

Data type Annual precipitation Annual average temperature Min. temperature (winter) Max. temperature (summer) Population Area Number of houses Family size

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Category Climate

Fig. 1 A hypothetical village showing the sewer network used in some scenarios.

ACCEPTED MANUSCRIPT Table 2 Average composition of urine, feces, and gray water in the study area. Unit

Urine

Feces

Gray water

kg/person/day l/person/day

1.59

0.315 (wet) 0.315

83

g/person/day g/person/day g/person/day

13 9.6 1.1

60 1.4 0.55

12 1.2 0.12

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Flow Quantity Mass Volume Considered constituents Chemical oxygen demand (COD) Total nitrogen (TN) Total phosphorous (TP)

2.2. Definition of goal and scope

The goal of this study is to compare source-separation systems with other domestic

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wastewater management systems from a life cycle perspective in the specific context of rural Tianjin, China. As shown in section 2.1, we use local data rather than general data. Water use

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and wastewater composition were derived from Chinese data, which differ substantially from those in Western and developed countries. For example, the daily gray water generation in China selected in this study (83 l/person/day) is much lower than for Australia (113 l/person/day; Morel and Diener, 2006) and Switzerland (110 l/person/day; Morel and Diener, 2006), while drier locations such as South Africa show much lower values (20 l/person/day; Morel and Diener, 2006). In consideration of local conditions, the work is intended to identify

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the comparative advantages and disadvantages of each system and, in particular, to determine whether source-separation systems are environmentally superior to more common wastewater management systems under given assumptions, as is often advocated (Langergraber and

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Muellegger, 2005). 2.2.1. Functional unit

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The primary function of the systems investigated is to collect and dispose of domestic wastewater, thus achieving effluent water quality of 25 mg/l biological oxygen demand (BOD), which is usually considered as the effluent standard for treated sewage in several countries such as France. The functional unit was set as the wastewater (urine, feces, and gray water) discharged annually by one person. 2.2.2. Wastewater management scenarios Four scenarios were developed for reuse of nutrients from domestic wastewater, as shown in Fig. 2. (1) Baseline-1: Offsite treatment (B1) represents the traditional, centralized approach to wastewater management, which is recognized as a reasonable system for densely populated areas. Flush toilets are used to collect urine and feces. Combined domestic wastewater is

ACCEPTED MANUSCRIPT transported by gravity sewers to a nearby small-scale wastewater treatment plant for processing via complete-mix activated sludge (CMAS). The effluent is chlorinated and reused for irrigation, and excess sludge is transported by a truck to an anaerobic digestion (AD) plant, 30 km from the village. The produced biogas is combusted and converted to energy, which offsets the heating energy required for the AD tank. Finally, the liquid residue from digestion

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is utilized as a fertilizer for the agricultural land surrounding the village. (2) Baseline-2: Onsite treatment (B2) represents a typical advanced, decentralized management system. Flush toilets are used to collect urine and feces. Combined wastewater is treated onsite at the household level via “Gappei-shori johkasou,” which is a popular

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treatment system in rural areas of Japan (Funamizu, 2010; Japan Education Center of Environmental Sanitation; Yang et al., 2001). Johkasou comprises various onsite domestic wastewater treatment systems, developed in Japan since the 1960s to respond to the growing

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demand for flush toilets, mainly in rural areas inaccessible to the sewer network.

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(d) Alternative 2: Pour-flush toilet use

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Fig. 2 Wastewater management scenarios

The main advantages of johkasou systems are high treatment performance, low vulnerability to natural disasters, ease of installation, and reuse of the treated effluent and sludge (Japan Education Center of Environmental Sanitation). In Japan, more than 30 million people are served by johkasou (Yang et al., 2001). After chlorination, the treated effluent is reused as irrigation water. The sludge is regularly collected by a truck and treated as in scenario B1. (3) Alternative-1: A1 represents a wastewater management system with source-separation. Urine, feces, and gray water are separately treated. Urine-diverting dry toilets (UDDT) are used to collect urine separately from feces without flush water. The urine is stored in a tank

ACCEPTED MANUSCRIPT and reused as a liquid fertilizer, while the feces is stored in a dehydration vault and kept dry before being buried in soil. Gray water is transported via gravity sewers to a nearby small-scale treatment plant consisting of vertical-flow constructed wetlands (VFCWs) with primary treatment. The effluent is used as irrigation water, the sludge is treated in sludge-drying reed beds (SDRBs), and the dried sludge is applied to land after 10 years of

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operation, during which sanitization is completed. (4) Alternative-2: A2 represents a centralized wastewater management system that emphasizes resource saving. Pour-flush toilets, which use less water for flushing than standard toilets, are used to collect urine and feces. Combined domestic wastewater is

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transported by gravity sewers to a nearby small-scale treatment plant equipped with rotating biological contactors (RBC), which generally consume less energy than a conventional activated sludge system. The treated effluent is chlorinated and reused for irrigation. The

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excess sludge is treated in SDRBs and subsequently applied to agricultural land. 2.2.3. System boundaries

The study considered both the construction and operation phases. To ensure that all scenarios were comparable, the “system expansion” method was used to account for secondary functions by adding supplementary functions. For example, as organic fertilizers

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are produced from wastewater in each scenario, the system boundary was expanded to include the production of synthetic fertilizers that supply equivalent products, so that each scenario provides the same amount of equivalent fertilizers. Similarly, the provision of irrigation water as treated effluent in each scenario was taken into account by adding groundwater irrigation in

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the system as an equivalent process. Even though they were mathematically equivalent, an “additive” system expansion was chosen rather than a “subtractive” one, to avoid possible

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misinterpretation of the results. 2.2.4. Data sources

The data for the target systems and raw materials were obtained from various literature sources (Section 3). For example, the carbon intensity of electricity in China is higher because of coal use. To take into account these local conditions, the background processes used the Chinese Life Cycle Database (CLCD), developed by Sichuan University and IKE Environmental Technology Co. Ltd. The data are based on the Chinese market and data quality is assured (Greenhouse Gas Protocol). Data for fiber-reinforced plastic (FRP) and rubber, which were not available from the CLCD at the time of this study, were taken from the Japan Environmental Management Association for Industry (JEMAI) database and the European Life Cycle Database (ELCD), respectively.

ACCEPTED MANUSCRIPT 2.2.5. Life Cycle Impact Assessment The life cycle impact assessment (LCIA) was completed using LIME-2 (LCIA based on endpoint modeling, ver. 2) developed in Japan (Itsubo and Inaba, 2010). We only considered the midpoint impact categories, such as global warming, acidification, and eutrophication, without aggregation into damage categories. Global warming potentials corresponding to a

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time horizon of 100 years (GWP100) were used as characterization factors to measure global warming impact (Intergovernmental Panel on Climate Change: IPCC, 2007). For acidification and eutrophication, the characterization factors used were the deposition-oriented acidification potential (DAP) and the eutrophication potential with fate model (EPF),

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respectively, which reflect the Japanese environment (Itsubo and Inaba, 2010). Characterization factors specific to Chinese local conditions have not been proposed yet; therefore, we used the factors for another East Asian country, Japan, rather than factors based

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on European conditions proposed by Heijungs et al., (1992) and Huijbregts (1999). It should be noted that emissions of NH3 to air and water are differentiated for eutrophication in LIME-2.

3. Life Cycle Inventory Analysis

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To determine the inventory inputs for each scenario, some components or processes were designed, based on wastewater characteristics, so that the consumption of resources could be estimated.

3.1. Wastewater collection

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The following sections describe the data sources and design parameters used to assess the

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first stages of the wastewater collection processes, such as toilets and sewers. 3.1.1. Toilets

The squatting pan was not included in the inventory, as it was present in all scenarios. Data from Gandy et al. (2012) were used for the inventory of construction materials used in flush toilets. For the dehydration vaults in the UDDT, the amount of concrete was calculated from the dimensions and indications of an existing project in China (Kumar, 2009). The amount of water used for flushing in the operational phase was estimated as shown in Table 3.

ACCEPTED MANUSCRIPT Table 3 Assumed flush water volumes for each scenario. Scenario B1 B2 A1 A2 b

Flush Flush UDDT Pour-flush

Flush volume [l/flush] 9a 9a 0 3b

Flushing frequency [flushes/person/day] 7.75a 7.75a 0 7.75a

Flush water volume [l/person/day] 69.8 69.8 0 23.3

Gandy et al. (2012) UNEP-Division of Technology, Industry, and Economics (DTIE)International Environmental Technology Center (IETC), 2000

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a

Toilet

3.1.2. Gravity sewers

The gravity sewers were designed based on the total volume of wastewater and a

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minimum self-cleansing velocity of 0.6 m/s, taking into account a peaking factor of 5 (Alberta Environment and Sustainable Resource Development, 2013). The selected diameters of the

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main sewer pipes and tributary pipes were 250 mm and 200 mm respectively in B1, and 200 mm and 150 mm respectively in A1 and A2 (Fig. 1). Sewer pipes were assumed to be made of polyvinyl chloride (PVC), which is commonly used in small settlements. The amount of PVC was calculated using data from Harvel Plastic, Inc. (2004).

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3.2. Wastewater treatment

The settings of each wastewater treatment processes were determined from the detailed design of each process.

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3.2.1. Johkasou

The johkasou system considered in this study was a typical model for BOD removal, designed for five persons based on an anaerobic filter-contact aeration process with

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recirculation, including a final disinfection step (Kubota Co., Japan). The quantity of FRP, the main material of the body, was estimated by subtracting the blower weight (4.7 kg) from the product weight (240 kg) (Kubota Co., Japan). This estimate was consistent with the value of 73.7 kg for a separating wall made of FRP inside the johkasou used by Nishimura (2010). Electricity consumed by the air blower (35 W) and for continuous chlorine disinfection was included in the inventory of the operational phase. As data on the chlorine tablets (90% available Cl) used in johkasou were not available from accessible life cycle inventory datasets, a solution of NaClO (15% available Cl) was used as a proxy, assuming a yearly consumption of 3.5 kg of chlorine tablets (Ministry of the Environment of Japan; WEPA).

ACCEPTED MANUSCRIPT 3.2.2. Complete-mix activated sludge (CMAS) Activated sludge is the most common suspended-growth biological process for municipal wastewater treatment because of its high effluent quality, reliability, and small space requirement. After a primary treatment step to remove settleable solids, wastewater enters an aeration tank where organic matter is biodegraded under aerobic conditions by an active mass

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of microorganisms. The resulting mixed liquor then flows to a secondary clarifier to allow the microbial suspension to settle while the treated effluent flows out. The settled biomass is returned to the aeration tank and excess biomass is periodically removed (Metcalf and Eddy, 2004). A typical CMAS process for BOD removal was considered here.

Table 4 Basic settings for the target CMAS process.

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Original value

2.5c 36.7 53d

hour m3/(m2 day) m3

1.36 4.5e 68d,f

kg BOD/(m3 day) m m3

85c lb BOD/(103ft3.day)

24.4 4.9c 68d

m3/(m2 day) m m3

600e gal/(ft2.day)

2g 45d

900c gal/(ft2.day)

hour m3

Calculated from data in Tables 1–3; b Assuming a COD/BOD5 ratio of 2.0 (Mara and Horan, 2003); c Crites and Tchobanoglous (1998); d Assuming a freeboard above waterline of 0.5 m; e Metcalf and Eddy (2004); f Assuming a BOD removal rate of 35% during primary settling; g USEPA (2002).

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a

Unit m3/day kg/day

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Value 310a 85a,b 1.5c

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Parameter Average flow rate Average BOD5 load applied Peak day factor Primary sedimentation Detention time Surface overflow rate (avg. flow rate) Calculated tank volume Aeration Volumetric loading rate Side water depth Calculated tank volume Secondary sedimentation Surface overflow rate (avg. flow rate) Side water depth Calculated tank volume Chlorination Detention time Calculated contact tank volume

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The basic settings for the target CMAS process are summarized in Table 4.

The amount of reinforced concrete to construct the tanks was estimated from the design criteria shown in Table 4 (Crites and Tchobanoglous, 1998; Metcalf and Eddy, 2004; United States Environmental Protection Agency: USEPA, 2002). The consumption of electricity by the air blower and recirculation pump, and consumption of chlorine for disinfection were included in the inventory. Electricity consumption of 109 kWh/day was calculated for the air blower, using an oxygen requirement of 0.8 (kg O2)/(kg COD applied) and an actual oxygen transfer efficiency of 5% (Crites and Tchobanoglous, 1998). The electricity consumption for sludge pumping was estimated at 2.8 kWh/day, assuming a recycling ratio of 0.6 (Crites and Tchobanoglous, 1998), a sludge density of 1020 kg/m3 (Grundfos Co., Denmark), a pump

ACCEPTED MANUSCRIPT head at duty point of 4.6 m (originally shown as 15 ft), and an overall efficiency (motor + pump) of 0.85. The quantity of NaClO (15% available Cl) was estimated using a chlorine dose of 30 mg/l (USEPA, 2002).

3.2.3. Rotating biological contactors (RBC) In the RBC process, the degradation of organic matter is accomplished by a biofilm

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formed by microbes on a series of closely spaced circular plastic disks that are partially submerged in wastewater and rotated around a shaft. Aeration occurs by exposure to the atmosphere as the disks slowly rotate. RBC plants have encountered numerous problems, mostly caused by inappropriate mechanical design, but well-designed systems have operated

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successfully.

RBC units are sometimes covered to prevent algae growth, exposure of the plastic disks to

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ultraviolet radiation, and excessive heat loss (Metcalf and Eddy, 2004), but this was not considered in this study. To achieve BOD concentration of 25 mg/l in the effluent, a configuration of one stage with two treatment trains without recirculation was selected. The amount of reinforced concrete for construction of the tanks was calculated from the design criteria shown in Table 5 (Chen and Liew, 2003; Crites and Tchobanoglous, 1998; Metcalf and Eddy, 2004; USEPA, 2002; Water Environment Federation, 2008). Each treatment train

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contained a standard module of high-density polyethylene (HDPE) of diameter 3.7 m, length 7.6 m, and a media surface of area 9,300 m2 (Metcalf and Eddy, 2004). In addition to reinforced concrete, the amount of HDPE for the contactor media was calculated assuming a density of 60 kg/m3 for the media, based on typical values for the density of plastic media

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used in trickling filters (Metcalf and Eddy, 2004). The consumption of electricity for disk rotation (motor) and of chlorine for disinfection were included in the inventory. The power of

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one mechanical shaft was estimated at 1.8 kW, assuming a disk rotational speed of 1.3 rpm (Chen and Liew, 2003), based on an empirical formula from Patwardhan (2003). The quantity of NaClO (15% available Cl) was estimated using a chlorine dose of 30 mg/l (USEPA, 2002).

ACCEPTED MANUSCRIPT Table 5 Basic settings for RBC. Value 217a 85a,b 1.5c

Unit m3/day kg/day

Original value

2.5c 36.7 37d

h m3/(m2 day) m3

4.1 1.52 61d

l/ m2 m m3

0.1e gal/ft2 5f ft

24.4 3.4e 34d

m3/(m2 day) m m3

600g gal/(ft2.day)

2h 32d

hour m3

Calculated from data in Tables 1–3; b Assuming a COD/BOD5 ratio of 2.0 (Mara and Horan, 2003); c Crites and Tchobanoglous, 1998; d Assuming a freeboard above waterline of 0.5 m; e WEF, 2008; f Chen and Liew, 2003; g Metcalf and Eddy, 2004; h USEPA, 2002.

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a

900c gal/(ft2.day)

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Parameter Average flow rate Average BOD5 load applied Peak day factor Primary sedimentation Detention time Surface overflow rate (avg. flow rate) Calculated tank volume RBC Tank volume/media surface Side water depth Calculated volume for one tank Secondary sedimentation Surface overflow rate (avg. flow rate) Side water depth Calculated tank volume Chlorination Detention time Calculated contact tank volume

3.2.4. Vertical-flow constructed wetlands (VFCW)

VFCWs with a primary sedimentation step were used for gray water treatment. A VFCW

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consists of a sand/gravel filter bed that is planted with aquatic reeds that are common in rather cold climates. Wastewater is pumped intermittently to the surface and percolates vertically through the unsaturated filter substrate to a drainage system in the bottom. The intermittent loading period is followed by a long resting period. The main role of the reeds is to maintain

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good hydraulic conductivity in the bed and to improve the transfer of oxygen into the root zone. A plastic liner is placed under the bed to prevent wastewater infiltration into

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groundwater. VFCWs have been increasingly used for their high removal of organic matter and nitrification, as well as their low operational energy consumption and requirements, where enough building lot is available (Hoffmann et al., 2011). The amount of reinforced concrete for the tanks and sand/gravel substrate for the beds were estimated from the basic settings shown in Table 6. The plastic liner was also included in the inventory, assuming a commonly used 30-mil PVC liner (USEPA, 1999). The electricity consumption of the pump for intermittent gray water loading into the beds was estimated at 0.09 kWh/day.

ACCEPTED MANUSCRIPT Table 6 Basic settings for VFCW.

a

10e 2f 31d 20e 1214g 15 10 60 20

Original value

hour m3/(m2 day) m3 doses/day m m3 g COD/(m2‧day) m2 cm cm cm cm

900c gal/(ft2.day)

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2.5c 36.7 28d

Unit m3/day kg/day

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Value 166a 12a,b 1.5c

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Parameter Average flow rate Average BOD5 load applied Peak day factor Primary sedimentation Detention time Surface overflow rate (avg. flow rate) Calculated tank volume Wet well Intermittent loading frequency Side water depth at peak flow rate Calculated tank volume VFCW Limit of organic loading in cold climates Calculated total area of beds Bed compositione Freeboard for water accumulation Gravel layer Sand layer Gravel layer covering drainage pipes

Calculated from data in Tables 1–3; b Assuming a COD/BOD5 ratio of 2.0 (Mara and Horan, 2003); c Crites and Tchobanoglous (1998); d Assuming a freeboard above waterline of 0.5 m; e Hoffmann et al. (2011); f Grundfos Co.; g Assuming a COD removal rate of 35% during primary settling.

3.3. Sludge treatment

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The quantity of sludge (dry and wet weights) produced during wastewater treatment was calculated for each scenario based on typical values taken from the literature: 0.144 kg dry solids(DS)/m3 (5% total solids: TS content) for primary sludge (Crites and Tchobanoglous, 1998); 0.092 kg-DS/m3 (1% TS content) for waste-activated sludge (Crites and

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Tchobanoglous, 1998); 0.079 kg-DS /m3 for waste sludge from RBC based on a typical value for trickling filters (Crites and Tchobanoglous, 1998) with a TS content of 2.5% (Metcalf and

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Eddy, 2004); and 10.5 g-DS /person/day for johkasou (Ichinari et al., 2008) with a typical TS content of 2%. The calculated quantities of sludge are shown in Table 7.

Table 7 Sludge production. Scenario B1 B2 A1 A2 *

Sludge type

Primary + WAS* Johkasou sludge Primary sludge Primary + RBC sludge

Dry weight [kg DS*/day] 73 21 5 48

WAS: waste-activated sludge, DS: dry solids, WS: wet solids.

Wet weight [kg WS*/day] 3,753 1050 97 1,767

ACCEPTED MANUSCRIPT 3.3.1. Anaerobic digestion (AD) AD is commonly used for sludge digestion. In this process, organic matter is degraded under anaerobic conditions in a sealed digester tank and is transformed into biogas, mostly composed of CO2 and CH4. A high-rate digester operated in the mesophilic range (temperatures of 30–38°C) was assumed, in which the sludge is heated and completely mixed.

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The biogas produced was burned directly in a central heat and power (CHP) engine that produces electricity and heat simultaneously and was used as a source of energy for sludge heating and mixing. Sludge dewatering or thickening were not considered in this study.

For the inventory of construction materials, data for a large-scale AD plant designed for

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100,000 persons were used (Jungbluth et al., 2007; Ronchetti et al., 2002), the basic settings of which are shown in Table 8. Only the amounts of concrete and steel for the digester tanks and gas holder were included in the inventory. The quantities of materials were allocated in

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proportion to the quantity of sludge (as dry solids) produced in the village and treated at the large-scale AD plant. Biogas combustion was assumed to generate a modest amount of energy that was only sufficient to offset that required for the anaerobic digestion process. Table 8 Basic settings for anaerobic digestion.

467b 86b 45,000a

Unit kg DS/day m3 m3 kg

Ronchetti et al. (2002); b Calculated using data from Ronchetti et al. (2002) and Jungbluth et al. (2007).

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a

Value 8,500a

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Parameter Capacity Digester tank Concrete Gas holder Concrete Steel

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3.3.2. Sludge-drying reed beds (SDRBs) SDRBs, which are similar to VFCWs, usually consist of shallow gravel/sand beds planted with aquatic plants, commonly reeds. Sludge is spread onto the beds following a semi-continuous regime: after a feeding period of a few days, the bed rests to allow sludge dewatering, while sludge is spread on another bed, before starting a new cycle of feeding and resting. Sludge accumulates on the bed in layers. When the sludge layer reaches the maximum height, the bed stops receiving sludge and rests for a final period of several months to improve sludge dryness and mineralization, before being emptied and starting a new operating cycle. A full operating cycle usually lasts for 10 years. Plants enhance water evaporation and oxygen transfer through the filter substrate. The leachate is collected by a network of perforated pipes located in the bottom of the bed (which also enhances aeration in

ACCEPTED MANUSCRIPT the various substrate layers) and is returned to the wastewater treatment headwork. The bed is usually lined with a membrane or plastic liner (Uggetti et al., 2010). The amount of filter substrate was estimated using the design criteria and bed composition from Uggetti et al. (2010) as shown in Table 9, and the amount of plastic liner was calculated as for VFCW. Sludge is fed by gravity, thus no energy flow was included in the inventory for

Table 9 Basic settings for SDRB.

a

Uggetti et al., 2010

Unit kg·DS/(m2·year)

32 321

m2 m2

1.7 15 25 20

m cm cm cm

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Value 55a

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Parameter Sludge loading rate Calculated total area of beds Scenario A1 Scenario A2 Bed compositiona Freeboard for sludge accumulation Sand layer Gravel layer Gravel layer covering drainage pipes

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the operation phase.

Table 10 Assumptions for estimation of direct air emissions in all scenarios. Scenario

A1

A2

Abiotic CO2 [g CO2]

CH4 [mg CH4]

N2O [kg N2O-N] 0.00037a

NH3 [kg NH3-N]

NOx [mg NO2]

SOx [mg SO2]

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inf. kg TN mg 0.085b 3 Nm 1,087c 1.1*10−6 c 1,817c kg N 0.012a,b 0.26a kg N 1,570e 0.012a,b 0.15e 6f f p/year 1.106*10 0.0168 mg 0.085b Nm3 1,087c 1.1*10−6 c 1,817c kg N 0.012a,b 0.26a kg N 0.012a,b 0.26a e a,b kg N 1,570 0.012 0.15e inf. mg BOD VFCW 0.0015g 0.00043g or kg TN SDRB m2.d 1000h 1.3*10−4 h 0.20b Dried sludge* kg N 0.010a,b 0.05i a,b Irrigation kg N 0.012 0.26a a,b Urine* kg N 0.012 0.15d,e e a,b Urea* kg N 1,570 0.012 0.15e RBC inf. kg TN 0.00037 a SDRB m2.d 1000h 1.3*10−4 h 0.20b a,b Dried sludge* kg N 0.010 0.05i Urea* kg N 1,570e 0.012a,b 0.15e *: Land application, a Doka (2003); b Hobson (2000); c Nielsen and Illerup (2003); d Johansson et al. (2001); e Nemecek and Kagi (2007); f MEJ (2007); g Søvik et al. (2006); h Uggetti et al. (2012); i Assumed

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B2

CMAS AD CHP Liquid sludge* Urea* Johkasou AD CHP Liquid sludge* Irrigation Urea*

Per unit

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B1

Process

65c

65c

ACCEPTED MANUSCRIPT 3.4. Direct emissions to air during operation Direct emissions to air during the operational stage, and indirect emission of N2O from the direct emission of nitrogen species were accounted for in the inventory. The considered species were CO2 (abiotic), CH4, N2O, NH3, NOx, and SOx. All assumptions are listed in Table 10. To estimate direct emissions to air during anaerobic digestion and biogas

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combustion, biogas production rate was assumed as 0.196 m3/kg-DS (Environmental Protection Research Institute of Light Industry, 2011; Ronchetti et al., 2002). NH3-N emission ratio during agricultural irrigation was assumed as 26% of nitrogen in the effluent, which was also assumed for the anaerobically digested sludge. This assumption derives from Saez et al.

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(2012), who measured ammonia losses between 15% and 35% (average 22%) when irrigating

3.5. Background processes

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with secondary-treated effluent.

As explained in section 2.2.4, CLCD data were used for the background systems. This section discusses the background systems that required specific settings, such as

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transportation, irrigation, and fertilizer use.

3.5.1. Transportation of sludge by truck

Transportation by truck was inventoried as payload distance (t km/year). It was assumed

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that the large-scale AD plant was located 30 km from the village and that the agricultural land

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surrounding the village was located 500 m from the SDRBs.

3.5.2. Irrigation from groundwater (system expansion) In North China, groundwater is commonly used to irrigate double-cropping wheat in winter, while maize is grown without irrigation in summer. The quantity of groundwater for irrigation, required to make each scenario comparable, was calculated after estimating the energy use rate based on the estimated pump lift of 58 m in Tianjin (Table 1) and the following formula from Wang et al. (2012):     (ℎ⁄ ) =

9.81 × H × 1000 3.6 × 10 × η

where H = pump lift (m), η = (pump + motor) efficiency (0.85)

ACCEPTED MANUSCRIPT 3.5.3. Mineral fertilizer supply (system expansion) Urea and triple superphosphate are commonly used as mineral fertilizers for double-cropping wheat and maize in North China (Zhao et al., 2006). For each scenario, the amounts of plant-available nitrogen and phosphorus from organic fertilizers relative to mineral fertilizers were estimated from the literature, after taking into

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account nitrogen losses during wastewater and sludge treatments. As shown in Table 11, scenarios B1 and A2 returned more nutrients as organic fertilizer than other scenarios, resulting in lower usage of mineral fertilizer.

For nitrogen, the common indicator used is the nitrogen fertilizer replacement value

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(NFRV), which is the amount of mineral nitrogen fertilizer (kg N/ha) needed to replace 100 kg total N/ha in the residue to obtain the same response in crop yield or N off-take (Petersen, 2003). Based on reported NFRVs of digested slurry applied to land (Pedersen, 2001; Schröder

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et al., 2007), a nitrogen plant availability of 60% relative to mineral fertilizers was assumed for liquid anaerobically digested sludge. For dried sludge from SDRBs, an NFRV of 30% was used, based on the NFRV of 31% for surface-applied farm manure reported by Schröder et al. (2007). A nitrogen plant availability of 100% relative to mineral fertilizers was assumed for urine (Jönsson et al., 2004), as nitrogen is mostly in the form of urea; and for the treated effluent, as nitrogen is almost completely in the form of NH4+ or NO3−, which are readily

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available to plants. For phosphorus, plant availability of 60% was assumed relative to mineral fertilizers for liquid anaerobically digested sludge and dried sludge from SDRBs, based on the literature review undertaken by Coker and Carlton-Smith (1986). Phosphorus plant

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availability of 100% relative to mineral fertilizers was assumed for urine (Jönsson et al., 2004) and treated effluent, as phosphorus is mostly in the form of phosphate that is readily available

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to plants.

ACCEPTED MANUSCRIPT Table 11 Plant-available nutrients from organic and mineral fertilizers.

B1

Organic fertilizer

B2

A1

Phosphorous (kgP/year) 156

4,775 5,494 0 547

1,042 1,199 0 281

1,776 2,323 3,172 15 358 4,121 4,494 1,000 263

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A2

Liquid anaerobically digested sludge Treated effluent Total Mineral fertilizer Organic Liquid anaerobically digested fertilizer sludge Treated effluent Total Mineral fertilizer Organic Dried sludge from SDRB fertilizer Treated effluent Urine Total Mineral fertilizer Organic Liquid anaerobically digested fertilizer sludge Treated effluent Total Mineral fertilizer

Nitrogen (kgN/year) 719

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Fertilizer

834 1,115 83 10 37 562 610 589 156

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Scenario

4,911 5,175 320

1,042 1,199 0

Typical nitrogen and phosphorus removal rates were assumed for the wastewater treatment processes. For johkasou, removal rates of 72% and 36% were assumed for nitrogen and

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phosphorus, respectively (Yang et al., 2001). For the CMAS and RBC processes, 25% of the influent nitrogen and 20% of the influent phosphorus are transferred to sludge in the investigated scenarios, based on the assumptions made by Doka (2009). The literature shows that VFCW performance is highly dependent on local conditions (Zhang et al., 2009).

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Nitrogen and phosphorus removal rates were assumed to be 40% and 45% respectively, based on values reported for VFCW plants in China (Song et al., 2006; Zhang et al., 2009; Zhao and

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Liu, 2013).

The results of our inventory of resource consumption are summarized in Table 12.

ACCEPTED MANUSCRIPT Table 12 Inventory of resource consumption for all scenarios.

Waste -water treatment

C

O

Sludge treatment

C

O System expansion

O

Freshwater Concrete RS** Gravel Sand PVC FRP HDPE Electricity NaClO*** Concrete RS** Gravel Sand PVC Electricity PD**** Electricity Urea TSP

0.7 57,809 2,188 125 125 50,918 132 15

Scenarios B2 A1 713 0.7 38,770 2,188 125 125 46 5 747 1,220 1,241

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38,770

16,973 141 16

147,063 40,792 27 5 0.4

191,625 16 1 0.1

32

9,449 31,681 19

30 8 60

296 81 379

0 82,191 0 0 0

0 22,995 0 7 0.3

8 9,787 2 2

80 6,343 1 0

* C: Construction, O: Operation, **: Reinforcing steel, ***: 15% av. Cl, ****: Payload distance

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A2

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O

m3 t kg kg kg kg m3/y m3 t t t kg kg kg kWh/y t/y m3 t t t kg kWh/y t.km/y kWh/y t/y t/y

B1

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C

Concrete Cast iron PVC HDPE LDPE Rubber

Unit

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Waste -water collection

Material

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Phase*

ACCEPTED MANUSCRIPT 4. Results Based on the above settings, the local impacts from direct fresh water use, and the lifecycle environmental impacts, including global warming, acidification, and eutrophication, were estimated.

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4.1. Direct freshwater use Fig. 3 shows local environmental impact, as indicated by direct freshwater use. When flush toilets are used, approximately 25 m3/person/year of water is utilized, which is significant in relation to the local water availability of 113 m3/person/year in Tianjin (Table 1). Thus, the types of system and toilets used in water-scarce areas should be carefully

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considered, as greater comfort for the user can lead to larger abstraction of local water resources, resulting in increased pressure on the local environment.

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Direct freshwater use (m3/person/year)

30 25 20 15 10

0 B1

B2

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5

A1

A2

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Fig. 3 Water use for all scenarios. 4.2. Global warming impact

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The total global warming impacts and the individual contributions of each stage are shown in Fig. 4 and Fig. 5, respectively. The Baseline 2 scenario (onsite treatment) scores the highest by far, because of the high electricity consumption and direct greenhouse gas emissions of johkasou compared to other treatment systems. Alternative scenario 1 (source-separation) has the lowest impact, the most important contributors being air emissions from the land application of urine and equivalent products from system expansion. The Baseline 1 (offsite treatment) and Alternative 2 (“resource saving”) scenarios have similar scores, where electricity consumption for wastewater treatment and air emissions from irrigation have the largest effects. Economy of scale is achieved by centralized systems in the Baseline 1 and Alternative 2 scenarios compared to the Alternative 1 scenario, in terms of resource consumption for

ACCEPTED MANUSCRIPT construction, while operation (including system expansion) is the most important phase for all scenarios.

System expansion Operation Construction

200

150

100

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50

B2

A1

A2

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0 B1

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Global warming (kg CO2eq/p/year)

250

Fig. 4 Global warming impact for all scenarios 70

50 40 30 20

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10 0

Equivalent products (SE) Irrigation with treated effluent Land application of liquid sludge Chlorination Sludge transportation Anaerobic digestion + CHP Complete mix AS (direct air emissions) Complete mix AS (electricity consumption) Complete mix AS (construction) Sewers PVC 250/200mm Flush toilet

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Global warming (kg CO2eq/p/year)

Baseline 1 (B1) 60

Operation

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Construction

200

Baseline 2 (B2) Equivalent products (SE) Irrigation with treated effluent Land application of liquid sludge Chlorination Sludge transportation Anaerobic digestion + CHP Johkasou (direct air emissions) Johkasou (electricity consumption) Johkasou (construction) Flush toilet

175

Global warming (kg CO2eq/p/year)

150 125 100

75 50 25 0 Construction

Operation

ACCEPTED MANUSCRIPT 35 Alternative 1 (A1)

Mineral fertilizers supply (SE) Irrigation from groundwater (SE)

25

Land application of urine Irrigation with treated effluent

20

Land application of dried sludge Sludge transportation

15

Sludge drying reed beds Primary settler + VFCW

10

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Global warming (kg CO2eq/p/year)

30

Land application of mineral fertilizers (SE)

Sewers PVC 200/150mm

5

Urine-diverting dry toilet

0 Construction

Operation

Alternative 2 (A2)

50

Mineral fertilizers supply + land application (SE) Irrigation from groundwater (SE) Irrigation with treated effluent Land application of dried sludge Chlorination Sludge transportation Sludge drying reed beds RBC (direct air emissions) RBC (electricity consumption) RBC (construction) Sewers PVC 200/150mm Pour flush toilet

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Global warming (kg CO2eq/p/year)

60

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70

40 30 20

0 Construction

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10

Operation

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Fig. 5 Contribution of each stage to total global warming impact

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4.3. Acidification impact

The Baseline 1 and Alternative 2 scenarios achieve similar scores for acidification, followed by the Baseline 2 scenario and the Alternative 1 scenario, the latter representing about half of the highest score (Fig. 6). For all scenarios, the majority of the environmental impact is associated with the land application of fertilizers (organic and synthetic). For the Baseline 1 and Alternative 2 scenarios, irrigation with treated effluent is the most significant contribution to acidification. For the Baseline 2 scenario, irrigation with treated effluent and the land application of mineral fertilizers have almost equal effects on acidification, while in the Alternative 1 scenario, the land application of urine leads to the highest impact.

ACCEPTED MANUSCRIPT 7

5 4 3 2

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1 0 B1 B2 A1 A2 Land application of mineral fertilizers Land application of urine Irrigation with treated effluent Land application of sludge Operation (others) Construction

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Fig. 6 Acidification impact for all scenarios

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Acidification (kg SO2eq/p/year)

6

4.4. Eutrophication impact

The results for eutrophication show the same tendencies as for acidification in terms of the relative score between scenarios and the share between processes in each scenario (Fig. 7). The Alternative 2 scenario scores highest, closely followed by the Baseline 1 scenario. The Baseline 2 scenario scores lower than the Baseline 1 and Baseline 2 scenarios in terms of

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acidification impact, while the Alternative 1 scenario has the lowest impact (50% of the highest score). The land application of nitrogen fertilizers results in emission of NH3 directly to air; this contributes to both acidification and eutrophication and is the main reason for the

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similar impacts of acidification and eutrophication.

0.15

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Eutrophication (kg PO43-eq/p/year)

0.20

0.10

0.05

0.00 B1

B2

A1

A2

Land application of mineral fertilizers Land application of urine Irrigation with treated effluent Land application of sludge Operation (others) Construction

Fig. 7 Eutrophication impact for all scenarios

ACCEPTED MANUSCRIPT 5. Discussion The results indicate several environmental advantages of source-separation systems associated with natural wastewater treatment processes, compared to the traditional centralized system based on the activated sludge process. This supports the findings of several previous LCA studies with similar system boundaries (Lundin et al., 2000; Remy, 2010;

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Tillman et al., 1998). Despite lower reuse of nutrients from wastewater compared to the traditional wastewater management system, the source-separation scenario still outperforms all other scenarios for all considered impacts because the present analysis includes treated effluent as a source of organic fertilizer. Source-separation demonstrated the most efficient

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use of resources from a life cycle perspective, while achieving good treatment performance and significant reuse of nutrients from wastewater to agricultural land, thus confirming the environmental benefits of wastewater management based on source-separation. Remy (2010)

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found that source-separation had higher acidification impact than traditional systems, while Benetto et al. (2009) found that they had larger impacts on climate change and acidification. However, our study considered a different configuration of source-separation systems, given our target area, and it should be kept in mind that the results of LCA studies are very much dependent on the selected technologies, system boundaries, and assumptions made; thus,

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careful consideration is needed when comparing studies.

The environmental impacts of the operational phase were much larger than the construction phase, as reported in several previous studies (Lundin et al., 2000; Remy, 2010; Tillman et al., 1998). Additionally, the traditional centralized wastewater management

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systems (scenarios B1 and A2) achieved economies of scale during the construction phase, compared to the source-separation system (scenario A1), which is in agreement with the

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conclusions of Tillman et al. (1998) and Lundin et al. (2000). Despite the inclusion of more energy-saving and local treatment solutions, the “pour-flush toilet use” scenario (A2) did not present a significant advantage compared to the more traditional offsite treatment scenario (B1) for life cycle environmental impacts. Generally, the RBC process used in A2 scenario has been considered as one of the lower-energy systems. However, as considered in this study, in order to achieve the same wastewater treatment performance (represented by effluent BOD concentration) as other systems, RBC requires additional disks; this resulted in the higher energy consumption of RBC in the A2 scenario. As shown in Remy (2010) and Renou et al. (2008), electricity use can account for a significant contribution to the global warming impact of wastewater treatment systems even when using European and IPCC data. However, the current global warming impacts, driven

ACCEPTED MANUSCRIPT by electricity consumption based on higher carbon intensity in China, can be decreased by substituting greener fuels such as natural gas and renewable energy; this might result in competitive global warming impact between scenarios A2 and A1 scenario. Uggetti et al. (2011) showed lower environmental impact of a decentralized SDRB system. However, as seen in the comparison between scenarios B1 and A2, lower greenhouse gas

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emissions from centralized anaerobic sludge digestion with CHP can compensate the higher greenhouse gas emissions driven by sludge transportation in scenario B1. Scenario A2 produces negligible greenhouse gas emissions from sludge transportation; however, it emits more greenhouse gases during sludge treatment. Thus, the two scenarios are comparable

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overall. If energy is not recovered by CHP, B1 results in much larger global warming impacts from sludge treatment. Righi et al. (2013) compared centralized and decentralized sludge treatment systems, and showed larger contribution of transportation. However, they also

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showed that combining anaerobic digestion with CHP can compensate for transportation impacts.

For the onsite treatment scenario (B2), johkasou seems to involve significant environmental drawbacks, as it uses much more electricity in operation and emits a large amount of greenhouse gases due to the high degree of sophistication, while achieving low reuse of nutrients from wastewater in agriculture because of the loss of nitrogen to the air.

of scenarios B1 and A2.

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However, the effects on acidification and eutrophication were significantly lower than those

As reported by Tidåker et al. (2007), urine application can have significant impacts on

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acidification and eutrophication for all scenarios, because of large emissions of NH3 to the air. Additionally, the source-separation scenario also showed significant global warming impact; however, benefits were also found from avoiding mineral fertilizer use. As seen in Figures 6

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and 7, scenarios B1 and A2 can avoid the significant acidification and eutrophication impacts associated with mineral fertilizer use, as found in several other studies (Johansson et al., 2008; Lundin et al., 2000; Remy et al., 2010; Tidåker et al, 2007).

6. Conclusions Four wastewater scenarios were evaluated based on detailed wastewater system designs for achieving the same effluent quality as the functional unit. Chinese local data, such as water demand, wastewater composition, and background data, were used for local and lifecycle environmental evaluation.

ACCEPTED MANUSCRIPT Within the selected framework and assumptions, the results show that the source-separation system (scenario A1) had the best environmental performance in all impact categories assessed. Electricity consumption accounted for a significant proportion of global warming impact. This trend has been reported in other previous studies; however, higher carbon intensity in China also results in higher contribution of electricity consumption to the

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total global warming impact. Johkasou has cumulative disadvantages compared to other systems, because of its higher energy consumption, higher emissions of greenhouse gases, and lower reuse of nutrients. The “resource saving” scenario did not present any significant environmental benefits compared to

negative impacts on acidification and eutrophication.

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the more traditional offsite treatment scenario, except for water use, and even had the greatest

In all scenarios, the reuse of nutrients from wastewater also resulted in significant adverse

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environmental impacts from air emissions of greenhouse gases and air pollutants. However, avoiding the use of mineral fertilizers also provides important benefits that should be taken into account.

This study has limitations of available statistical and LCA data. Further studies are necessary to more closely analyze the functions of the assessed systems, the determination of

results.

References

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system boundaries, and the assumptions and data used, which can significantly influence the

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ACCEPTED MANUSCRIPT Doka, G., 2003. Life cycle inventories of waste treatment services. Ecoinvent report no. 13. Swiss Center for Life Cycle Inventories. Dubendorf, Switzerland. http://www.doka.ch/13_I_WasteTreatmentGeneral.pdf (accessed in July, 2014). Emmerson, R.H.C., Morse, G.K., Lester, J.N., Edge, D.R., 1995. Life-cycle analysis of small-scale sewage treatment processes. Water and Environment Journal 9(3), 317–325.

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Funamizu, N., 2010. Current status of wastewater technologies for small communities in Japan. In: Hao, X., Novotny, V., Nelson, V. (Eds) Water infrastructure for sustainable communities–China and the world. IWA Publishing. Gao, X. Z., Shen, T., Zheng, Y., Sun, X., Huang, S., Ren, Q., Zhang, X., Tian, Y., Luan, G., 2002. Practical manure handbook. (In Chinese). Chinese Agriculture Publishing House. Beijing, China. Gandy, S., Dolley, P., Parry, G., 2012. Developing an evidence base on flushing toilets and urinals. Task 4: Base case assessment. Final draft report. AEA, http://susproc.jrc.ec.europa.eu/toilets/docs/ Task4_Report_Draft_may12.pdf (accessed in July, 2014). Greenhouse Gas Protocol, “Chinese Life Cycle Database”, http://www.ghgprotocol.org/Third-Party-Databases/CLCD (accessed in February, 2015)

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Grundfos Water Utility Solutions. The sewage pumping handbook. http://net.grundfos.com/ doc/webnet/waterutility/assets/downloads/sewage-handbook.pdf (accessed in July, 2014). Haase P.H., Zhao, J., Wang, S., Godavitarne, C., 2011. Guide for wastewater management in rural villages in China. World Bank, http://water.worldbank.org/sites/water.worldbank.org/files/publication/WATERGuide-Wastewater-Management-Rural-Villages-China.pdf (accessed in July, 2014). Plastic, Inc., 2004. PVC CPVC & LXT Engineering and Design Data. http://www.plascowelding.com/pdf2007/PVC&CPVCPipeandFittings/PVC&CPVCPipe/PVC,CPVC&L XT-EngineeringandDesignData.pdf (accessed in July, 2014).

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Harvel

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Heijungs, R., Guinée, J.B., Huppes, G., Lankreijer, R.M., Udo de Haes, H.A., Wegener Sleeswijk, A., Ansems, A.M.M., Eggels, P.G., Duin, R. van, Goede, H.P. de, 1992. Environmental Life Cycle Assessment of products. Guide and Backgrounds (Part 2). Centre of Environmental Sciense (CML), Leiden University, Leiden. https://openaccess.leidenuniv.nl/handle/1887/8062 (accessed in July, 2014). Hobson, J., 2000. CH4 and N2O Emissions from Waste Water Handling. IPCC Good Practice Guidance and Uncertainty Management in National Greenhouse Gas Inventories, 441-454. http://www.ipcc-nggip. iges.or.jp/public/gp/bgp/5_2_CH4_N2O_Waste_Water.pdf (accessed in July, 2014). Hoffmann, H., Platzer, C., Winker, M., Muench, E.v., 2011. Technology Review of Constructed Wetlands. Subsurface Flow Constructed Wetlands for Greywater and Domestic Wastewater Treatment. Deutsche Gesellschaft für Technische Zusammenarbeit GmbH (GTZ) Sustainable sanitation-Ecosan program, http://www.susana.org/lang-en/library?view=ccbktypeitem&type=2&id=930, (accessed in July, 2014). Huijbregts, M., 1999. Life cycle impact assessment of acidifying and eutrophying air pollutants. Calculation of equivalency factors with RAINS-LCA. Interfaculty Department of Environmental Science, Faculty of Environmental Science, University of Amsterdam, The Netherlands. http://media.leidenuniv.nl/legacy/ Life-cycle%20impact%20assessment.pdf (accessed in July, 2014).

ACCEPTED MANUSCRIPT Ichinari,T., Ohtsubo, A., Ozawa, T., Hasegawa, K., Teduka, K., Oguchi, T., Kiso, Y., 2008. Wastewater treatment performance and sludge reduction properties of a household wastewater treatment system combined with an aerobic sludge digestion unit. Process Biochemistry 43, 722–728. Institute for Global Environmental Strategies, 2007. Sustainable Groundwater Management in Asian Cities: A Final Report of Research on Sustainable Water Management Policy. IGES Freshwater Resources Management Project, Institute for Global Environmental Strategies, Hayama, Japan, p. 97. http://pub.iges.or.jp/modules/envirolib/upload/981/attach/00_complete_report.pdf (accessed in July, 2014).

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Itsubo, N., Inaba, A., 2010. LIME 2-Life-cycle Impact Assessment Method based on Endpoint Modeling, Japan Environmental Management Association for Industry, Tokyo. http://lca-forum.org/english/pdf/ No13_C0_Introduction.pdf (accessed in July, 2014).

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Japan Education Center of Environmental Sanitation website. Wastewater Treatment with Johkasou Systems, https://www.jeces.or.jp/en/technology/index.html (accessed in July, 2014). Johansson, K., Perzon, M., Fröling, M., Mossakowska, A., Svanström, M., 2008. Sewage sludge handling with phosphorus utilization-life cycle assessment of four alternatives. Journal of Cleaner Production 16, 135– 151. Johansson, M., Jönsson, H., Höglund, C., Richert Stintzing, A., Rodhe, L., 2001. Urine separation – closing the nutrient cycle. Stockholm Water Company. Stockholm, Sweden. http://www.swedenviro.se/ gemensamma_se/documents/Urinsep_eng.pdf (accessed in July, 2014).

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Jönsson, H., Stintzing, A.R., Vinnerås, B., Salmon, E., 2004, Guidelines on the use of urine and faeces in crop production, Stockholm Environment Institute, EcoSan Res Programme, Stockholm, Sweden. http://www.ecosanres.org/pdf_files/ESR_Publications_2004/ESR2web.pdf (accessed in February, 2015).

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Jungbluth, N., Chudacoff, M., Dauriat, A., Dinkel, F., Doka, G., Faist Emmenegger, M., Gnansounou, E., Kljun, N., Schleiss, K., Spielmann, M., Stettler, C., Sutter, J., 2007. Life cycle Inventories of Bioenergy. Ecoinvent report No 17, Swiss Centre for Life Cycle Inventories, Dübendorf, Switzerland. http://www.bfe.admin.ch/php/modules/publikationen/stream.php?extlang=en&name=en_917750084.pdf (accessed in July, 2014).

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ACCEPTED MANUSCRIPT Metcalf and Eddy, 2004. Wastewater Engineering: Treatment, Disposal, Reuse. Fourth edition. McGraw-Hill, New York. Ministry of the Environment of Japan, 2007. National Greenhouse Gas Inventory Report of Japan 2007. Greenhouse Gas Inventory Office of Japan (GIO), Center for Global Environmental Research (CGER), National Institute for Environmental Studies (NIES). http://www-gio.nies.go.jp/aboutghg/ nir/2012/NIR-JPN-2012-v3.0E.pdf (accessed in July, 2014).

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Morel, A., Diener, S., 2006. Greywater Management in Low and Middle-Income Countries, Review of different treatment systems for households or neighbourhoods. Swiss Federal Institute of Aquatic Science and Technology (Eawag). Dubendorf, Switzerland. http://www.eawag.ch/forschung/sandec/publikationen/ewm/dl/GW_management.pdf (accessed in April, 2015) National Bureau of Statistics of China. China Statistical Yearbook. China Statistics Press. http://www.stats.gov.cn/english/statisticaldata/AnnualData/ (accessed in July, 2014).

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National Bureau of Statistics and Department of Environmental Protection of China (2006). China Environmental Statistical Yearbook 2006. China Statistics Press.

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Nemecek, T., Kagi, T., 2007. Life cycle inventories of agricultural production systems. Ecoinvent report no. 15 Swiss Center for Life Cycle Inventories, Dübendorf, Switzerland. http://www.upe.poli.br/~cardim/PEC/ Ecoinvent%20LCA/ecoinventReports/15_Agriculture.pdf (accessed in July, 2014). Nielsen, M., Illerup, J.B., 2003. Emission factors and inventory of emissions for decentralized CHP (In Danish, with an English summary). Eltra PSO projekt 3141. Kortlægning af emissioner fra decentrale kraftvarmeværker. Delrapport 6. Danmarks Miljøundersøgelser. 116 s. –Faglig rapport fra DMU nr. 442. Nishimura, T., 2010. Life Cycle Assessment of Gappei-shori Johkasou system. Journal of Life and Environmental Sciences 2, 31-40 (in Japanese).

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Nwaneri, C.F., Foxon, K.M., Bakare, B.F., Buckley, C.A., 2008. Biological degradation processes within a pit latrine. WISA 2008 Conference. http://www2.gtz.de/Dokumente/oe44/ecosan/en-biological-degradation -processes-2008.pdf (accessed in July, 2014). Pan, T., Zhu, X.D., and Ye, Y.P., 2011. Estimate of life-cycle greenhouse gas emissions from a vertical subsurface flow constructed wetland and conventional wastewater treatment plants: A case study in China. Ecological Engineering 37, 248–254.

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Patwardhan, A.W., 2003. Rotating biological contactors: A review. Industrial & Engineering Chemistry Research 42(10), 2035-2051.

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Pedersen, C.Å., 2001. Oversigt over Landsforsøgene (Report of field trials) (In Danish) National Crop Committee, Aarhus, Denmark. Petersen, J., 2003. Nitrogen fertilizer replacement value of sewage sludge, composted household waste and farmyard manure. Journal of Agricultural Science 140, 169–182. Putnam, D.F., 1971. Composition and concentrative properties of human urine. NASA, Washington, USA. http://ntrs.nasa.gov/archive/nasa/casi.ntrs.nasa.gov/19710023044_1971023044.pdf (accessed in July, 2014). Remy, C., 2010. Life Cycle Assessment of conventional and source-separation systems for urban wastewater management. PhD thesis, Technical University of Berlin, Institute of Environmental Technology, Berlin, Germany. http://opus4.kobv.de/opus4-tuberlin/frontdoor/index/index/docId/2429 (accessed in July, 2014). Renou, S., 2006. Life cycle assessment applied to wastewater treatment systems, PhD thesis, Institut National Polytechnique de Lorraine (INPL), Nancy, France (in French). Renou, S., Thomas, J.S., Aoustin, E., Pons, M.N., 2008. Influence of impact assessment methods in wastewater treatment LCA. Journal of Cleaner Production 16, 1098-1105.

ACCEPTED MANUSCRIPT Righi, S., Oliviero, L., Pedrini, M., Buscaroli, A., Casa, C.D., 2013. Life Cycle Assessment of management systems for sewage sludge and food waste: centralized and decentralized approaches. Journal of Cleaner Production 44, 8-17. Ronchetti, C., Bienz, P., Pridal, R., 2002. Ökobilanz Klärgasverstromung. (Life cycle assessment of sewage gas electrification) (in German), Swiss Federal Office of Energy, Bern, Switzerland. Saez, J.A., Harmon, T.C., Doshi, S., Guerrero, F., 2012. Seasonal ammonia losses from spray-irrigation with secondary-treated recycled water. Water & Science Technology 65(4), 676–682.

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Schönning, C., 2002. Evaluation of microbial health risks associated with the reuse of source-separated human urine. EcoSanRes, Sweden. http://www.diva-portal.org/smash/get/diva2:8844/FULLTEXT01.pdf (accessed in July, 2014). Schröder, J.J., Uenk, D., Hilhorst, G.J., 2007. Long-term nitrogen fertilizer replacement value of cattle manures applied to cut grassland. Plant and Soil 299, 83–99.

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Song, Z., Zheng, Z., Li, J., Sun, X., Han, X., Wang, W., Xu, M., 2006. Seasonal and annual performance of a full-scale constructed wetland system for sewage treatment in China. Ecological Engineering 26, 272– 282.

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Søvik, A.K., Augustin, J., Heikkinen, K., Huttunen, J.T., Necki, J.M., Karjalainen, S.M., Kløve, B., Liikanen, A., Mander, U., Puustinen, M., Teiter, S., Wachniew, P., 2006. Emission of the Greenhouse Gases Nitrous Oxide and Methane from Constructed Wetlands in Europe. Journal of Environmental Quality 35, 2360– 2373. Takahashi, Y., Wydeven, T. and Koo, C., 1989. Subcritical and supercritical water oxidation of celss model wastes. Advances in Space Research 9(8), 99–110. Tianjin Municipal Bureau of Statistics, 2012. Tianjin Statistical Yearbook 2012. China Statistics Press.

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Tidåker, P., Mattsson, B., Jönsson, H., 2007. Environmental impact of wheat production using human urine and mineral fertilisers - a scenario study. Journal of Cleaner Production 15, 52-62. Tillman, A.M., Svingby, M., Lundström, H., 1998. Life cycle assessment of municipal wastewater systems. International Journal of LCA 3, 145–157. Thibodeau, C., Monette, F., Bulle, C., Glaus, M., 2014, Comparison of black water source-separation and conventional sanitation systems using life cycle assessment, Journal of Cleaner Production, 67, 45-57.

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Tidåker, P., Kärrman, E., Baky, A., Jönsson, H., 2006, Wastewater management integrated with farming- an environmental systems analysis of a Swedish country town, Resources, Conservation and Recycling, 47, 295-315.

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Udert, K.M., Larsen, T.A., Biebow, M., Gujer, W., 2003. Urea hydrolysis and precipitation dynamics in a urine-collecting system. Water Research 37, 2571–2582. Uggetti, E., Ferrer, I., Llorens, E., Garcia, J., 2010. Sludge treatment wetlands: A review on the state of the art. Bioresource Technology 101, 2905–2912. Uggetti, E., Ferrer, I., Molist, J., and Garcia, J., 2011. Technical, economic and environmental assessment of sludge treatment wetlands. Water research 45, 573–582. Uggetti, E., Ferrer, I., Arias, C., Brix, H., Garcia, J., 2012. Carbon footprint of sludge treatment reed beds. Ecological Engineering 44, 298–302. UNEP, 2006. Human Development Report 2006, Beyond Scarcity: Power, poverty and the global water crisis. http://www.undp.org/content/undp/en/home/librarypage/hdr/human-development-report-2006.html (accessed in February, 2015). UNEP-DTIE-IETC, 2000. International Sourcebook for Environmentally Sound Technologies for Wastewater and Stormwater Management. http://www.unep.or.jp/Ietc/Publications/TechPublications/ TechPub-15/2-4/4-1-3.asp (accessed in July, 2014).

ACCEPTED MANUSCRIPT USEPA, 2002. Onsite Wastewater Treatment Systems Manual. Cincinnati, Ohio. http://water.epa.gov/aboutow/owm/upload/2004_07_07_septics_septic_2002_ osdm_all.pdf (accessed in July, 2014). USEPA, 1999. Constructed wetlands treatment of municipal wastewaters. Manual. Cincinnati, Ohio. http://water.epa.gov/type/wetlands/restore/upload/ constructed-wetlands-design-manual.pdf (accessed in July, 2014).

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Vinnerås, B., Palmquist, H., Balmér, P., Weglin, J., Jensen, A., Andersson, Å., Jönsson, H., 2006. The characteristics of household wastewater and biodegradable waste – a proposal for new Swedish design values, Urban Water 3, 3–11. Wang, J., Rothausen, S.G.S.A., Conway, D., Zhang, L., Xiong, W., Holman, I.P., Li, Y., 2012. China’s water– energy nexus: greenhouse-gas emissions from groundwater use for agriculture. Environmental Research Letters 7, 14–35.

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Water Environment Federation, 2008. Operation of municipal wastewater treatment plants. Manual of practice n11. Sixth edition. Mcgraw Hill.

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WEPA : Water Environment Partnership in Asia. “Datasheets of Johkasou for residence“, http://www.wepa-db.net/technologies/individual/datasheet/jpn/18_hanematsu.htm (accessed in July, 2014). WHO/UNICEF, 2012. Progress on drinking water and sanitation 2012 – update. World Health Organization and United Nations Children’s Fund Joint Monitoring Programme for water supply and sanitation (JMP), Geneva, Switzerland. Wilderer, P.A., 2001. Decentralized versus centralized wastewater management. In: “Decentralized sanitation and reuse: concepts, systems and implementation”, Lens, P., Zeeman, G., Lettinga, G. (eds.), IWA Publishing. Meteorological Organization, World weather information service: http://www.worldweather.org/001/c00353.htm (accessed in July, 2014).

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Yang, X.M., Yahashi, T., Kuniyasu, K., Ohmori, H., 2001. On-site systems for domestic wastewater treatment (johkasous) in Japan. In: “Decentralized sanitation and reuse: concepts, systems and implementation”, Lens, P., Zeeman, G., Lettinga, G. (eds.), IWA Publishing.

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Yildirim, M., Topkaya, B., 2012. Assessing environmental impacts of wastewater treatment alternatives for small-scale communities. Clean-Soil, Air, Water 40, 171–178. Zhang, D., Gersberg, R.M., Keat, T.S., 2009. Constructed wetlands in China. Ecological Engineering 35, 1367– 1378.

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Zhang, Q.H., Wang, X.C., Xiong, J.Q., Chen, R., Cao, B., 2010. Application of life cycle assessment for an evaluation of wastewater treatment and reuse project – Case study of Xi’an, China. Bioresource Technology 101, 1421–1425. Zhao, R.F., Chen, X.P., Zhang, F.S., Zhang, H., Schroder, J., Romheld, V., 2006. Fertilization and Nitrogen Balance in a Wheat–Maize Rotation System in North China. Agronomy Journal 98, 938–945. Zhao, X., Liu, L., 2013. A comparative estimate of life cycle greenhouse gas emissions from two types of constructed wetlands in Tianjin, China. Desalination and Water Treatment 51(10–12), 2280–2293. Zeng, Q.,Wang, Y., Liu, H., Feng, L., 2008. Survey and Countermeasures on Latrine in Rural Areas of Tianjin. Journal of Preventive Medicine Information 10, 766–768.