Techno-environmental assessment of integrating polyhydroxyalkanoate (PHA) production with services of municipal wastewater treatment

Journal of Cleaner Production 137 (2016) 1368e1381

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Techno-environmental assessment of integrating polyhydroxyalkanoate (PHA) production with services of municipal wastewater treatment Fernando Morgan-Sagastume a, *, Sara Heimersson b, Giuseppe Laera c, Alan Werker a, d, €m b Magdalena Svanstro €ngsva €gen 11A, SE-226 47 Lund, Sweden Veolia Water Technologies AB (AnoxKaldnes), Klostera €gen 10, SE-412 96 Gothenburg, Sweden Department of Chemistry and Chemical Engineering, Chalmers University of Technology, Kemiva Water Research Institute, IRSA-CNR, Viale De Blasio 5, 70132 Bari, Italy d School of Chemical Engineering, University of Queensland, Brisbane, Queensland 4072, Australia a

b c

a r t i c l e i n f o

a b s t r a c t

Article history: Received 19 January 2016 Received in revised form 1 August 2016 Accepted 1 August 2016 Available online 4 August 2016

In this paper, the potential impacts in the techno-environmental performance of a municipal wastewater treatment plant with integrated mixed-microbial-culture polyhydroxyalkanoate (PHA) production are presented for the first time. A life cycle assessment was conducted based on mass and energy balances. The techno-environmental performance was evaluated for five wastewater treatment configurations: a reference case and four alternative processes producing PHA-rich biomass using influent municipal wastewater as the only organic carbon source. The integration of PHA-rich biomass production into a municipal wastewater treatment plant with sludge digestion sustains the overall conversion yield for total products of biogas and PHA-rich biomass (around 0.26 gCOD products per influent gCOD treated). PHA production integration has the potential to improve the overall environmental performance with respect to the reference case. Even when no benefits were accounted for substitutions related to the biogas and PHA-rich biomass, similar or improved environmental performances were estimated for all four alternatives for global warming potential, acidification potential, terrestrial eutrophication potential, and photo-oxidant formation potential. When benefits were accounted from substitutions of electricity and heat co-generated from biogas and of PHA-rich biomass by pure-culture PHA-rich biomass from sugar fermentation, gains were even higher due to the diversion of carbon from biogas to PHA-rich biomass. Freshwater and marine eutrophication potentials were dependent on effluent specifications. Case-by-case process configurations influence the mass and energy balance and trade-offs of process integration. The production and export of PHA-rich biomass decreased the aeration requirements for COD and nitrogen removal; however, increased demands for heat, power and chemicals were incurred for the generation of volatile fatty acids from primary solids fermentation. The choice of nitrogenremoval approach (nitrification-denitrification vs. anammox) also impacted energy consumption. Using influent wastewater as the sole carbon supply, the energy balance and PHA production were sensitive to the efficiency of primary treatment and available flux of volatile fatty acids into PHA production. Other regional inputs of organic residuals may improve carbon recovery in the treatment facility. The improved environmental performance of the treatment configurations motivates the idea that individual municipal wastewater treatment plants may become suppliers of renewable raw materials of higher value than that of biogas and/or energy and heat today. © 2016 Elsevier Ltd. All rights reserved.

Keywords: Polyhydroxyalkanoate (PHA) Life cycle assessment (LCA) Activated sludge Resource recovery Nitrogen/phosphorus management Bio-based

1. Introduction * Corresponding author. E-mail addresses: [email protected] (F. Morgan-Sagastume), [email protected] (S. Heimersson), [email protected] (G. Laera), [email protected] (A. Werker), [email protected] €m). (M. Svanstro http://dx.doi.org/10.1016/j.jclepro.2016.08.008 0959-6526/© 2016 Elsevier Ltd. All rights reserved.

During the past decade, the technical feasibility of the production of biodegradable thermoplastic polyesters, polyhydroxyalkanoates (PHAs), by open mixed microbial cultures

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Abbreviations AP CHP COD EP GWP LCA MAD MMC NREU PE PHA POFP sCOD TSS VFA VSS WWT WAS

acidification potential co-generation of heat and power chemical oxygen demand eutrophication potential global warming potential life cycle assessment mesophilic anaerobic digestion mixed microbial culture non-renewable energy use process element polyhydroxyalkanoate photo-oxidant formation potential soluble COD total suspended solids volatile fatty acid volatile suspended solids wastewater treatment waste activated sludge

(MMCs) has been repeatedly demonstrated using waste and residual carbon (C) sources as substrates (as reviewed in Laycock et al., 2013; Nikodinovic-Runic et al., 2013). These MMC systems have generally included three biological process elements (PEs): PE1 - acidogenic fermentation, PE2 - enrichment and production of biomass with PHA-storing capacity, and PE3 - PHA accumulation using PE2 surplus biomass and feedstocks with easily degradable organics (Fig. 1) (e.g., Serafim et al., 2008; Morgan-Sagastume et al., 2014). Polymers may be readily recovered from a PHA-rich mixed culture biomass in a fourth process element (PE4 - ArcosHern andez et al., 2015) by means of solvent extraction (Werker

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et al., 2014). MMC PHA production concurrent with meeting demands of municipal and industrial wastewater treatment represents a cornerstone for combining environmental engineering with the recovery of renewable resources. Biological treatment of wastewater and sludge management for recovering wastewater organic C as PHAs is a route to transform end-of-pipe environmental protection infrastructure into bio-refineries. Integration strategies for MMC PHA production within wastewater treatment processes have been proposed for industrial process wastewater (as in a sugar factory by Anterrieu et al., 2014) and municipal wastewater (Coats et al., 2011; Valentino et al., 2015) treatment. The technical feasibility of producing biomass with PHA-accumulation potential from municipal wastewater treatment (PE2) and fermented waste sludge (PE1) as feedstock for PHA accumulation (PE3) has been demonstrated at pilot scale (Morgan-Sagastume et al., 2014, 2015; Bengtsson et al., 2016). The technical advancements on MMC PHA production in the research and patent literature over the past 10 years anticipate benefits of producing excess activated sludge biomass yielding biopolymers as by-products of wastewater and sludge treatment. However, a full-scale production of mixed-culture PHAs integrated into waste management services does not exist today, and the environmental benefits that this route of resource recovery may bring remain to be more clearly understood. An environmental assessment of such process integration within an existing PHA value chain is therefore lacking in the literature. Notwithstanding, the technical impacts of such technology integration in a municipal WWT plant can be assessed by modelling the processes by means of mass and energy balances (e.g., Mininni et al., 2015), which can then be used to evaluate the potential environmental impacts of the technology integration. Life cycle assessment (LCA) has been used as a tool for evaluating the environmental performance of PHA-production processes by means of pure-culture bacterial fermentations, as reviewed by

Fig. 1. Generic schematic integration of MMC PHA production into wastewater and residuals management services with four different process elements (PE1 to PE4). PE4 can be ndez et al., 2015. The processes outside the dashed area were outside the scope of the centralised and serve several PHA-rich biomass production sites. Modified from Arcos-Herna present study. Optional stream flows are indicated by dashed lines.

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Heimersson et al. (2014). Energy consumption due to the production of refined sterile substrates or PHA recovery (Harding et al., 2007; Gerngross, 1999; Rostkowski et al., 2012), and the fermentation (Akiyama et al., 2003), have been identified as the principal contributors to the environmental impacts of pure-culture processes. Nevertheless, the published LCA results for PHA production are challenging to compare due to distinct differences in LCA methodological approaches and highly specific case-by-case assumptions in process considerations (Yates and Barlow, 2013; Heimersson et al., 2014). Consequently, the limited available published outcomes for the environmental performance of pure and mixed culture PHA production methods are unfortunately inconsistent to the point that conflicting conclusions have been brought forward. Several studies have found that pure-culture PHA production may have lower non-renewable energy use (NREU) and/or global warming potential (GWP) compared to the production of petrochemical polymers (e.g., Essel and Carus, 2012). However, when considering other impact categories, the outcomes have not been consistent (Yates and Barlow, 2013). For MMC PHA production using wastewater as substrate, three studies have addressed the question of environmental performance with comparative LCAs. MMC PHA production from the organic C content of industrial wastewater has been shown to be comparable to or to outperform biogas production in terms of GWP, while having higher GWP and NREU than a sugar-based pure-culture PHA production (Gurieff and ndez-Dacosta et al., 2015). The GWP of MMC PHAs Lant, 2007; Ferna was lower with reference to high-density polyethylene production (Gurieff and Lant, 2007), but it was generally higher in relation to the production of polyethylene terephthalate (Fern andez-Dacosta et al., 2015). Heimersson et al. (2014) underlined pitfalls and sources of misinterpretation due to LCA methodological choices. Allocation methods employed to account for system multifunctionality and the energy modelling, in particular, can significantly affect the outcomes of an environmental assessment (Heimersson et al., 2014). Therefore, the reporting of LCA results should provide a detailed and transparent description of assumptions and resultant biases. When comparing and drawing conclusions among LCAs on MMC or pure-culture PHA production and recovery methods, the underlying assumptions of these studies should be discussed. For example, the choice of energy use impact category, as the source (non-renewable or also renewable) and type (primary or secondary) of energy used, is sometimes poorly motivated or described (Heimersson et al., 2014). Also the quality and type of PHA produced is sometimes compared to conventional oilbased polymers without specifying whether this comparison is valid for the type of PHA that could be eventually produced given the potential diversity of PHAs and specific polymer applications. The choice of LCA allocation method is especially critical to specify when the PHA production is embedded within a context of providing concurrent essential services of waste (resource) management. ndezThe few published studies (Gurieff and Lant, 2007; Ferna Dacosta et al., 2015) of environmental impacts of MMC PHA production suggest that downstream processing or PHA recovery (PE4) contributes largely to the environmental impacts due to anticipated energy requirements. However, these LCA outcomes are not based on validated PE4 performance data of processes that are necessarily applicable to PHA recovery from MMCs. Furthermore, the assumptions on PHA recovery tend to overlook the critical influence of recovery methods on the PHA polymer type (i.e., PHA monomer composition and its origin from pure-culture or MMC biomass) and the requisite polymer quality (e.g., polymer properties such as thermal stability, average molecular mass, and melt colour) with respect to a well-defined application of the recovered materials.

Such considerations of the specific type of PHA coupled to the intent of application will determine the choice of PHA recovery methods and this will in turn influence LCA outcomes. Nevertheless, PE1, PE2 and PE3 can supply a PHA-rich-biomass within the boundaries of a WWT plant (Fig. 1), thus, transforming a WWT facility into an infrastructure for combined water quality improvement and renewable raw-material supply. The principal objective of the present investigation was to understand the most important technical and environmental opportunities and challenges involved with integrating MMC-PHA process elements PE1-3 into a full-scale reference municipal WWT process. We aimed to help answer the question of whether it makes sense from a technical and LCA perspective for a municipal WWT plant to become a supplier of PHA-rich biomass raw material. This raw material could, for example, be managed in a regional centralised biorefinery serving several suppliers in the region, upgrading PHA-rich biomass into biopolymers, platform chemicals and other valuable materials. To this end, we considered four alternative process configurations for municipal WWT. The influence of one strategy for integrating technology for MMC PHA production was considered on the technical and environmental performance of a WWT plant. The WWT plant receives municipal wastewater as an input raw material and generates outputs of treated effluent, digested sludge and PHA-rich biomass. The technical performance of each process configuration (reference and four alternatives) was evaluated from mass and energy balances. Mass and energy balances provided inputs to the evaluation of the environmental performance by means of an LCA. The four different alternative process configurations provided for a sensitivity analysis on the type of WWT process. The reported technical assumptions and performance results from the energy and mass balances and LCA were intended to improve the understanding about the impacts of integrating MMC PHA production into a municipal WWT plant. Since there is no industrial-scale production of MMC PHA today, and the commercial value chain of products and services associated to it is non-existent, we chose to consider just the supply of PHA-rich biomass as raw material for PHA production. We strove to minimise the risk of a speculative LCA by not including PHA recovery (PE4) in the assessment, as further explained and motivated later. Nonetheless, PHA recovery and associated LCA methodological considerations are discussed in the present investigation. The relation of PE4 to the other PEs is also considered regarding the potential transformation of a WWT plant into a supplier of PHA-rich-biomass as raw material of a regional biobased PHA value chain. 2. Approach and methods 2.1. Wastewater treatment (WWT) processes In total, five model full-scale process configurations for municipal wastewater treatment (WWT) were evaluated by means of mass and energy balances. A WWT process comprising aerobic biological treatment for C and nitrogen (N) removal and anaerobic digestion of sludge for biogas production was adopted as the nonPHA producing reference process configuration. The reference WWT process configuration (Reference) was based on established methods for wastewater and sludge management (Metcalf and Eddy, 1991; Qasim, 1994). This Reference was compared to four alternative WWT process configurations (Alternatives 1e4) that comprised integrated PEs for the MMC production of PHA-rich biomass as raw material for PHA production. The same municipal wastewater influent and treated-effluent water qualities were imposed in the mass and energy balances of all the process configurations. The mass and energy balances for the cases with MMC

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PHA production were evaluated based on data and practical experience from recent pilot-scale technology prototype operations (Morgan-Sagastume et al., 2015). 2.1.1. Treatment performance The WWT process was designed for treating an influent wastewater daily flow of 136,500 m3/d with a nominal capacity for 500,000 population equivalents. The wastewater loads were defined based on a specified model influent quality for a mediumstrength untreated domestic wastewater (Metcalf and Eddy, 1991) (Table 1). The effluent quality (Table 1) was selected based on the standards for a sensitive recipient according to the European Urban Waste Water Treatment Directive 91/271/EC. The Reference and Alternative WWT processes were modelled to treat the model influent wastewater and to comply with these effluent quality standards, which were used as the limits for maximum allowable discharge concentrations. The Reference process configuration generated digestate sludge from mesophilic anaerobic digestion (MAD) as the single solids output. When PHA-rich biomass production was integrated, both PHA-rich biomass and MAD digestate sludge were produced. All solids exports were dewatered to 30% (w/w) dry solids, as raw materials for off-site incineration (digestate) or PHA recovery (PHArich biomass). 2.1.2. Integration of mixed-culture PHA production PHA-rich biomass production comprised three main process elements (PEs) (Fig. 1) (Morgan-Sagastume et al., 2015; ArcosHern andez et al., 2015). PE1 converts primary sludge into a volatile-fatty-acid (VFA) rich liquid stream and fermented solids. PE2 converts influent municipal wastewater soluble organic matter into a surplus biomass with enhanced PHA-storage capacity. PE3 produces a PHA-rich biomass by exploiting the PHA-storage capacity of the harvested surplus PE2 biomass when fed with PE1 VFA-rich liquor. 2.1.3. Process configurations The five WWT process configurations are presented in Fig. 2. The Reference process encompassed: primary settling, activated sludge process (C removal and N removal by biological nitrification and methanol-based denitrification), chemical treatment (phosphorus (P) removal by precipitation with FeCl3 addition), sludge digestion (MAD of primary and waste activated sludge (WAS)), cogeneration of heat and power (CHP) from the biogas, and dewatering of the sludge digestate. Four alternative process configurations (Alternatives 1e4) integrated municipal WWT with the production of PHA-rich biomass. Alternative 1 (Morgan-Sagastume et al., 2014) included a similar process like that of the Reference with some modifications for advanced primary treatment (coagulant/flocculant addition), the production of PE2 functional biomass (or WAS), separate PE1 acidogenic fermentation of enhanced

Table 1 Characteristics of the model influent municipal wastewater and treated effluent quality standards considered for modelling all the WWT process configurations. The effluent quality standards were used as guidelines for maximum allowable discharge concentrations with which each WWT process was made to comply. Parameter

Influent (mg/L)

Effluent (mg/L)

Total suspended solids (TSS) Volatile suspended solids (VSS) Chemical oxygen demand (COD) Total nitrogen (TN) NHþ 4 -N NO 3 -N Total phosphorus (TP)

220 165 500 40 e e 5

14 10 50 10 0.5 8 1

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primary sludge, and PHA accumulation in a separate activated sludge PE3 post-process (C, and some level of N and P removal from the liquid fermented liquor). The remaining alternatives (2e4) entailed modifications to Alternative 1. In Alternative 2, the chemical-based advanced primary treatment of Alternative 1 was exchanged with sieving using microfiltration screens. Alternative 3 evaluated N removal by nitritation/anammox instead of conventional nitrification/denitrification, as a potential new process. Alternative 4 integrated nitrification and denitrification in PE2, thereby avoiding methanol addition. Alternative 4 integration of PE2 N removal has been successfully demonstrated at pilot scale in Leeuwarden, The Netherlands (Bengtsson et al., 2016) and has been benchmarked at full scale (2015e2016) in a Dutch national demonstration project called PHARIO (Topsector Energy program of the Dutch Ministry of Economic Affairs). 2.2. Mass and energy balances 2.2.1. Mass balance calculation Mass balances for the Reference and Alternative process configurations were performed according to the procedure described in Mininni et al. (2015). Each WWT process was modelled considering the water and sludge processing lines, and their reciprocal exchanges and interdependencies, such as recirculation of sidestreams. Plant-wide mass balances were iteratively solved for average conditions of load (Section 2.1.1) considering mass flow rate (Q), total and volatile suspended solids (TSS and VSS), chemical oxygen demand (COD), total N (TN), total P (TP), and the soluble fractions of COD (sCOD), and of TN and TP. Average process performance for each unit operation of the water and sludge lines were obtained either from the literature, from available PE1-PE3 laboratory-scale and pilot-scale data, or direct practical experience. Specific assumptions for modelling the Reference process are reported elsewhere (Tables 3 and 4 in Mininni et al., 2015), while the performance parameters for the Alternative processes have been summarised in Table 2. 2.2.2. Energy balance and chemical consumption calculations The consumption of electric power, thermal energy, and chemicals and energy generation (Table 3) were estimated based on mass flow rates for different unit operations in the water and sludge treatment lines of the five WWT process configurations (Fig. 2). Process performance factors and parameters in Tables 2 and 3 were derived from theoretical or empirical process data available in the literature (Metcalf and Eddy, 1991; Qasim, 1994), from PE1PE3 laboratory-scale and pilot-scale testing data (MorganSagastume et al., 2015), and from the direct practical experience of the authors. A nominal temperature of 25
C was assumed for the water and sludge lines entering all the process units. Heating requirements from 25 to 35
C were considered for the sludge lines entering PE1 acidogenic fermentation and MAD, which were modelled to operate at 35
C. All heating requirements were calculated considering heat losses of 4.5 kW h/m3 sludge (Metcalf and Eddy, 1991; Qasim, 1994). The electricity requirements for mixing during PE3 chemical treatment, and flocculation mixing for dewatering of the MAD effluent and the PE3 PHA-rich biomass were estimated based on a mixing power requirement equation considering velocity gradient, viscosity and volume (Qasim, 1994). MAD was considered to lead to a volatile solids (VS) reduction of 45% with a methane yield of 0.5 Nm3 CH4/kg VSreduced, and CHP conversion efficiencies of 35% for electric power and of 45% for heat were assumed, in keeping with typical specifications (Metcalf and Eddy, 1991; Qasim, 1994; Speece, 2008). Aeration with associated electric energy requirements for aerobic unit processes (activated sludge, PE2, PE3) were estimated following fundamental equations

Fig. 2. Schematic process flow diagrams of the five WWT process configurations evaluated in this study. Process stream flows with a high content of solids (sludge or biomass streams) are indicated by dashed lines.

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Fig. 2. (continued).

of oxygen transfer and uptake by activated sludge with diffusers (Metcalf and Eddy, 1991; Qasim, 1994): an a oxygen transfer factor of 0.5, a salinity surface tension factor b of 0.95, a standard oxygen transfer efficiency of 0.26 (for ceramic grids, rigid porous plastic tubes, or perforated membrane grids), and power requirements estimated based on an adiabatic blower (Qasim, 1994). Energy and chemical demands for the off-site incineration of sludge digestate were calculated following the published data presented by Larsen et al. (2010) (Table 3). 2.3. Life cycle assessment (LCA) The environmental LCA was performed according to recommendations in standards and guidelines (ISO 14040 and ILCD guidelines: European Commission Joint Research Centre, 2010) and using the software GaBi Professional 6. The assessed environmental impact categories were global warming potential (GWP), acidification potential (AP), eutrophication potential (EP) for freshwater,

marine and terrestrial ecosystems, and photo-oxidant formation potential (POFP). These impact categories are commonly assessed in these types of LCAs (Corominas et al., 2013) and were considered to be relevant also for the present investigation. Other impact categories that could have been relevant are human toxicity and ecotoxicity, but due to both the immaturity of impact assessment modelling for these categories and a lack of relevant data for the studied system and its different variations, these categories were not considered meaningful to include in the present work. Methods of characterization for the included impact categories were chosen based on recommendations from the ILCD Handbook (European Commission Joint Research Centre, 2011). The default method for climate impact assessment does not consider carbon dioxide of biogenic origin, and all C in wastewater and sludge was considered to be of biogenic origin. The LCA modelling included the WWT processes with the respective PEs, as described above (Fig. 2), as well as production of all process inputs (energy and chemicals) and sludge transport. The

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Table 2 Assumed performances for different unit operations in the Alternative (A) process configurations. Unit operation

Performances

WWT process

Source

Advanced primary treatment

Separation efficiency TSS 88%, VSS 88%, COD 70%, TN 28%, TP 90% Chemical sludge production 4 gTSS/gFedosed Separation efficiency TSS 60%, VSS 50%, COD 40%, TN 10%, TP 10% Solubilization VSS 35%, COD 17% VFA yield ¼ 0.2 gCODVFA/gCODin TSS post-separation efficiency ¼ 90% Removals sCOD 65%, TN 25% Biomass yield ¼ 0.3 gVSS/gCODrem TSS post-separation efficiency ¼ 65% Removals sCOD 80%, TN 70% Biomass yield ¼ 0.24 gVSS/gCODrem TSS post-separation efficiency ¼ 63% sCOD removal ¼ 90% Yields 0.15 gVSS/g sCODrem 0.30 gPHA/g sCODrem TSS post-separation efficiency Thickening 90% Dewatering 90% Biomass yield ¼ 0.60 gVSS/gNrem Biomass yield ¼ 0.10 gVSS/gNrem Chemical sludge production 6 gSS/gPrem VS reduction ¼ 30% TSS separation efficiency ¼ 95%

A1, A3, A4

Ødegaard, 2000; Helness et al., 2005

Microfiltration PE1

PE2

Modified PE2

PE3

Post-denitrification Nitritation-Anammox Chemical P removal (tertiary) MAD Dewatering digested sludge

functional unit was the treatment of the same influent wastewater flow rate with treatment to comply with effluent water quality standards for all process configurations (Table 1). Other WWT outputs comprised biogas, MAD sludge digestate and PHA-rich biomass. Biogas and PHA-rich biomass were further managed in the LCA using a so-called substitution approach to account for differences in functions and outputs between the five different WWT process configurations, as recommended for multifunctional systems according to the ILCD Handbook (European Commission Joint Research Centre, 2010). Substitution implies crediting the system under evaluation for functionalities beyond the function represented by the functional unit by using conventional options for the generation of similar basic functions. Construction of facilities, vehicles or other infrastructure was not included in the LCA since this normally contributes only a minor part of the total environmental impacts (e.g., Peters and Rowley, 2009), and since only small differences associated with construction are anticipated among the five investigated process configurations. The inlet boundary was the wastewater entering the WWT facilities; thus, the wastewater collection system was not considered. Biogas is generated to varying degrees in all the WWT process configurations. This biogas was modelled to be utilised on site for CHP. The produced electricity was assumed to replace EU-27production of grid electricity while the heat was assumed to replace thermal energy from natural gas combustion. Dewatered MAD sludge digestate was considered to be transported (50 km by truck) for off-site incineration. The utilization on site at the WWT plant of the electricity or heat potentially produced from off-site incineration of this digestate may be site dependent. Therefore, due to uncertainties about specific site conditions, off-site incineration was not considered to be a net energy producer for the WWT plant of the present work. As such, off-site incineration did

Ødegaard, 2000 A2

Rusten and Lundar, 2006; Ljunggren, 2006

A1-4

Morgan-Sagastume et al., 2015

A1-4

Morgan-Sagastume et al., 2015

A4

Veolia Water Technologies AB (AnoxKaldnes)

A1-4

Morgan-Sagastume et al., 2015

Metcalf and Eddy, 1991

A1, A2 A3 A1-4

Metcalf and Eddy, 1991 Henze et al., 2008 Paul et al., 2001

A1-4

Metcalf and Eddy, 1991; Speece, 2008 Metcalf and Eddy, 1991

not require any substitution. The PHA-rich biomass produced in Alternatives 1e4 was assumed to be dewatered to 30% for subsequent use in a PHA recovery facility (off site), but transport and subsequent treatment was not accounted for in this study. In such a PHA-recovery facility, a biopolymer product would be generated that would supply biopolymers and/or platform chemical products. The PHA-rich biomass was instead credited for the avoided production of the corresponding amount of PHA by pure culture fermentation, but no transport was considered for this flow since transport would be similar for both biomass types. MMC PHAs may have similar but also different properties and applications than PHAs currently marketed and produced from pure culture methods; however, pure-culture PHAs are the closest available PHAs comparable to MMC PHAs, and pure-culture PHA production methods are better described. The calculations associated with the pure-culture PHA production were drawn from the process described by Akiyama et al. (2003) using glucose from corn as substrate. To the best knowledge of the authors, Akiyama et al.'s work is one of the few sources of transparent process data from a pure-culture PHA production process usable in an LCA. Their mass and energy results have thus been widely (re)used in other studies, as stressed by Heimersson et al. (2014). Improved process data from pure-culture or even MMC PHA production may influence the substitution and overall assessment. Literature emission factors were used to calculate nitrous oxide and methane emissions from WWT (Foley et al., 2008; Westling, 2011) and from the release of the treated effluent into a receiving water body (Foley et al., 2010; IPCC, 2006a; IPCC, 2006b). Emissions from CHP were assumed as reported by Nielsen et al. (2010). Sludge incineration was modelled following Larsen et al. (2010). The data for the production of electricity, heat, chemicals and transportation were taken from the GaBi 6 Professional and the EcoInvent 3.1 databases.

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Table 3 Consumption/generation of energy and dosage of chemicals assumed for each unit operation in the energy balance calculations of the different WWT process configurations and the LCA. “A” stands for Alternative process configuration. Electric energy Unit operation Mixing Primary settling Advanced primary treatment PE1 Secondary treatment PE2 PE3 Microfiltration Primary treatment (Microsieves) Thickening Primary treatment (Rotary drum thickener) PE3 (Gravity belt thickener) Dewatering MAD (Decanting centrifuge) PE3 (Decanting centrifuge) PE1 (Belt filter press) Off-site Incineration

Consumption

WWT process

Sourcea

0.0005 kW/m3

e

0.006 kW/m3 0.001 kW/m3

Reference A1, A3, A4 A1-4 Reference A1-4 A1-4

0.01 kW h/m3

A3

Remy et al., 2014

0.833 kW/(ton TSS/h) 0.23 kW/(m3/h)

Reference A1-4

e

1.875 kW/(m3/h)

e

0.34 kW/(m3/h)

All A1-4 A1-4

Metcalf and Eddy, 1991 e

330 kW h/ton sludge TS

All

Larsen et al., 2010

Thermal energy Unit operation

Consumption

WWT process

Source

Off-site Consumption for incineration (30% dry solids content) Generation for incineration

820 kW h/ton sludge TS 620 kW h/ton sludge TS

All

Larsen et al., 2010

Chemicals Unit operation

Dose

WWT process

Sourcea

11 mg Fe3þ/L 3 mg PAM/L

A1, A3, A4

Helness et al., 2005

7.5 mg Al3þ/L

1

e

7.5 mg Al3þ/L 0.00495 kg AlCl3/kg TS

A1-4 All A1-4

e Ryu et al., 2000

11 mg Fe3þ/L

All

e

1.9 g MeOH/g NO 3 -Nrem

Reference, A1, A2

e

355 meq/pH/kg TS

A1-4

Morgan-Sagastume et al., 2015,b

0.070 kg/ton TS 27.7 kg/ton TS 0.65 kg/ton TS 0.05 kg/ton TS 0.64 kg/ton TS 33.78 kg/ton TS 1.62 kg/ton TS 2 kg/ton TS 1.2 kg/ton TS 0.59 kg/ton TS

All

Larsen et al., 2010

Advanced primary treatment FeCl3 as coagulant Polyacrylamide (PAM) Thickening AlCl3 (PAC) e Primary treatment Dewatering AlCl3 (PAC) e PE1 AlCl3 (PAC) e MAD AlCl3 (PAC) e PE3 Chemical P removal FeCl3 N removal Methanol (MeOH) as C source for denitrification Chemical treatment H2SO4 e PE3 Off-site incineration Flocculant Quarry sand Calcium chloride Metal capturing agent (TMT15/Na3T) Sodium persulphate Sodium hydroxide Ammonia Hydrochloric acid Sulphuric acid Sodium chloride a b

Assumed values for this work based on heuristics/professional experience are left blank (e) in the source column. Testing data generated during the referenced study, but not directly cited there.

3. Results and discussion This study evaluated the relative influence of integrating MMC PHA-rich biomass production with respect to a Reference WWT process configuration. A broadened perspective was brought forward by considering the sensitivity of techno-environmental impacts from four different Alternative process configurations. 3.1. Overall process performance The resultant overall levels of COD removals in all the modelled

WWT process configurations were similarly close to 63 ton COD removed per day (see variations in COD removed in Table 4). Among Alternatives 1e4, the biomass PHA content ranged from 0.40 to 0.47 g PHA/g VSS (Table 4). The minor differences in treatment mass balance results related to differences in process configurations (Fig. 2), in which the flux of C towards PHA production was influenced by the constraints of overall influent and effluent quality in combination with design and conversion specifications (Table 2). The estimated range of biomass PHA contents was in agreement with practical experience at pilot scale (and full scale) from municipal WWT (Morgan-Sagastume et al., 2015; Bengtsson

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et al., 2016). Estimated differences in output-stream mass flow rates of MAD methane and dewatered digestate, as well as of PHA-rich biomass (Table 4) were due to the modelled extent of primary treatment, the integration of PHA production, and the selected respective processes for nutrient removal (Fig. 2). The largest flows of methane and digestate were generated, not surprisingly, in the Reference WWT process configuration. The Alternative PHA-producing WWT process configurations generated less methane because of a diverted flux of organic C into PHAs. Between the Alternative WWT configurations, the predicted flow rates and yields of methane and digestate were similar, albeit marginally lower in Alternative 2 and higher in Alternative 4 (Table 4). The PHA production potentials among Alternatives were similar (2.5 ton PHA/d) except for Alternative 2 (1.4 ton PHA/d). Alternative 2 included microfiltration for primary treatment, and the lower PHA production potential for Alternative 2 was due to lower amounts of primary solids being directed to acidogenic fermentation. Primary treatment with micro-sieves was modelled with lower solids separation efficiency (60% TSS and 40% COD removals) when compared to advanced primary treatment using coagulants and flocculants with higher solids separation efficiency (88% TSS and 70% COD removal) (Table 2). The lower flux of organic C to PE1 and VFAs reduced the PHA-production potential in Alternative 2 because the PE3 feedstock (PE1 effluent) became limiting. This is a result of the model limitation of VFA production based solely on the available organic C from the influent wastewater. Assessment of PHA production through the use of other municipal, industrial, or agricultural organic residuals was beyond the scope of the present investigation, but it is foreseen as an effective means for managing and valorising organic residuals regionally. The available flux of VFA-COD as feedstock for PE3 directly impacts the PHA production potential of a municipal WWT plant. Nevertheless, the PHA-production potential of such a WWT plant could be enhanced with the import of regionally available organics as a C supply for PE1 VFA production while, at the same time, sustaining methane production, as further discussed below. The largest process energy demands were from diffused aeration (75e86%), followed by MAD (6e8%) and dewatering (5e8%)

(Fig. 3). This ranking of energy consumption fits well with published predictions that aeration, dewatering, mixing and pumping may account for as much as 70e75% of the total plant energy requirements for conventional municipal WWT plants (Bertanza et al., 2015). Energy demands associated with dewatering were similar among the configurations because the total flows of MAD digestate plus PHA-rich biomass were similar, and in the range from 18.8 to 23.8 ton VSS/d. Correspondingly, the requirements of AlCl3 for thickening and dewatering were also similar between configurations (Fig. 3). In the Alternatives, higher energy demands were incurred in mixing and thickening compared to the Reference due to additional mixing for advanced primary treatment (except Alternative 2) and for PE1, and due to an extra gravity belt thickener in PE3. While all the WWT configurations required a net energy input (electricity), only the Reference yielded a net excess of heat (Table 3) produced by CHP from MAD biogas. Heat and electrical power were both estimated to be produced in larger amounts in the Reference since more organic C was directed to biogas than in the Alternatives. In reality, the efficacy of heat utilization depends on the internal capacity of a WWT plant for heat reuse; excess heat would be useful only if delivered to a heating network off-site, which is not always the case. Therefore, the efficacy of heat utilization and its influence on the environmental performance in practice may require specific case-by-case consideration. 3.2. Impact of PHA-production integration into WWT process configurations The integration of MMC PHA production into a municipal activated sludge WWT process with primary and secondary treatment and MAD, such as the Reference process of this study, will introduce a balance of opportunity between producing just methane and diverting part of the wastewater organic C flux into higher value platform chemicals for bio-based products. A decrease in methane yield due to C flux to PHA production (Table 4) will influence the WWT mass and energy balances. However, the question is if diverting wastewater C flux from methane into PHAs brings a net environmental benefit to the WWT plant. From a wastewater C flux

Table 4 Summary of estimated mass flow rates of main outputs and calculated heating requirements for the Reference and the four PHA-producing, Alternative WWT process configurations. All configurations were assumed to treat a wastewater flow rate of 136,500 m3/d. Parameter

Reference WWT process

Alternative WWT processes 1

2

3

4

62.9 22.9 4.2 0 0 0

62.7 17.3 1.9 6.3 0.40 2.5

62.6 15.6 1.7 3.2 0.44 1.4

62.7 16.4 1.8 6.3 0.40 2.5

63.3 18.5 2.0 5.3 0.47 2.5

0.27 0.27

0.12 0.26

0.11 0.18

0.11 0.26

0.13 0.25

5900 8324 2424

10,122 3408 3773 9757

5638 3674 3299 6012

10,109 2490 3571 9028

10,138 4085 4017 10,206

Mass flows COD removed (ton/d) Digestate (ton TSS/d) Methane (ton CH4/d) PHA-rich biomassa (ton VSS/d) Biomass PHA content (g PHA/gVSS) PHA production potential (ton PHA/d) Yieldsb Methane yield (gCOD/gCODremoved) Overall yield for PHA-rich biomass þ methane (gCOD/CODremoved) Heating requirements (MWh/y)c Heating (PE1) MAD CHP Net a b c

Dewatered to 30% dry content. The PHA produced is assumed to be a 3HB-co-3HV with 3% 3HV mass content (Morgan-Sagastume et al., 2015). COD conversions based on: 4 gCH4-COD/gCH4; 1.42 gPHA-rich-biomass-COD/gVSS; 1.75 gPHA-COD/gPHA where PHA ¼ 3HB-3HV with 3% 3HV mass content. A negative value implies that heat is generated.

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Fig. 3. Calculated requirements of chemicals (a) and energy (b) for the Reference and the four PHA-producing Alternative WWT process configurations. Energy production from CHP is reported as negative numbers.

perspective, the global yields of influent WWT C into by-products as combined methane and PHA-rich biomass were found to be similar among the configurations (0.25e0.27 g COD/g CODremoved, Table 4). Therefore, one may conclude that since the C waste-toproduct conversion is similar, the question of environmental sustainability (and potential economic motivation) will depend on the role the final products will fulfil in regional value chains. PHA-rich biomass is a raw material comprising platform chemicals (PHAs, lipids, proteins, carbohydrates) that can be utilised for its chemical and energy contents. Recovered PHA polymers used for bioplastics and specialty renewable materials are considered to be higher up on the so-called value pyramid for biomass conversion. The pyramid base comprises lower valued energy/heat and fuel products like methane (Lange, 2014). The embedded complementary value of the non-PHA components of the PHA-rich biomass (energy,

lipids and minerals), which can be recovered (Werker et al., 2012, 2014) as part of the PHA recovery process, has not yet been considered in the scientific literature. The comparison of benefits from different routes of organic C conversion and recovery from WWT is complicated because biogas and digestate management technologies are so wellestablished within the industry (Kleerebezem et al., 2015). A mix of emerging opportunities, which divert some C flux away from methane into chemical products such as PHAs, can bring corollary benefits that are not fully recognised today within the current WWT mind-set. For example, the diversion of secondary activated sludge away from biogas digestion and into PHA production results in a decrease in the loading of nutrients associated with liquid streams in need of on-site management compared to the ones biogas production currently brings. The

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nutrients contained within the PHA-rich biomass are exported, and thus present not only an opportunity for value exploitation, but also a need for management as part of PE4. Additionally, the diversion of secondary activated sludge away from biogas and into PHAs frees on-site MAD capacity. This freed MAD capacity could be directed towards the VFA platform (PE1) or biogas since the WWT owners could consider to import other regionally available organic residuals for co-digestion with the benefit of sustained or increased biogas production (Mata-Alvarez et al., 2014). For simplicity of the present investigation, we considered no net change in the imported quantities of organic C to the WWT facility. Therefore, in the cases of the four Alternative configurations with a fixed wastewater influent C flux, the integration of PHA production into a Reference WWT system resulted in a similar net potential for conversion of input COD into COD as products distributed between methane and PHA-rich biomass (Table 4). Decreased methane production, but the generation of other value-added products, such as PHA-rich biomass to be exported as a value-added material, brings potential opportunities for regional renewable raw material supplies that do not exist today. Under the assumption of CHP from biogas of the present work, the integration of PHA production resulted in similar or even lower net levels of power consumption due to a decrease in electricity demands associated with aeration for C and N oxidation and sludge MAD (Table 4). On the other hand, net heat requirements increased due to heating requirements for acidogenic fermentation (PE1) of primary solids and the reduction in heat generation from biogas despite the decrease in heating requirements for MAD (Table 4). The PE1 sludge fermentation was assumed to be at 35
C, but fermentation is also feasible at lower temperatures, albeit with potentially reduced VFA yields (Pittmann and Steinmetz, 2013; Morgan-Sagastume et al., 2015). Operating PE1 sludge fermentation for VFA production at a lower temperature (e.g., 30
C) with a trade-off in VFA yields (for PHA production), leaving some C flux for MAD, may be one strategy to consider for the best overall C flux conversion, resource balance and environmental performance. Therefore, when multiple products are to be produced (e.g, PHArich biomass as well as methane), the optimization in the balance for one product or one individual process unit within the material flow may not necessarily provide the optimal net process tuning for renewable resource generation. Furthermore, the selection of a primary unit process (as in Alternative 2) and influent-wastewater C limitation may constrain the extent and effectiveness of C utilisation. However, opportunities for fluxes of organic C residuals into the WWT plant could eventually become attractive for optimising residuals management in a region. The extent of primary treatment and the process of choice for N removal were estimated to exert a relatively significant effect on the energy balance and demands of chemicals. Primary treatment (Alternative 2) compared to advanced primary treatment (Alternatives 1, 3 and 4) led to lower PHA production potential, as discussed above, but benefited the energy balance with respect to decreased heating requirements (Table 4). Advanced primary treatment implied additional resources for handling the larger resultant flow of advanced primary solids through the acidogenic fermenter (PE1) for VFA feedstock generation. Nonetheless, the largest estimated impact towards reducing power consumption was predicted by applying nitritation/anammox for N removal (Alternative 3). The combined biological N removal with COD removal and the production of excess biomass with PHA-storage capacity in a single PE2 process (Alternative 4) gave treatment with similar power requirements (as in Alternative 2), with the benefits of not requiring the supply of external C source, such as methanol, for denitrification.

3.3. Environmental impacts The results from the life cycle impact assessment for the different impact categories are given in Fig. 4, as the gross environmental impacts derived from contributions from the different parts of the WWT. Substitutions (negative bars in Fig. 4) were introduced in order to compensate for additional functions generated in the systems, i.e., CHP and the production of PHA-rich biomass, beyond the traditional base functions expressed by the functional unit. These substitutions were found to exert a strong influence on several of the impact categories; therefore, the results were also evaluated without the substitutions to assess for the sensitivity of the outcomes to the assumptions made. Looking at the environmental impacts without the substitutions, i.e., looking only at the bars above zero in Fig. 4, it is clear that even without accounting for any benefits from the substitutions, there are gains to or unchanged environmental performance for all the Alternatives in comparison to the Reference for all relevant environmental impact categories. The environmental performance results for EP-marine are dominated by influences from N discharges and for EP-freshwater by P discharges in the effluent (Fig. 4). Since N and P effluent levels were a specified effluent quality standard (Table 1), the impacts for these two categories for the Reference and the four Alternative process configurations were essentially the same and comparisons are not relevant in this context. For the other impact categories, the positive outcome is a combined effect of a decrease in electricity use, and in the amounts of biogas for combustion and of sludge for incineration. The Alternative configurations were estimated to incur a lower electric energy use and this was mainly due to reduced aeration requirements (Fig. 3); aeration requirements decreased with reduced loading of COD and N to PE2. The reduced loads were owing to the diversion of COD solids towards acidogenic fermentation as a result of advanced primary treatment, and to the export of nutrients with the PHA-rich biomass from the WWT plant. The Alternative configurations produced less biogas and sludge for incineration with lower associated emissions. The outcome is not changed by the increase in the use of thermal energy for larger flows of advanced primary solids through the acidogenic fermenter (PE1) (Table 4). Nor is it changed by the varying effects of changes in chemical consumptions that were estimated for the Alternatives. Comparing the substitutions between the Reference and the Alternative PHA-generating configurations, some C flux shifted from methane towards PHA, decreasing the gain from CHP from biogas, but generating gains due to PHA generation. However, these estimated effects also relate to how the substitutions were modelled, i.e., what these additional functions were considered to replace. Biogas was considered to be used on site for combustion and CHP, and the substitutions of EU-27 electricity and thermal energy from natural gas combustion were deemed both realistic and robust in terms of availability of high-quality data. However, methane can also be used as a vehicle fuel, which is common in some countries like Sweden. The selected approach for biogas valorization will also influence scores of electric power consumption and heating requirements and, consequently, the LCA results. This choice of methane utilization for either CHP or vehicle fuel has been shown to lead to different LCA outcomes in a food-waste MAD system producing biogas and refuse for incineration (Carlsson et al., 2015). Carlsson et al. (2015) found that GWP impacts were more dependent on the conversion efficiencies of biogas into heat and power for the CHP case rather than on the marginal energy considered for refuse incineration in the case of vehicle fuel. In € m et al. (2016) showed that the assumed use of addition, Svanstro

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Fig. 4. Life cycle impact assessment results for the Reference and the four Alternative WWT process configurations given as partial contributions for six environmental impact categories: GWP ¼ global warming potential, AP ¼ acidification potential, EP ¼ eutrophication potential, and POFP ¼ photo-oxidant formation potential. Wastewater processing ¼ impacts from nitrous oxide and methane emissions from biological wastewater treatment. Sludge processing ¼ impacts from methane emission from MAD.

biogas for either vehicle fuel or CHP was influential to the overall LCA results of sludge management. Therefore, case-specific evaluations are anticipated to be necessary in practice for fine-tuning assessments and understanding gain opportunities. The substitution for the MMC PHA-rich biomass with pureculture PHA-rich biomass relies on the assumptions that a centralised PHA recovery facility in the region will receive and further manage the flow of PHA-rich biomass, and that transport, PE4 residuals management, and the final use of all recovered products are comparable. The PHA content in the biomass from Alternatives 1e4 from this study is lower than that in the assumed pure-culture PHA-rich biomass (contents of 50 and up to 80%, gPHA/gTS), and the substitution was done on a kg to kg of PHA basis. Therefore, the substituted impact may be underestimated because both recovery and transport are likely done at a lower environmental cost for the pure-culture biomass than for the MMC PHA-rich biomass generated in this study. The evaluation of the purposeful utilization of non-PHA residuals (which become larger, the lower the biomass PHA content is) and of methods for PHA recovery that can more readily valorise them was outside the scope of the present study. Nevertheless, such evaluation is understood to affect and be relevant to consider in an LCA with the specific focus on a facility recovering PHAs from PHA-rich biomass sources. The PHA recovery process and implications of different product types, qualities and applications for this study are further discussed in the following section; however, we can state that the substitution done for the MMC PHA-rich biomass bears some uncertainty that is related to the type of system that is replaced (pure-culture PHA production) and the quality of data describing this system. This uncertainty reflects especially the general scarcity of full-scale process data for the production of PHA-rich biomass and PHA recovery. Notwithstanding, whether substitutions are included or not, the environmental impacts were lower for the Alternative process configurations than for the Reference configuration (Fig. 4). Overall, this LCA revealed potential for environmental benefits

when producing PHAs in a municipal WWT process compared to a Reference system with only biogas production. The net environmental impacts were up to 40% less for GWP, 70% less for AP, 90% less for terrestrial EP, and 75% less for POFP., when considering the four Alternatives for each category with respect to the Reference. Alternative 3 was predicted to have the lowest environmental impacts among the PHA-producing Alternative configurations. Disregarding substitutions, the other PHA-producing Alternatives ranked, from lowest to highest impacts, as Alternative 2, 1 and 4, whereas with substitutions, they ranked as Alternative 1, 4, and 2. Electricity consumption for primary-treatment micro-filtration in Alternative 2, combined with a lower PHA-yield, contributed to the relatively poorer environmental performance for this Alternative process configuration. Conversely, lower electricity use in the case of applying an anammox process for N removal benefited Alternative 3. Overall, electricity consumption led to the largest contributions of environmental impacts in the categories of AP, EP and POFP (Fig. 4). Electricity together with N2O and methane emissions (wastewater processing line) contributed greatly to GWP for all the process configurations (Fig. 4). Therefore, processes leading to decreased overall electricity consumption, such as with Alternative 3, would naturally improve the general environmental performance of a WWT plant. 3.4. PHA-rich biomass from WWT for PHA production in a regional bio-based economy In a bio-based economy, the ultimate value of the by-products from WWT C conversion and resource recovery is anticipated to be dictated by the position of these products on the value pyramid of renewable resources (Lange, 2014). From the results of the present investigation, the challenge for a WWT facility to engage in PHA-rich biomass production may not relate to changes in environmental performance. Instead, the challenge may lie on initiating PHA-mediated value chains that can bring benefits for a WWT

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plant. Apart from a relatively higher value of PHA polymers as specialty platform chemicals on the biomass conversion value pyramid (Lange, 2014), there is currently no commercial-scale production of MMC PHAs, let alone production integrated into municipal WWT. While PHA production integrated into municipal wastewater quality management is promising in theory, the predicted benefits have not been brought into practice today as there is no commercial-scale flow of PHA-rich biomass within biopolymer value chains. Such value chains would require a regionally established facility to process PHA-rich biomass. The Alternative configurations (1, 3 and 4) produce in the order of 1000 ton PHA per year. A regionally centralised PHA recovery facility may benefit from scales of economy, and therefore from the operation within a network of PHA-rich biomass inputs from different municipal and industrial WWT plants. Hence, the most appropriate PHA recovery processes are anticipated to be those that can accommodate a variety of biomass sources, biomass PHA contents and PHA types (Werker et al., 2014), and that are readily scalable at a starting capacity in the order of 1000 ton PHA per year. From the present study, we found that it makes similar sense to produce biogas or the combination of biogas and PHA-rich biomass. This outcome was found based on the assumption that the PHArich biomass will be used (not wasted) as a raw-material from the WWT plant within a regional network of PHA-rich biomass supplies (Fig. 1). The WWT plant has the potential to be a regional supplier of PHA-rich biomass, but it will not be likely in the business of being a polymer producer. PE4 therefore deserves its own independent techno-environmental evaluation that could tackle the question of whether it makes sense environmentally to establish a centralised PHA recovery facility that could process and valorise a supply of PHA-rich biomass from a regional network. Furthermore, the techno-economic viability of such PE4 PHA recovery facility could be also addressed, although with the added complexity of lack of prices on MMC PHAs since they are nonexistent in the market nowadays. The nutrient and organic contents exported as PHA-rich biomass from the WWT plant and handed over to a regional PHA-recovery facility (PE4) require management, as for example in the case of centralised sludge management facilities today. This management is not only with respect to the PHA recovery and valorization but also to the exploitation of the resource value embodied in the non-PHA biomass. The non-PHA residual biomass from PHA recovery presents a latent opportunity for the recovery of other chemicals and/or energy as well as nutrients. Nevertheless, these opportunities need to be evaluated with respect to the choice of PHA recovery method and associated inputs and outputs to the PE4 system boundaries. While some studies have addressed the challenges of PHA recovery from biomass (Yu and Chen, 2006; ndez-Dacosta et al., 2015), such studies do not generally Ferna consider the critical influence of recovery methods on the specific type of PHA (Werker et al., 2012, 2014), nor is there substantive consideration of the quality of the polymers with respect to its intended ultimate application. Given that MMC PHAs are not commercially available today, fullscale validation performance data from a PE4 process for MMC PHA recovery are unavailable. Hence, large uncertainties may be associated with LCAs considering MMC PHA recovery, and even if a particular PHA product could be established, there would still be some uncertainty related to the choice of allocation methodology, as discussed by Heimersson et al. (2014). The specific applicability of given PHA recovery methods to MMC PHAs with regards to polymer type, quality, and defined applications remains largely undefined in the scientific literature and specifically in the currently available LCA literature on MMC PHA production (Gurieff

ndez-Dacosta et al., 2015). Accordingly, curand Lant, 2007; Ferna rent LCA studies on MMC PHA production might be best understood as indicative of anticipated “hot-spots” requiring attention towards improving the environmental performance of these processes. One may further anticipate that the LCA insights will evolve with the benefit of more context-relevant process-performance data being derived from a growing number of pilot-scale developments (Tamis et al., 2014; Morgan-Sagastume et al., 2014, ndez et al., 2015; Bengtsson et al., 2016) and 2015; Arcos-Herna within case-study demonstration projects. 4. Conclusions PHA-rich biomass production integrated into municipal WWT can result in an improved WWT plant environmental performance. For four alternative process configurations integrating PHA-rich biomass production, a consistent reduction in environmental impacts was estimated compared to a Reference plant. With LCA modelling with substitutions for biogas CHP and MMC PHA-rich biomass, less net environmental impacts were estimated for the four Alternative configurations of up to 40% less for GWP, 70% less for AP, 90% less for terrestrial EP, and 75% less for POFP. Reduced electricity consumption in the Alternative processes, especially when using anammox for N removal (Alternative 3), has the potential to improve the WWT plant environmental performance. From a technical perspective, the integration of PHA-rich biomass production can provide for a combination of biogas and PHA-rich biomass by-products without sacrificing the overall conversion of influent wastewater organic C into products (around 0.26 gCOD as products per gCOD treated). The production and export of PHA-rich biomass from the WWT plant decreases the aeration requirements for COD and N removal, but these benefits are counterbalanced by increased demands for heat, power and chemicals in the handling of primary solids for VFA-rich feedstock production. The net electricity consumption will be sensitive to the type of N removal technology applied. In general, and with only the influent wastewater as an organic C source, the energy balance and PHA production potential is sensitive to the efficiency of primary treatment and to the flux of VFA-COD feedstock to PHA accumulation. Thus, the global outcomes in the techno-environmental performance are likely to be influenced by the existing WWT infrastructure intended to be integrated for PHA-rich biomass (and biogas) production, alongside specific approaches for biogas valorization. This study has assumed that the WWT plant could be a local raw-material supplier within an existing regional bio-based network. However, such regional networks demanding PHA-rich biomass as a raw material do not exist today. The outcomes of environmental impact evaluations with the process configurations of the present work do motivate the idea that individual municipal WWT plants could become suppliers of renewable raw materials of higher value other than just biogas and/or energy and heat. The environmental performance of a regional network of raw material suppliers feeding a refinery within a chemical product value chain remains to be explored and understood. Acknowledgements Part of this work was funded by the EU ROUTES project (Contract No 265156, FP7 2007-2013, THEME [ENV.2010.3.1.1e2] Innovative system solutions for municipal sludge treatment and management), and by The Adlerbert Research Foundation. References Akiyama, M., Tsuge, T., Doi, Y., 2003. Environmental life cycle comparison of

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