a critical review on sustainability assessment of recycled water schemes

Science of the Total Environment 426 (2012) 13–31

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Science of the Total Environment

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Review

A critical review on sustainability assessment of recycled water schemes

Zhuo Chen, Huu Hao Ngo ⁎, Wenshan GuoSchool of Civil and Environmental Engineering, University of Technology Sydney, Broadway, NSW 2007, Australia

Abbreviations: AAS, atomic adsorption spectrophotomicrofiltration and disinfection; DAR, depletion of abioinput–output; EP, eutrophication potential; ERA, envirochromatography; GWP, global warming potential; HRA,assessment model; LC, liquid chromatography; LCA, lifematerial flow analysis; MIET, missing inventory estimatienvironmental concentration; PNEC, predicted no effectSP, salinisation potential; TEP, terrestrial ecotoxicity; UF,plant.⁎ Corresponding author. Tel.: +61 2 9514 2745; fax:

E-mail address: [email protected] (H.H. Ngo).

0048-9697/$ – see front matter © 2012 Elsevier B.V. Alldoi:10.1016/j.scitotenv.2012.03.055

a b s t r a c t
a r t i c l e i n f o
Article history:Received 8 February 2012Received in revised form 20 March 2012Accepted 22 March 2012Available online 20 April 2012

Keywords:Recycled waterEnvironmental assessment toolsIntegrated approachesMulti-criteria decision making framework

Recycled water provides a viable opportunity to supplement water supplies as well as alleviateenvironmental loads. To further expand current schemes and explore new recycled water end uses, thisstudy reviews several environmental assessment tools, including Life Cycle Assessment (LCA), Material FlowAnalysis (MFA) and Environmental Risk Assessment (ERA) in terms of their types, characteristics andweaknesses in evaluating the sustainability of recycled water schemes. Due to the limitations in individualmodels, the integrated approaches are recommended in most cases, of which the outputs could be furthercombined with additional economic and social assessments in multi-criteria decision making framework. Thestudy also proposes several management strategies in improving the environmental scores. The discussionand suggestions could help decision makers in making a sound judgement as well as recognising thechallenges and tasks in the future.

© 2012 Elsevier B.V. All rights reserved.

Contents

1. Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 142. Environmental assessment tools on recycled water schemes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 15

2.1. Material Flow Analysis (MFA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 152.1.1. Types of MFA models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 152.1.2. Application of MFA models on environmental sanitation improvement . . . . . . . . . . . . . . . . . . . . . . . . . . . . 152.1.3. Characteristics and weaknesses of MFA models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 17

2.2. Life Cycle Assessment (LCA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172.2.1. Types of LCA models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 172.2.2. Application of LCA models on recycled water schemes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 192.2.3. Characteristics and weaknesses of LCA models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

2.3. Environmental Risk Assessment (ERA) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 212.3.1. Types of ERA models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 222.3.2. Application of ERA models on recycled water schemes . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 222.3.3. Characteristics and weaknesses of ERA models . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24

3. Integrated assessment tools on recycled water schemes. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 243.1. MFA coupled with LCA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 253.2. LCA coupled with ERA . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 263.3. Multi-criteria decision making . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 26

3.3.1. Economic assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 273.3.2. Social assessment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 273.3.3. MCA implementation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

meter; AP, acidification potential; CAS, conventional activated sludge; CED, cumulative energy demand; CMF, ozonation,tic resources; ECOSAR, ecological structure activity relationships; EIA, environmental impact assessment; EIO, economicnmental risk assessment; ETP, ecotoxicity potential; FAETP, freshwater aquatic ecotoxicity; FWU, freshwater use; GC, gashuman health risk assessment; HTP, human toxicity; ICP, inductively coupled plasma; KOREOCORisk, Korea ecological riskcycle assessment; LCIA, life cycle impact assessment; MAETP, marine aquatic ecotoxicity; MBR, membrane bioreactor; MFA,on tool; MS, mass spectrometry; ODP, ozone depletion potential; ORWARE, Organic Waste Research model; PEC, predictedconcentration; PHO, photochemical oxidation; RO, reverse osmosis; RQ, risk quotient; SAR, structure activity relationships;ultrafiltration; WEST, water-energy sustainability tool; WSP, wastewater stabilisation pond; WWTP, wastewater treatment

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14 Z. Chen et al. / Science of the Total Environment 426 (2012) 13–31

4. Conclusion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29Acknowledgement . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 29

1. Introduction

As far as the early 2000s, many cities and regions of developedcountries including North America, Australia, the Middle East, theMediterranean, Asia and Africa, have considered recycled water as analternative water resource to combat water scarcity issues associatedwith population increase, surface water quality deterioration, ground-water depletion and climate change. With increasing knowledge andunderstanding on the merits of recycled water (e.g., alleviation of thepressure on existing water supplies, reduction of effluent disposal tosurface and coastal waters and provision of more constant volume ofwater than rainfall-dependent sources), some planned recycled waterschemes have also been observed in many developing countries,especially in intensive agricultural areas (Fatta-Kassinos et al., 2011a).Moreover, the technical possibility and economic affordability toproduce recycled water of virtually drinking water quality have evenbroadened its application from non-potable uses (e.g., irrigation,industry, environmental flow, residential use, etc.) to indirect anddirect potable reuses (Rodriguez et al., 2009). Currently, thousands ofrecycling projects and pilot studies are being carried out worldwidewith many more in the planning and construction stages.

With the water recycling targets being more aggressive, long-termsustainability of the recycled water scheme becomes critical forfurther project expansion and new end use exploration. The currentenvironmental assessment models are playing vital important roles infast and reliable evaluation of existing or future recycling schemesfrom a perspective of environment-related considerations. Severalstudies have applied Material Flow Analysis (MFA) to calculate thesystematic material flow of pollutants and nutrients in environmentalsanitation systems over a given period of time whereas others haveused Life Cycle Analysis (LCA) to identify environment-related issuesof different wastewater treatment technologies or water resources onthe ecosystem and natural resources in life cycle. Since risk is one ofthe determinative factors to the success of recycling schemes, theEnvironmental Risk Assessment (ERA) studies have also been con-ducted to analyse the potential environmental risks (e.g., excessivepharmaceuticals and xenobiotic compounds on soil, surface water

Legal and institutional framework

Stakeholders’ roles and

responsibilities

Householders’ perceptions and

needs, policy makers’ views, etc.

Infrastructure, managerial and

financial arrangements

Scenarios development

Wastewater management

…e

…

Healthimpac

Fig. 1. Proposed procedure for multi-stakeholdModified from Agnes et al. (2007).

and groundwater) resulted from recycled water projects (Ahmed,2007; Meinzinger et al., 2009; Urkiaga et al., 2008). While mostenvironmental studies have been carried out using a single environ-mental tool, the integrated models have been increasingly developedto compensate the weaknesses of individual ones (Benetto et al.,2007; Jeppsson and Hellstrom, 2002; Montangero et al., 2006a).However, the investigations on the selection and implementation ofappropriate integrated models for particular recycling schemes arestill limited and not well documented.

Besides, failure to manage the existing recycling systems optimal-ly has also introduced many challenges to local areas. For example,some of the schemes sought to achieve great environmental savingsthrough minimising water consumption and maximising waterrecycling regardless of utility, economy and feasibility in particulargeographical conditions whereas other recycling activities thatcaused the degradation of ecological habitats might have political orfinancial underpinnings (Chapagain and Orr, 2009; Ku-Pineda andTan, 2006). A systematic and comprehensive assessment in theproject planning or management processes could be significant whichis to investigate the trade-off among a variety of issues (e.g., ambientecosystem, engineering feasibility, infrastructure cost, energy con-sumption, recycled water pricing policies, community attitudes, etc.).Fig. 1 outlines how these different aspects can coordinate with eachother and finally contribute to the environmental sustainability(Agnes et al., 2007; Levine and Asano, 2004; Meneses et al., 2010).

Consequently, this review aims to summarise and discuss thecharacteristics and applications of each assessment model in differentrecycled water schemes. The authors emphasise the importance ofintegrated assessment approaches and the need to combine withother non-environmental-related models when further holistic plan-ning analysis or management is required. Additionally, the authorsstress contextual complexities inmodel integration that have not beengiven sufficient attention, yet which compromise the achievement ofdesired environmental sustainability improvements. The paper con-cludes by highlighting knowledge gaps (e.g., system boundary, modelintegration and performance) that future researchers may need toaddress.

Acceptance by users

Costs and socio-conomic benefit

Impact assessment

Environmental impact and

resource recovery

Environmental sustainability

Environmental sanitation

t

er environmental sustainability planning.

15Z. Chen et al. / Science of the Total Environment 426 (2012) 13–31

2. Environmental assessment tools on recycled water schemes

2.1. Material Flow Analysis (MFA)

As MFA examines the material flows and their transformation inregional environmental systems over a given period of time, itaddresses the importance of water recycling by linking adverseenvironmental impacts with possible resource recovery and reusesolutions (Brunner and Baccini, 1992; Jeppsson and Hellstrom, 2002).Generally, MFA consists of four steps: (1) definition of a systemwhichcomposes of material flows, stocks and processes; (2) measurementof mass fluxes and element concentrations of all goods; (3)calculation of the element fluxes; and (4) schematic presentationand interpretation of the results (Sinsupan, 2004). Based on the lawof the conservation of matter, environmental impacts of a particularflow can be calculated by a simple mass balance of all associatedinputs, outputs and storage. The results can then be interpretedagainst environmental standards or can be linked to other assessmenttools for further analysis.

2.1.1. Types of MFA modelsQualitative MFA models are simple and can be quickly performed

which help decision makers understand the metabolism of theirregion and provide early warning signals for future environmentalissues. Nonetheless, it can only be regarded as an initial assessment as

Industries Hou

On-site sanitation

Open drainage

Surface water, groundw

Ex

Slud

geWas

tew

ater

(a)

Industries

Oimu

NWW

const

Surface water, groundw

WW

Greywatertreatment

Gre

ywat

er

Ind. WWTP

Tre

ated

WW

Rec

ycle

d W

W

(b)

Fig. 2. Simplified systems representing nitrogen flows in the current (Modified from Agnes et al. (2007).

numerical material flow data are not available (Agnes et al., 2007;Schneider et al., 2002). Comparatively, quantitative MFA modelsnormally employ mathematical equations to quantify the processesand flows of transformation, production and consumption as well asthe mass and/or water balance within the system, which offers morespecific and reliable information in decision making. Depending onthe variance of the flows over time, they can be further classified intostatic and dynamic forms. The static model, where the flows areassumed to be invariable, is suitable to estimate the flows with noprimary data and calculate the effectiveness of adopting differentpolicy scenarios in sustainability improvement. When the system isfound to be unsustainable, the model is unable to tell when it becameunsustainable due to its high uncertainties. On the other hand, thedynamic approach accounts for time dependence and analyses theflows of materials or any accumulation in stocks over a period of timebased on mathematical probabilistic distributions (Jeppsson andHellstrom, 2002; Park et al., 2011; Tangsubkul et al., 2005a).

2.1.2. Application of MFA models on environmental sanitationimprovement

MCA models have been increasingly applied to environmentalsanitation planning in several developing countries such as Columbia,Ghana and Vietnam (Belevi, 2002; Binder and Pazel, 2001; Brunnerand Baccini, 1992; Gumbo, 2005; Huang et al., 2007; Neset et al.,2006). With the water supply, sanitation, solid waste management

seholds

ater, soil

Agr

icul

ture

, aq

uacu

ltur

e

creta

Gre

ywat

er+

excr

eta

Eff

luen

t

Households

n-site sanitation (e.g. proved septic tank or

rine diversion latrines)

eighbourhood TP (e.g. ponds,

ructed wetlands)

ater, soil

Agr

icul

ture

, aq

uacu

ltur

e

Excreta

Eff

luen

t

Slud

ge

a) and improved (b) sanitation system in urban areas in Vietnam.


and urban agriculture being considered in an integrated way, they areable to identify the key flows or processes associated with huge waterconsumption and serious environmental pollution. For instance,Agnes et al. (2007) evaluated the current environmental sanitationsystem in Vietnam as well as the effects of new sanitation concepts ormeasures. Fig. 2 depicts two conceptual MCA models where thethickness of arrows indicated the relative importance of the flows.The small scale decentralised wastewater treatment facilities (e.g.,constructed wetlands or waste stabilisation ponds) for greywatertreatment were proved to be effective in water sustainabilityimprovement as treated water can be reused in agricultural irrigationand aquaculture and the amount of open drainage can be greatlyreduced. The study also suggested treating industrial wastewaterseparately from domestic sources and then reusing it internally.Likewise, Schneider et al. (2002) analysed the metabolism of waterwithin a region of Portugal. As can be seen from Fig. 3, significantimprovements in water sustainability can be achieved from lesswater consumption, increased internal or external water recyclingand reuse, reduced wastewater discharge, lower variability, etc.Although several recycled water end uses (e.g., agriculture, house-hold, industry and services) have been proposed in this water flowanalysis, future more detailed quantitative assessments are required.

Further, Sinsupan (2004) applied a static MFA model to investi-gate the nitrogen (N) fluxes in wastewater and organic wastes inenvironmental sanitation planning at Pak Kret, Thailand, wherearound 1900 kg/day of N from wastewater and septage wasdischarged into the environment. To improve this situation, twonew sanitation scenarios were put forward. The scenario 1 is to instalthe wastewater treatment plants (WWTP) and reuse the treated

Fig. 3. Metabolism of water in thAdapted from Schneider et al. (20

effluent and sludge for peri-urban agriculture whereas scenario 2 is toinstal the WWTPs and compost the municipal organic solid wastes.According to the mass balance calculations in MFA, the scenario 1 wasable to reduce the chemical fertiliser consumption by 57% and the Nloading in wastewater by 31%, compared with 51% and 45% inscenario 2 respectively. Likewise, Tangsubkul et al. (2005a) investi-gated the phosphorus (P) and water reuse management strategies inthe Sydney region for the year 2000. The results demonstrated thatthe combination of greywater recycling, composting toilet andhuman behaviour change (e.g., using P-free detergent and adoptinga vegetarian diet) was the most effective solution since around3600 tonnes/year of P can be prevented from entering the wastewa-ter system. Besides, nearly all of the P in wastewater could berecovered in this combined system. Despite the data gaps in socio-cultural, economic and health issues, both of the studies haveidentified the importance of conducting wastewater treatment andreuse in environmental sustainability. However, when temporalchanges of the flows are of interest, the dynamic approach shouldbe adopted rather than static ones.

Montangero and Belevi (2008) reported three important mathe-matical equations in dynamic MFA approach. The most essential oneis the mass balance equation:

dM jð Þi

dt¼ ∑

rAi;r−j−∑

sAi;j−s ð1Þ

where i is the indicator substance, j is the process number, Mi(j)

is thestock of substance i in process j, t is the time, r is the source process, sis the destination process, Ai,r− j is the input flow of substance i from

e socio-economy of a region.02).

Table 1Major environmental impact categories in LCA.a

Type Impact category or indicator Units

Squandering ofresources

Depletion of abiotic resources (DAR) kg antimony eqBiotic resources year−1

Pollution Acidification potential (AP) kg SO2 eqGlobal warming potential (GWP) kg CO2 eqEutrophication potential (EP) kg PO4 eqPhotochemical oxidation (PHO) kg formed ozoneOzone depletion potential (ODP) kg CFC-11 eqEcotoxicity potential (ETP) kg 1,4-DCB eqFreshwater use (FWU) m3

Cumulative energy demand (CED) MJOdour m3 polluted airNoise Pa2·s

Affection Ecosystem and landscape m2·sDeath –

a Modified from Andrae (2010), Pasqualino et al. (2010), Rebitzer et al. (2004), andTukker (2000).


the source process r to the destination process s. The left side of theequation represents the stock change rate of substance i within theprocess j while the right side expresses the difference between inputand output flows of substance i to and from process j. Ai,r− j can bederived as follows:

Ai;r−j ¼ f p1;p2…pnð Þ ð2Þ

where p1, p2…pn refer to parameters based on scientific and expertknowledge. Additionally, transfer coefficient is also commonly usedin modelling material flows, which describes the partitioning of asubstance in a process and provides the fraction of the total input of asubstance transferred to a specific output good (Eq. (3)).

k jð Þi;g ¼ A jð Þ

i;g=∑rAi;r−j: ð3Þ

The study also addressed the importance of expressing the modelinputs as probability distribution when limited data are given.

Based on these equations, Montangero et al. (2007) carried out acase study to evaluate water and nutrient management strategies(e.g., household consumption patterns, type of sanitation infrastruc-ture and wastewater reuse practices) in Hanoi, Vietnam. The modelindicated that reusing a fraction of greywater for toilet flushing wouldreduce the water consumption from 140 L to 113 L per capita per dayby 2015, which was tantamount to a 16% decrease in groundwaterabstraction. Nevertheless, some important factors (e.g., the fate oforganic matter, toxic substances, economic and social conditions)were not considered (Montangero et al., 2006a, 2006b). Moreover,Cencic and Rechberger (2008) introduced a user-friendly softwarenamed STAN which supports performing MFA according to theAustrian standard ÖNORM S 2096 under consideration of datauncertainties. Predefined elements such as processes, flows, systemboundaries and text fields can be imported from Microsoft Excel orinput manually whereas uncertain data are assumed to be normallydistributed. With these input data, the graphical MFA model can beautomatically translated into a mathematically model using thefollowing equations:

Balance equation : ∑inputs ¼ ∑outputsþ change in stock ð4Þ

Transfer coefficient equation : outputx¼ transf er coef f icientto output x �∑inputs ð5Þ

Stock equation : Stockperiodiþ1 ¼ Stockperiodi þ change in Stockperiodi ð6Þ

Concentration equation : masssubs tance¼ massgood � concentrationsubs tance: ð7Þ

Finally, STAN expresses the mass flows as Sankey arrows whichare proportional to their mass flow values. In addition to performingdynamic MFA, STAN is also capable of evaluating the economic,resource and environmental value of the materials.

2.1.3. Characteristics and weaknesses of MFA modelsWith respect to the scope, MFA models are not only restricted to

flows within the region but also trace the flows beyond the boundaryas far as they are relevant to regional activities thereby enabling thedetection of unexpected side-effects within the region to otherregions or other time periods. The dynamic approaches are morecomplicated than conventional methods as the temporal changes,transfer coefficient and economic conditions are involved as well.With quantitative MFA models becoming more accurate andadvanced, they are likely to give more realistic pictures of theregional environmental status (e.g., nutrient flow, recycled waterconsumption and wastewater discharge). Yet the manipulation ofsophisticated MFA models still provides a big challenge to decision

makers. Despite of these strengths, MFA can only deal with onesubstance and the related environmental interventions at a particulartime in one area. The side-effects to other substance chains areoutside the study scope (Brunner and Rechberger, 2004; Hendrikset al., 2000; Wrisberg and Udo de Haes, 2002). As such, there is a riskthat a critical problem might be overlooked if a wrong judgement ismade on the goods/substance selection (Tangsubkul et al., 2005a).While most of recent MFA studies have recognised the importance ofwater recycling and reuse in environmental sanitation management,the downstream assessments and discussions on the feasibility andsuitability of particular recycled water schemes are still essential.More work on the different fractions of water sources arrived at theWWTP, subsequent treatment technologies, the effluent quality,possible end uses and potential risks to human health and theenvironment should be done in the following analyses.

2.2. Life Cycle Assessment (LCA)

When dealing with water reuse issues, LCA mainly focuses on theenergy and material requirements throughout an entire life cycle ofthe treatment process as well as the quality of treated effluentassociated with fit-for-purpose end uses (Muñoz et al., 2009a).Table 1 lists several environmental impact categories or indicators inLCA, the analytical results of which are able to give an overall pictureof the system performance or contributions to decision makers(Ahmed, 2007; Hermann et al., 2007; Liamsanguan and Gheewala,2007).

2.2.1. Types of LCA modelsThere are generally three types of LCA models at present, which

are process-based LCA, economic input–output (EIO) LCA and hybridLCA. The initial and simplest approach is the process-based LCAwhichis usually carried out in four steps: (1) goal and scope definitions,(2) life cycle inventory analysis, (3) life cycle impact assessment(LCIA) and (4) life cycle improvement analysis and interpretation(Stokes and Horvath, 2006; Tangsubkul et al., 2005b). These foursteps are inter-connected and can quickly contribute to strategicplanning, technology development and improvement as well asdecision making and marketing (ISO, 14040, International Standard,1997; Pennington et al., 2004). Table 2 describes the purposes andmajor contents of each step. Still, there are several issues that need tobe addressed. First, in Step 1, the inputs (e.g., materials and energysources) and outputs (e.g., emissions and wastes) should be clearlyidentified and insignificant contributions in the system boundaryshould be excluded or ignored so as to minimise the time and effort ininformation collection in Step 2 (Matthews, 2011).

Table 2Purposes and major contents in each step of a LCA study on water reuse.a

Steps Purposes and major contents

Goal ● To evaluate the environmental impacts of the wastewater treatment technologies and their combinationsFunctional unit ● To deliver a certain amount of recycled water produced at the WWTP for reuse applications, which enables comparison between different

scenarios delivering the same amount of the waterSystem boundary Phases that are considered: WWTP construction phase (Optional); WWTP operation phase and WWTP demolishment phase (Optional). Others

need to be considered: materials, chemical additives, energy consumption, transport, etc.Life cycle inventoryanalysis (LCI)

To quantify the environmental relevant inputs and outputs of all processes over the life cycle. Data can be collected from internal reports, personalinterviews with WWTP staff or previous literature. The LCI databases have already been equipped in some LCA softwares (e.g., Gabi and Simapro).

Life cycle impactassessment (LCIA)

● To interpret the inventory results into their potential environmental impacts, such as GWP, EP, HTP, etc.● To integrate various types of impacts and express them as a single indicator

Life cycle assessmentand interpretation (LCAI)

● Contribution analysis (to examine which process, substance and impact category is most important based on the LCI and LCIA)● Consistency check (to check whether the assumptions, methods, models and data are consistent with the goal and scope of the study)● Completeness check (to check whether all processes and data are complete)● Sensitivity/uncertainty analysis● Conclusions and recommendations

a Modified from Andrae (2010), Meneses et al. (2010), Sharaai et al. (2009), Suh et al. (2004), and Vlasopoulos et al. (2006).


As for Step 3, several LCIA methods (Table 3) are normally used toquantify the different environmental indicators. Dreyer et al. (2003)demonstrated that EDIP 97 and CML 2001 are both midpointapproaches, which showed only minor differences for most impactcategories except for the ones that describe toxicity to humans and theecosystems. The results of the Eco-indicator 99 and EDIP 97 reachedopposite directions for some inventories as the former one is an endpointmethod,where the patterns ofmost important contributors to theweighted and aggregated impact scores are quite different from EDIP97. Both Dreyer et al. (2003) and another study by Renou et al. (2008)concluded that more work was required on human and ecosystem

Table 3LCIA models and associated characteristics.

LCIA methods Characteristics

CML 2000 ● It is a mid-point approach● This method consists of characterisation, normalisation and evaluation● It focuses on two main classes of the environmental impact: the exhau

of raw materials and energy (abiotic and biotic resources) and pollutio(focusing on GWP, ODP, ETP, HTP, smog, AP, EP, solids).

● Grouping and weighting were not included in this method.Eco-points 97 ● It is based on the critical targets derived from the Swiss policy and calc

the impact assessment with a single score.● It allows a comparative weighting and aggregation of various environm

interventions by use of so-called eco-factors and evaluates emissions inwater and top-soil/groundwater as well as energy sources

● One can choose different normalisation factor (normalisation based onvalue or actual emission), so that different equations can be applied.

Eco-indicator 99 ● It is an end-point approach, and weighting wassimplified using three points.

● Nitrogen oxides are the main contributors for thismethod.

EDIP 97 ● It is a thoroughly documented mid-point approach.

● The normalisation is based on person equivalents whereas the weightibased on political reduction targets for environmental impacts.

● Ecotoxicity and human toxicity are modelled using a simple key-propeapproach where the most important fate characteristics are included insimple modular framework.

USES-LCA ● It is an integrated multimedia fate, exposure and effects model● It has been developed for use in risk assessment of non-polar organi

chemicals as well as other substances such as metals● It contains a database of 3396 chemicals

● It calculates characterisation factors for ETP and HTP on both themidpoint and endpoint level

● Due to the large uncertainty of modelling metal behaviour in theenvironment, it involves testing the sensitivity of the metalcharacterisation factors

toxicity indicators as LCIA methods did not converge toward similarresults. Currently, by applying commercial LCA softwares such as Gabiand SimaPro, one can easily choose one or several different LCIAapproaches (e.g., CML 2000, Eco-points 97, Eco-indicator 99, EDIP 97,LIME and USES-LCA) for assessment. Comparatively, EIO-LCA is aneconomic-based technique which is to capture all economic trans-actions, resource requirements and transportations to produce recycledwater of certain quality and then calculate the associated environmen-tal emissions and wastes (e.g., energy use, toxic air emissions andhazardous waste) in terms of economic expenditures. Furthermore,to overcome the disadvantages of conventional process-based and

Advantages and disadvantages References

Since the different substances can have considerablevariation in the characterisation step, the resultingeffect scores may not be completely reliable.

Gawor (2009)steps.stionn

ulates ● It is detailed and substance-specific Gawor (2009)

entalto air,

● Only a few impacts are assessed

target ● The end-point system is normally based on policylevels instead of sustainability levels, which areusually a compromise between political andenvironmental considerations.Despite the simplicity, weighting was the keyproblem.

Goedkoop et al. (2008)

● It covers most of the emission-related impacts,resource use and working environment.

Dreyer et al. (2003)

ng is ● The updated version-EDIP 2003 covers a largerpart of the impacts and lies closer to adamage-oriented approach.

rtya

● It is an easy-to-use model. Dreyer et al. (2003)and van Zelm et al.(2009)

c ● Both midpoint and endpoint characterizationfactors are calculated.

● Apart from FAETP, MAETP and TEP are alsoaddressed.

● Various scenarios can be tested by changingscenario settings.

Table 4Electricity production models and associated compositions.a

Electricity production models Thermal Nuclear Hydroelectric

European model 43.3% 40.3% 16.4%French model 11.4% 72.9% 15.7%Norwegian model 0.5% 0.3% 99.2%Portuguese model 80.8% 2.6% 16.6%

a Adapted from Ortiz et al. (2007).


EIO-LCA models, hybrid LCA has been developed over the years, whichcombines the accuracy of process-based LCA and the completeness ofEIO-LCA (Mattila et al., 2010).

2.2.2. Application of LCA models on recycled water schemesPresently, some studies have applied these three types of LCA

models in comparing different water resources in terms of energyconsumption and environmental concerns. Pasqualino et al. (2010)pointed out that replacing both potable and desalinated water byrecycled water for non-potable purposes (e.g., irrigation, industry,urban cleaning, fire fighting, recreation and groundwater recharge)could result in lower environmental impacts for all selected impactcategories, including AP, GWP, EP, PHO, DAR, ODP and ETP. Althoughthe environmental impacts caused by tertiary treatment were higherthan that of secondary treatment, they can be compensated for bysignificant freshwater and energy savings. In a similar study byMeneses et al. (2010), apart from the above-mentioned strengths, theenvironmental benefits of recycled water from reduced fertiliserrequirements in agriculture were also demonstrated.

Likewise, Stokes and Horvath (2006) compared recycled waterwith imported freshwater and desalinated water using a hybrid LCAmodel. The model was implemented in an Excel-based decisionsupport tool named Water-Energy Sustainability Tool (WEST). InWEST, to minimise time and data requirements, EIO-LCAwas adoptedto determine the effects related to material acquisition, transforma-tion and production whereas process-based LCA was used to assessthe environmental effects of treatment system construction andoperation. Several air emission indicators including GWP, nitrogenoxides, sulphur oxides, particulatematter, volatile organic compoundsand carbon monoxide were considered in the analysis. The studydemonstrated that the environmental impacts were mainly dominat-ed by energy consumption in the system operation. According to thecases studies on two California water utilities, desalination had 2–5times larger energy demand and caused 2–18 times more emissionsthan freshwater importation and water recycling while freshwaterimportation raised concerns about reliability and environmentprotection. Overall, the recycled water for irrigation and commercialcar washing was found to be more environmentally sustainable. Themost distinct advantages of WEST were that it successfully incorpo-rated EIO-LCA with process-based LCA and took into account theconstruction phase in addition to the operation phase. Stokes andHorvath (2009) further modified the previous study and verified thestrengths of recycled water for non-potable uses (e.g., irrigation,commercial and industrial applications) over imported water, desa-linated seawater and desalinated brackish groundwater in terms ofenergy consumption, air emission concerns, reliability, etc. Neverthe-less, recycled water quality issues were not addressed in these twopapers.

Having recognised and confirmed the advantages of recycledwater, other studies have also employed LCA models to select optimalwastewater treatment technologies or stages for recycled waterplanning in agriculture, industry and urban landscape irrigation orevaluate the environmental profiles of existing WWTP (Gallego et al.,2008; Lim and Park, 2009; Romero-Hernandez, 2005; Vlasopouloset al., 2006). Yet the full application of LCA in holistic recycled waterscheme assessment is still quite limited.

2.2.2.1. Agricultural uses. Agricultural irrigation represents the largestuse of recycled water throughout the world. LCA models, in this case,are proved to be helpful in initial water recycling planning and design,especially in some of the large-scale irrigation schemes in Australia,Europe and the Middle East. For instance, Ortiz et al. (2007) used aprocessed-based LCA in comparing four wastewater treatment sce-narios, including Conventional Activated Sludge System (CAS), CAS-UF,external membrane bioreactor (MBR) and immersed MBR. The con-struction, operation and dismantlement phases of the plant were

considered in the system boundary and the airborne emissions were ofprime concern. The results indicated an overall lower impact in CAS,followed by immersed MBR and external MBR. Additionally, whencombining the treatment technologies with electricity productionmodels (Table 4), CAS-Norwegian model demonstrated the lowestimpact as renewable energies aremuchmore environmental friendly inelectricity production comparedwith fossil fuels and nuclear resources.On the other hand, the systems expect for CAS produced high effluentquality, which not only allowed the water to be safely reused inirrigation but also enabled other applications such as groundwaterrecharge, household and industrial uses. Thus, considering both envi-ronmental impact and water quality, immersed MBR coupled with therenewable energy consumption pattern was optimal. Nevertheless, thisstudy did not take into account the environmental impacts on soil andwater nor consider the toxic and health effects.

To investigate the toxicity-related impacts of recycled water onagriculture, Muñoz et al. (2009a) evaluated four water treatment andreuse scenarios, including no reuse, reuse without tertiary treatment,reuse after tertiary (ozonation) treatment and reuse after tertiary(ozonation and hydrogen peroxide) treatment. Two LCIA approaches(USES-LCA and EDIP 97) have been applied for evaluation. Both ofthem showed that water reuse scenarios with tertiary treatment werepreferred and the ozonation approach arrived at lower toxicity scorecompared with the ozone-peroxide one. The study disregarded the EPpotential and microbiological quality of recycled water. Besides, theenvironmental results with high uncertainty and significant deviation(1.5 to 6 orders of magnitude) were also observed in USES-LCA. Yet,Meneses et al. (2010) found that chlorination–UV disinfection wassuperior over ozonation and ozonation-hydrogen peroxide systemsin terms of AP, GWP, EP, PHO, DAR, ODP, FWU and CED, whenconsidering tertiary treated water for agricultural and urban applica-tions. They also indicated that winter climate conditions contributedto lower environmental impacts due to the changes in populationhabits, water quality and water use.

Apart from processed-based LCA models, Tangsubkul et al.(2005b) initially used EIO-LCA to evaluate three treatment methods,including the CMF (Ozonation-MF-disinfection), MBR (MBR-RO) andWastewater Stabilisation Pond (WSP) systems for irrigation purposesat the Rouse Hill residential area in Sydney, Australia (Fig. 4). Withthe assistance of Missing Inventory Estimation Tool (MIET) 2.0 andGaBi3 softwares, all flows associated with the construction activitiesof treatment options were converted into a monetary value perfunctional unit. The environmental impact categories in considerationwere GWP, EP, HTP, FAETP, MAETP, TEP and salinisation potential(SP). The results revealed that MBR system could result in a low SPvalue but was likely to trigger soil dispersion. The WSP system wasregarded as the most suitable option for irrigation applications ifreductions to SP impact were made. Although MIET was proved to bean acceptable means of estimating the impacts caused by theconstruction phase of wastewater treatment, a major constraint isthat since the MIET relies on the 1996 US Input–output table, theresults might not be appropriate to other countries due to inter-industry cost structure differences.

2.2.2.2. Industrial uses. Recycled water has been successfully applied toindustry in Japan, the US, Canada and Germany since the Second

Material production

Material inputs

Chemical production

Chemical dosing

Raw wastewater

Construction

Treatment train

Recycled water

Agricultural applications, such as bean, almond,

apricot, plum and grape

Sludge treatment

Waste

Biosolidsfinal use

Landfill

System boundary

Input Output

Fig. 4. System boundary of LCA study on Rouse Hill water recycling scheme.Modified from Tangsubkul et al. (2005b).


World War for more than 70 years. Recently, industrial use is thethird biggest contributor to recycled water consumption. The majorindustrial categories associated with substantial water consumptioninclude cooling water, boiler feed water and industrial process water.In regard to cooling and boiler feed purposes, Vlasopoulos et al.(2006) investigated over 600 different treatment technology combi-nations on treating petroleum wastewaters for industrial andagricultural reuses. The study took into account the environmentalimpacts of different treatment options in terms of GWP, DAR, AP, EPand PHO during the construction and operation phase, using CML2000 method. The optimal treatment technology combinationsassociated with four end uses are shown in Table 5. As can be seen,compared with two agricultural end uses (categories 1 and 2) wherethe crop products can be eaten raw, cooling and boiler feed waterrequired lower water quality (categories 3 and 4). Specially, thecooling system allowed the use of three-staged treatment with moretechnology combinations due to lower water quality requirement forsodium. However, to select the optimal treatment technology for eachtreatment stage, the study highlighted that one should not onlyconsider the involved environmental impacts but also the recycledwater quality as well as indirect downstream benefits to subsequenttreatment stages.

Jørgensen et al. (2004) studied six alternatives for water recyclingand handling of associated residues at an industrial laundry companyin Denmark. It was concluded that onsite wastewater reuse scenariosusing UF or a bio-filter led to lower environmental impacts comparedwith current no reuse practice. As laundry process wastewater carriedsome pollutants (e.g., heavy metals), UF coupled with sludgevitrification was proved to be the optimal technology combination

Table 5Most environmentally friendly treatment technologies associated with different end use ca

End uses Water quality requirement Best technologcombinations(impact catego

(1) Barley-wheat 4-staged treatment (53 technologycombinations can meet the requirement)

MF–ORG–RO (DAF–CWL–DM(except EP)

(2) Citrus 4-staged treatment (53 technologycombinations can meet the requirement)

MF–ORG–RO–I(in EP)DAF–CWL–DM(except EP)

(3) Alfalfa–sorghum–

cotton–rhodes-boiler feed4-staged treatment (104 technologycombinations can meet the requirement)

DAF–CWL–DMMF–ORF–RO

(4) Cooling system feed 3-staged treatment (139 technologycombinations can meet the requirement)

DAF–ABS–ORG

Abbreviations: DAF = Dissolved Air Flotation; CWL = Constructed Wetlands; ABS = AbsorReverse Osmosis; and ION = Ion Exchange.

a Modified from Vlasopoulos et al. (2006).

in terms of lowest toxicity impact to the environment. The UFpermeate could be safely reused in the washing process. The resultsdemonstrated that LCA was able to identify the best treatmenttechnologies as well as long-term environmental benefits of waterreuse in laundries. Moreover, Zhang et al. (2010) have adopted ahybrid model to measure the life cycle benefit of treated water reusein industrial and domestic applications as well as the correspondinglife cycle energy consumption in the construction, operation anddemolishment phases of theWWTP in Xi'an, China. When quantifyingthe environmental impacts of different treatment stages as equivalentenergy consumption, the study employed the Eco-indicator 99method. However, unlike processed-based LCA, the system boundaryin this model was confined to the secondary and tertiary treatmentunits without any sub-boundary between them, which is capable ofreducing the possible difficulties in conventional approaches. Thestudy indicated that energy consumption in operating the tertiarytreatment (2065.28×109 kJ) can be significantly compensated by lifecycle benefit of water reuse in terms of reduced wastewater discharge(74.2×109 kJ) and freshwater saving (1598.4×109 kJ). Although thisstudy successfully linked the life cycle energy consumption withdirect benefits of recycled water reuse (e.g., wastewater reductionand freshwater saving), other indirect benefits such as ecosystemprotection and water cycle improvement were not considered.Besides, several important impact categories such as GWP, SP andETP were not evaluated in the model.

2.2.3. Characteristics and weaknesses of LCA modelsAmong three types of LCA models, process-based LCA is still the

most widely used approach which allows for specific treatment

tegories.a

y

ries)

Descriptions and comments

in EP) Although CWL reached higher environmental impact than ABS atsecondary stage, it resulted in smaller design and lower energyconsumption in subsequent treatments due to evaporation losses.

F–RO

ON ION was to achieve additional boron removal by 0.5 mg/L. ION onlycontributed to 1% of the overall environmental impact.

F–RO–ION

F–RO and Although this end use required less stringent water quality, quaternarytreatment was still needed to reduce the sodium concentration.Although SSF and DMF had better environmental performance thanORG in the treatment stage 3, they did not satisfy the required coolingwater quality.

bents; DMF = Dual Media Filtration; MF = Microfiltration; ORG = Organoclay; RO =


technology comparisons and selections for new equipment instal-lations. As the international standards (ISO 14040 and 14044)regarding principles, frameworks and requirements and guidelinesfor process-based approach have already been clearly documentedand the current commercial softwares (e.g., Gabi and SimaPro) tendto be more mature and robust in performing LCI and LCIA analyses,this method could be easily manipulated and give reliable andconvincing results to decision makers. Besides, process-based LCAcould also provide suggestions for future improvement and develop-ment since it is able to identify technology weak points in evaluationprocesses.

Yet one of themost serious difficulties is the choice of an appropriatesystem boundary. Some researches excluded the insignificant contri-butions within the system as considerable materials or processes caneasily lead to an overwhelming number of inputs and outputs whereasothers pointed out that such a narrow focus might ignore importanteffects and cause incorrect decisions (Hendrickson et al., 2011;Matthews, 2011). Moreover, this approach is time intensive and costly,especially in more complicated assessment. When a large variety ofimpact categories are required to be considered, much time will notonly be consumed on emission classification and characterization, butalso spent on normalisation and weighting processes to make theindicators dimensionless thereby enabling comparison and achievingan overall score. As the aggregation result is sensitive to the set ofweights, the lack of detailed site data and decision makers' lowconfidence on optimal weight assignment will reduce the reliabilityand robustness of the final score due to involved uncertainties. The highprices of commercial softwares also limit the model popularisation tosome extent. Additionally, it would be difficult to apply process-basedLCA to newly developed treatment technologies when the relevantmaterial and energy consumption are not available in LCI database.

Comparatively, EIO-LCA approach overcomes the system boundaryproblem. Although the boundary is very broad and inclusive, thetransactions and emissions of all processes among all phases areincluded. It is also relatively quick to be performed and has modestdata requirements. This approach is ideal for comprehensive assess-ments and systems-level comparisons since all environmental flowshave been converted into monetary values with an assumption of aproportional relationship between them thereby avoiding the tediousnormalisation step. Unlike site-specific results from process-basedLCA which sometimes involve a certain degree of confidentiality, theEIO-LCA results could be reproducible and publicly available as theyare economy-wide. However, as EIO-LCAmust link physical units withmonetary values, it mainly captures environmental impacts associat-ed with raw material acquisition and construction stages of recyclingfacilities rather than downstream impacts (e.g., water reuse, wasterecycling and end-of-life options). As a significant amount of inputsare needed, lack of data sometimes hinders the complete assessmentof environmental effects. It is also difficult to be applied to an openeconomy with substantial non-comparable imports. Besides, theuncertainties are likely to be high through indicator aggregation,monetary transactions between currencies and times and the possibleuse of outdated data on interactions and emissions (Hendricksonet al., 2006; McMichael, 2011; Stokes and Horvath, 2006).

Hybrid LCA is regarded to be a state-of-the-art method, whichinvolves the integration of more reliable bottom-up process-based LCAdata into the comprehensive top-down EIO-LCAmethods (Mattila et al.,2010). Although the hybrid approach can take advantage of bothmethods, themodel tends to be complicated and hard to be understoodor manipulated by decision makers in a short time. Another significantissue in hybrid LCA is to avoid double counting. Simply adding theresults of a process-based LCA and an EIO-LCA of the same system willerroneously cause the system components modelled twice. Further-more, some practical limitations of this method (e.g., model structurevariation, methodology uncertainties, data completeness, softwaredeficiency, etc.) also exist (Rowley et al., 2009).

2.3. Environmental Risk Assessment (ERA)

When considering the introduction of recycled water to local areasor the expansion of current recycling schemes, most of riskassessment studies have been devoted to human health riskassessment (HRA) with implications from pathogens (Donald et al.,2009; Hamilton et al., 2007;Westrell et al., 2004) or chemicals (Cao etal., 2010; Rodriguez et al., 2007; Weber et al., 2006). Various existingnational recycled water guidelines also mainly focus on the potentialrisks from pathogens and there is little mention of other tracepollutants. Nevertheless, from an environmental perspective, chem-ical pollutants in recycled water (e.g., excessive pharmaceuticals andxenobiotic compounds) to receiving ecosystems could be the firsttargets. Even if the scheme (e.g., agricultural and landscape irrigation,environmental and recreational impoundment flows and groundwa-ter recharge) is conducted far away from human targets or activityzones, it can induce environmental burdens unintentionally when therecycled water quality is unacceptable (Corwin and Bradford, 2008;Fatta-Kassinos et al., 2011b; Tiruta-Barna et al., 2007). The potentialenvironmental risks resulted from water reuse projects include(Urkiaga et al., 2008):

• Substantial alteration of land use;• Conflict with the land use plans or policies regulations;• Adverse impact on wetlands;• Affection of endangered species or their habitat;• Populations displacement or alteration of existing residential areas;• Antagonistic effects on a flood-plain or important farmlands;• Adverse effect on parklands, reserves, or other public landsdesignated to be of scenic, recreational, archaeological, or historicalvalue;

• Significant contradictory impact upon ambient air quality, noiselevels, surface or groundwater quality or quantity;

• Substantial adverse impacts on water supply, fish, shellfish, wildlife,and their actual habitats.

Due to the above concerns, ERA evaluates the ecological riskimpacts of environmental changes or multiple stressors in relevantsystem boundary over long periods. It integrates ecology, environ-mental chemistry, environmental toxicology, geochemistry and otherfundamental sciences in characterising the impacts of natural andman-made disturbances on ecological resources (Bartell, 2008).Particularly, for agricultural and landscape irrigation uses, concernsshould be given to inorganic species in recycled water (e.g., sodium,potassium, calcium, chloride and bromide, and trace heavy metals).Owing to high solubility, they are too difficult to be treated bymembrane processes, not to mention advanced oxidation processesand traditional biological treatment. For this reason, their potentialrisks to the environment have been reported as highly salineirrigation water can severely degrade the soil and the accumulationof heavy metals in soil can pose threats to food chain. Situations maybe even worse in dry climates since the concentrations of salts indrainage can be much higher due to evaporation, posing potentialthreats to groundwater quality.

Apart from inorganic chemicals, the trace organic pollutants (e.g.,pharmaceutical active compounds, endocrine disrupting compounds,antibiotics and xenobiotics) are also important substances as indus-trial and technological advances in production of chemicals haveoutpaced the regulatory practices (Bolong et al., 2009; Meisel et al.,2009; Fatta-Kassinos et al., 2011a, 2011b). Due to high hydrophilicityand low adsorption ability, they are poorly removed by conventionalactivated sludge and are likely to cause adverse biological effects inorganisms at part per trillion concentrations (Weber et al., 2006). Forinstance, Gagne et al. (2006) demonstrated that redox properties ofsome pharmaceuticals could influence the hepatocyte oxidativemetabolism of rainbow trouts as well as the plant growth and plant–


microorganism symbiosis. Besides, these organic compounds cantransport via recycled water and exhibit a significant cumulative effecton the ecosystem. Karnjanapiboonwong et al. (2011) found thatpharmaceuticals in recycled water during land application transportedboth vertically and horizontally in the soil and eventually reachedgroundwater at West Texas, the US. In another study by Drewes et al.(2002), antiepileptic drugs (e.g., carbamazepine and primidone)showed persistent during travel times of more than 6 years in thesubsurface where secondary and tertiary treated effluents wererecharged for indirect potable reuse purposes. Al-Rifai et al. (2007)also reported the high concentrations of pharmaceutical compoundsthrough partitioning to soil and possibly plants over long periods at theGerringong Gerroa Sewage Scheme in southeast Australia. Fig. 5 givesgeneral steps in performing ERA where the environmental risk ofparticular compound should be carefully identified and characterisedunless sufficiently low concentrations are observed. Once the re-quirements from risk managers and decision makers are fulfilled anddocumented, the processes end (Bartell, 2008; Carlsson et al., 2006).

2.3.1. Types of ERA modelsDespite of difficulties, numerous ERA models have been increas-

ingly developed, which are proved to be quite useful when empiricalmeasurements of toxic effects are unavailable, measured values arescarce or the exposure level is being projected into the future (Leeet al., 2007). The simplest ERA approaches normally apply chemicalanalytical instruments such as atomic adsorption spectrophotometer(AAS), inductively coupled plasma optical emission spectroscopy(ICP), liquid chromatography (LC), gas chromatography (GC) ormass spectrometry (MS) to determine the predicted environmentalconcentration (PEC) of the pollutant and then compare it with thepredicted no effect concentration (PNEC) guideline value so as toarrive at the potential risk. Comparatively, instead of using costlydetection instruments, some ERA models estimated the PEC usingmathematical equations, which are based on initial pollutant concen-tration, percentage removal rate, dilution factor and the volume ofrecycled water. Others more complicated approaches even take intoaccount biodegradation effect of the substance during environmentalexposure. On the other hand, some studies also adopted empiricalmodels in PNEC calculation when relevant data were not available. Forexample, the ecological structure activity relationship (ECOSAR)model has been routinely used by US EPA, which is able to calculatethe unknown aquatic toxicities of organic pollutant from the structureactivity relationships (SAR) of other similar compounds with known

Description of the potential sources (e.g. greywater, municipal and industrial wastewaters)

Quantitaassessme

Identification of the undesired ecological impacts in relation to the sources (e.g. adverse

impacts on wetlands, endangered species and residential areas)

Conduct of risk assessment (conceptual, qualitative and quantitative approaches)

Collaboration among risk assessors, risk managers, stakeholders, concerned members

of the public and other organisations

Establishment of risk control and risk management approaches

Fig. 5. Steps in perReferred from Bar

environmental toxicities. The SARs are based on test data and expressthe correlations between a compound's physical and/or chemicalproperties and its aquatic toxicity. Thus, toxicity values of thepollutant under study can be calculated by inserting the estimatedoctanol water coefficients (Kow) into the regression equation andcorrecting the resultant value for molecular weight of the compound(Jones et al., 2002).

2.3.2. Application of ERA models on recycled water schemes

2.3.2.1. Agricultural uses. ERA models are widely used to evaluate thepotential effects of waterborne hazards on soil and surroundinggroundwater quality for several agricultural recycled water schemes.Regarding inorganic chemicals, Xu et al. (2010) observed the longterm (3, 8 and 20 years) recycled water irrigation (56.78 ML/d,processed by primary sedimentation and oxidation pond) on agricul-tural soils for plots growing trees and feed crops at Palmdale,California, the US. Despite the benefits of nutrients and fresh watersaving, the ICP analytical results showed that trace metals include Cr,Cu, Ni and Zn in the upper horizons were accumulated over years,which eventually deteriorated the soil and groundwater quality. Asimilar study by Li et al. (2009) using AAS indicated that althoughirrigating with poorly treated industrial wastewater at Zhangshiirrigation area in Shenyang, China has been ceased since 1992, the Zn,Pb and Cu concentrations were still higher than or close to China'sgrade A standard due to 30 years' accumulation effects. Thus, the placeshould be abandoned for cultivated crops and bioremediation or othermeasures should be carried out. Likewise, Katz et al. (2009) reportedthe elevation of nitrate, boron and chloride concentrations ingroundwater samples from the Sprayfield aquifer, where secondarytreated municipal wastewater was supplied for agricultural irrigation.Peasey et al. (2000) and Jimenez and Asano (2008) also found acorrelation of risk problems with the proximity to farms whererecycled water had been applied.

Furthermore, Muñoz et al. (2009b) used a more complicated ERAmodel in risk characterization of heavy metals as well as pharma-ceuticals in recycled water for agricultural irrigation. An ERAframework for screening environmental risks in terrestrial ecosystemis depicted in Fig. 6. Both the risk exposure to soil organisms (PECsoil)and second poisoning to top predators via terrestrial food chain(PECpredator) were quantified by the European Commission TechnicalGuidance Document on Risk Assessment and the level III fugacitymodel while PNEC values were derived from the ECOSAR software.

Exposure assessment, in which the predicted environmental concentrations (PEC) of the pollutants of concern are

determined for different compartments (e.g. soil and freshwater)

Effects assessment, where a dose-response relationship is established, involving the determination of a predicted no-effect

concentration (PNEC)

Risk characterisation, which involves calculating the risk quotients (RQ). RQ = PEC/PNEC, where a value above 1 means

that adverse effects are likely to occur

tive nt

forming ERA.tell (2008) and Muñoz et al. (2009b).

Definition of target

substances

Wastewater effluent

sampling

Chemical analysis

Substance-independent data:-Environmental parameters

-Exposure parameters

Substance-dependent data:-Physical-chemical properties

-Toxicological properties

Effluent concentrations

Experimental work

Data collection

Soil

DegradationVolatilisation

Bio-concentrationLeaching

Irrigation PECsoil

PECworm

RCRsoil

RCRpredatorEffects assessmentPNECsoil

PNECoral

Exposure assessment (modelling)

Fig. 6. Proposed framework for screening environmental risks related to wastewater reuse.Adapted from Muñoz et al. (2009b).


The case studies on 27 pollutants in secondary treated effluents fromtwo Spanish WWTPs showed that both plants were likely to causeadverse effects on agriculture soil and predators. The nickel concentra-tions in recycled water from the Alcala de Henares WWTP whichreceives a mixture of domestic and industrial wastewater were toxic tosoil and predators whereas pharmaceuticals (sulfamethoxazole, cipro-floxacin, diclofenac, gemfibrozil and erythromycin) in both effluentsposed high risks on soil compared with diclofenac in effluents onpredators. Thus, additional treatments with membrane filtration,advanced oxidation and UV disinfection were recommended.

With respect to pharmaceuticals, the experimental and predictedsludge–water partition coefficients have been observed to bedifferent by several orders of magnitude (Jones et al., 2002; Stuer-Lauridsen et al., 2000). Hence, Lindberg et al. (2007) still used theinstrumental methods (ultrasonic-assisted liquid/solid extraction andLC–MS/MS) to investigate the concentrations of norfloxacin andciprofloxacin in dewatered wastewater sludge and pellets. Theleaching tests demonstrated that these compounds had limitedmobility as the accepting aqueous phases contained less than 1% ofthe pharmaceutical concentrations that were initially found in solidphases. Consequently, they would not affect groundwater quality ifthe wastewater sludge is reused as fertiliser in agricultural soil.

2.3.2.2. Environmental and recreational uses. To release reliableenvironmental and recreational flows and protect the downstreamhealth of rivers, many ERA studies have been conducted with a focuson pharmaceuticals in recycled water. For instance, Stuer-Lauridsenet al. (2000) investigated 25 pharmaceuticals in Denmark underworst case scenarios where all sold pharmaceuticals were assumed tobe used evenly at temporal and spatial scale in the same year andthen released to the sewage system without any attenuation.Accordingly, PECs were calculated as follows:

PECw ¼ A� 100−Rð Þ365� P� V� D� 100

ð8Þ

where A is the amount used per year (kg/yr), R is the removal inpercent (set to zero), P is the number of inhabitants in Denmark(5,200,000 in 1997), V is the volume of wastewater per day per capita(0.2 m3) and D is the dilution factor in the environment (a defaultvalue of 10 is used). The calculated PECs were generally consistentwith actual measured pharmaceutical concentrations. On the otherhand, as ecotoxicity data were only available for 6 compounds, thecorresponding PNECs were derived on the basis of Europe draftguideline document for medicinal products for human use with a

default safety factor of 1000. Finally, high risks were observed onacetylsalicylic acid, paracetamol and ibuprofen while low risks wereachieved for oestrogen, diazepam and digoxin. Although the math-ematical PECmodel is easy to perform, lack of chronic toxicity data forPNECs is the prime obstacle in the study. To partially solve thisproblem, Jones et al. (2002) used the ECOSAR model in PNECestimation when assessing the top 25 English prescription pharma-ceuticals to the aquatic environment. Four types of pharmaceuticalsincluding mefenamic acid, oxytetracycline, paracetamol and amoxi-cillin were shown to be of high risk. Nonetheless, risk quotients (RQs)of the pharmaceuticals might be overestimated in both studies sincethe worst case scenarios were adopted in calculating PECs as well asacute ecotoxicity data in quantifying PNEC.

Additionally, Carlsson et al. (2006) have made some improve-ments in estimating PECs of 27 pharmaceuticals in Sweden. Morespecifically, PECswere firstly calculated underworst case assumptionsusing Eq. (9) and then refined by a Simple Treat 3.1 model to reflectlocal realistic environmental conditions. To trace the steady-statepharmaceutical concentrations in recycled water, sludge or air, themodel requires the input of several physical–chemical parameters(e.g., molecular weight, Kow, vapour pressure, water solubility,dilution factor, degradation rates and acid–base dissociation con-stants). Owing to limited biodegradation data, the degradation ratewas assumed to be 0.1/h for paracetamol but zero for others. In regardto PNECs, when toxicities of the chemicals were known, they could bederived from the lowest available acute values to organisms (e.g., LC50of fish, EC50 of daphnia and IC50 of algae) divided by an assessmentfactor of 1000. Otherwise, the ECOSAR model should be applied.Overall, nine substances were considered to be dangerous while onlyoestradiol and ethinyloestradiol were likely to cause long termadverse effects to the aquatic compartment. Despite the considerationof local conditions in PEC estimation, most RQ values were over-estimated as chronic toxicity datawere only available for 4 substances.

PECsurface water ¼DOSEai� Fpen

WASTEWinhab� DILUTION� 100ð9Þ

where DOSEai is the highest recommended daily dose of thepharmaceuticals in question, and Fpen is the percentage of marketpenetration. WASTEWinhab is the amount of wastewater used perinhabitant per day and DILUTION is the dilution of sewage water insurface water.

Furthermore, Escher et al. (2011) proposed another ERA approachfor pharmaceuticals from hospital wastewater which might be


directly discharged to surface waters or infiltrated. The detailed PECcalculations under four scenarios were shown in Table 6. PNECs wereestimated from acute toxicity (EC50) values divided by 1000. The EC50

values were derived from Eqs. (10)–(12), where Dlipw values refer tothe liposome-water distribution coefficients at pH 7 and are availablein literature.

For green algae : log 1=EC50 Mð Þð Þ¼ 0:95� logDlipw pH 7ð Þ þ 1:53 ð10Þ

For water flea : log 1=EC50 Mð Þð Þ ¼ 0:90� logDlipw pH 7ð Þ þ 1:61 ð11Þ

For fish : log 1=EC50 Mð Þð Þ ¼ 0:81� logDlipw pH 7ð Þ þ 1:65: ð12Þ

The RQs of Top-100 pharmaceuticals in Switzerland from thegeneral hospital and psychiatric centre were then calculated under allscenarios. The results demonstrated that for chemicals withRQHWW>1, the dilution effect significantly decreased the RQ values(RQWWTP influentb1) and even had better performance than the actualelimination in WWTP. However, it is not the real fact as clotrimazoleand ritonavir were found to be highly removed (>80%) in biologicaltreatment. Due to lack of data, no elimination in WWTP was assumedfor 55 and 66 pharmaceuticals in the general hospital and psychiatriccentre respectively, which might introduce biases to the finalconclusions. Overall, the study allows setting priorities risks forfurther testing and the related equations are also good references forother ERA studies.

Another study by Lee et al. (2008) assessed the environmentalrisks of most concerned antibiotics at the Gapcheon WWTP inDaejeon, Korea. The exposure doses of the chosen 13 antibioticswere estimated by a Korea ecological risk assessment (KOREOCORisk)model. The model consists of a release rate estimation module, anexposure estimation module and an ERAmodule. The release rate andbiodegradation removal efficiencies of the WWTP were taken intoaccount in the release rate estimation module whereas the multime-dia fate model rather than dilution factor was applied for calculatingthe site specific regional PECs in the exposure estimation module.Comparatively, PNECs were collected from open literature, review,the ECOTOX database or calculated from the ECOSAR model. Themodel outcomes indicated that RQs of amoxicillin and erythromycinwere 151 and 3, respectively, which may chronically degrade theKorean aquatic environment. Lee et al. (2007) also verified theeffectiveness of KOREOCORisk in setting management priority amongindustrial chemicals in the aquatic environment. Nevertheless, thecalculations inside the model are complicated which involves callingsub-models and the uncertainties vary greatly (10 to 103).

2.3.3. Characteristics and weaknesses of ERA modelsAccording to the above studies, ERA models were able to identify

the environmental risks of most inorganic and trace organic

Table 6Hospital wastewater treatment scenarios and associated PEC calculations.a

Option Descriptions

1 Risk potential of wastewater from the hospital main wing, befo(full risk potential without any degradation or dilution)

2 Risk potential at inlet of the WWTP (reduction of risk potential3 Risk potential at outlet of the WWTP (reduction of risk potentia

sorption processes with dilution in sewers)4 Risk potential at the hospital main wing after hypothetical conv

(reduction of risk potential through degradation and sorption p

PECHWW was the concentration of active ingredient expected in hospital wastewater. M is theamounts mi of active ingredient consumed in the different drug preparations. mi can be deingredient contained in each unit, mUi. df was the dilution factor in the sewer and assumed

a Modified from Escher et al. (2011).

compounds in recycled water. Some models address only one or afew of numerous components of the physical processes regardinghazard exposure and degradation, while others attempt to take morecomprehensive forms. However, some weaknesses still exist that areneeded to be modified and avoided. For instance, apart frominstrumental methods, most of PEC values were calculated based onworst case scenarios so that the RQs are likely to be overestimated. Toarrive at more accurate and reliable PEC values, more data should becollected regarding the chemical metabolism, spreading routes,environmental biodegradation, bioaccumulation potential, partition-ing characteristics, human use patterns, wastewater treatment andcatchment conditions. Meanwhile, if data on site-specific conditionsare available, several models such as Simple Treat, PharmaceuticalAssessment and Transport Evaluation and Geo-referenced RegionalExposure Assessment Tool for European Rivers can be further appliedto refine the PECs towards more realistic.

As for PNECs, current ecotoxicity tests rely solely on limitedaquatic organisms (e.g., daphnia and algae) which can not representecotoxicological responses of the whole ecosystem. Besides, most ofecotoxicity tests are based on acute responses and are not able toreflect the potential chronic effects following long term exposure tosubacute levels. Although the acute to chronic ratios are available forseveral substances, they are empirical and sometimes the potency tocause chronic ecotoxic effects is not correlated with potency to causeacute effects. Moreover, since pharmaceuticals mostly exist ascomplex mixtures, synergistic effects may occur. For example,clofibric acid combined with carbamazepine as well as diclofenaccombined with ibuprofen exhibited a much stronger toxic effect thanthe sum of their individual effects (Cleuvers, 2003). Consequently,future work should address these difficulties and create a preliminaryrisk assessment database where the different grades of recycledwater versus corresponding environmental risks are clarified. A riskranking of a series of hazardous compounds that may threaten theenvironment should also be established (Carlsson et al., 2006; Cooperet al., 2008; Jones et al., 2002). Overall, as recycled water end uses areincreasingly being expanded, ERA will play a more and moreimportant part and if possible, it should be carried out togetherwith additional technical, social and economical assessments to forma holistic recycled water sustainability analysis (Fig. 7) (Benetto et al.,2007; Claassen, 1999).

3. Integrated assessment tools on recycled water schemes

In most cases, it is difficult to model or analyse all accumulativeand interactive environmental effects of different activities from asingle assessment tool as each approach evaluates the recyclingproject in different ways. For example, MFA can be seen as aneffective tool for the early recognition of environmental sanitationproblems as well as the assessment of control measures. It allowsdecision makers to obtain a first efficient screening of potential

Equations for PEC calculation

re discharging to the sewer PECHWW ¼ M⋅f excretedVHWW

M ¼ Pn

i¼1mi ¼

Pn

i¼1UimUi

through dilution in sewers) PECWWTP influent ¼ df ⋅PECHWW

l through degradation and PECWWTP effluent ¼ f e⋅PECWWTP influent

entional biological treatmentrocess without dilution)

PECWWTP effluent ¼ f e⋅PECHWW

amount of each active ingredient consumed in the hospital. It was summed up from allrived from the units consumed for each drug preparation, Ui, and the amount of activeto be 0.013. fe referred to the fraction eliminated in the treatment plant.

Risks

ENVIRONMENTAL FAILURE

SOCIAL DISRUPTION

EC

ON

OM

ICA

L

LO

SSE

S

TE

CH

NIC

AL

F

AIL

UR

E

EN

VIR

ON

ME

NT

ECONOMY

SOC

IET

Y

ENGINEERINGECONOMIC

EFFECTIVENESS

ENVIRONMENTAL SAFETY

SOCIAL EQUITY

TECHNICAL VIABILITY

Fig. 7. Risks and objectives for sustainable water reuse.Adapted from Urkiaga et al. (2008).


environmental effects. This could be vital important in the holisticrecycled water sustainability analysis. When the current local envi-ronmental sanitation condition is predicted to be sustainable with nochanges required, there would be no need to conduct furtherassessments or actions, saving lots of time and energy. On the otherhand, if the environment situation is proved to be unsustainable, theMFA model can be executed again under different environmentalmanagement scenarios and the effectiveness of water recycling andreuse in sanitation improvement could be verified eventually. Oncethe importance of recycled water has been increasingly noticed bydecision makers, more and more recycling schemes together withdetailed sustainability assessments are likely to be developed andimplemented. In some developed countries with a large number ofrecycling schemes being successfully conducted, the significance ofwater reuse has already been widely recognised. MFA, in this case,could be regarded as an optional tool in the holistic decision makingof existing or future recycled water projects.

Comparatively, LCA is a non-site-specific analysis. Although it isconceived to carry out detailed and complex analyses, it has broaderambitions and gives overall environmental results as both the input andoutput related interventions are considered and all the releasesof different substances (e.g., CO2 for the whole life cycle) to theenvironmental media (e.g., air, soil or water bodies) could besummarised. It can be regarded as an important tool in the context ofholistic recycled water sustainability analysis since the selection of anappropriate treatment technology (or technology combination) couldnot merely benefit the local environment directly but also guide thedownstreamwater quality assessment aswell as end use consideration.However, it would be unnecessary to carry out comprehensive LCAassessment in most recycling schemes which normally involves anumber of different types of environmental impact categories. To savetime and cost for the whole sustainability decision making process,the assessment categories in LCA should be narrowed down afterconsidering the potential downstream end uses in initial recycledwaterproject planning. For example, when the recycled water is planned forirrigation uses, more concerns should be given to the potentialenvironmental impacts on soil and groundwater (e.g., EP, SP, HTP andTEP). Other impact categories related to air quality (e.g., GWP, AP, PHOand ODP) might be more important for industrial applications ofrecycled water. Yet although the insignificant categories have been

excluded, some difficulties may still hamper the quantification ofenvironmental effects in LCA, including the subjectivity in calculation,lack of consideration on alternatives, politicisation of assessmentprocesses and competence of involved authorities (Asano et al., 2007;Pasqualino et al., 2010; Zhang et al., 2010). Moreover, additionalbenefits (e.g., water quality improvement, fertiliser consumption re-duction and saltwater intrusion prevention) should be further clarifiedin downstream studies.

Besides, ERA mainly evaluates the environmental impacts in asite-specific way, which considers the possible releases of a singlesubstance from the different sources and tries to predict the risks ofecotoxicity-related impacts from that particular substance. In theholistic recycled water sustainability analysis, ERA models could beadopted to investigate the particular inorganic and trace organiccompounds of concern so as to address the output-related environ-mental interventions caused by different recycled water end uses.This could provide useful information for decision makers to adoptfurther management strategies in recycling schemes. When thesedifferent types of tools can be linked together, a broader set of criticalenvironmental issues could be encompassed in study and theweaknesses of individual tools might be compensated for in analysisprogress (Sanne and Widheden, 2005; Udo de Haes et al., 2006;Wrisberg and Udo de Haes, 2002). The integrated approaches havealready been reported in several studies.

3.1. MFA coupled with LCA

As MFA has been successfully applied in the early recognition ofenvironmental problems, it is possible to be regarded as a prerequi-site for the implementation of LCA. Instead of evaluating a wide rangeof environmental impact categories, LCA, in this circumstance, canfocus on the impacts associated with one or several particularelements which are tracked by MFA, saving lots of evaluation timeand energy. Thus, the Organic Waste Research model (ORWARE) isdeveloped to calculate material flows and energy turn over forvarious treatment alternatives and transfer the results into environ-mental effects using LCA methodology. A total of 43 differentsubstances are considered in the model, where the related transfor-mation, transportation, energy and other external resource consump-tion can be simulated. As can be seen in Fig. 8, the flow data generated

Input data

ORWARE SIMULINK Model

Output data X

Factor Matrix

Effect Categories

Fig. 8. Simplified diagram of the ORWARE methodology.Adapted from Ramírez et al. (2002).


by ORWARE is aggregated in a form of effect categories, whichindicate the level of environmental damage and resource consumptionin terms of emissions to air and water, accumulation in landfill, flow ofrecycled products, etc. This static model can be further divided intoseveral sub-models (e.g., WWTP, incineration, landfill, compost,anaerobic digestion, truck transports), where WWTP model is one ofthemost important ones which is to calculate the emissions and energyfrom wastewater treatment and reuse. Besides, as ORWARE itself israther general, other new sub-models can be easily incorporated intothe system. Nevertheless, most sub-models are empirical and fail toconsider the site-specific conditions (e.g., geographical, industrial andsocial factors). Hence, advanced recalibration is required whenevernecessary (Dalemo et al., 1997; Sonesson et al., 1997).

More specifically, Jeppsson and Hellstrom (2002) used ORWAREmodel to evaluate two fundamentally different urban water systems inSweden. One is a centralised system while the other one equips sourceseparation of stormwater, greywater, black water and urine togetherwith onsite treatment facilities. To effectively analyse these twosystems, several parameters (e.g., COD fractions, exergy, PO4, particu-late P, particulate N, potassium, cadmium, etc.) that are closely relatedto the environmental performance were modelled. The schematicmaterial flows are illustrated in Fig. 9. Overall, the centralised systemwas more environmentally friendly in terms of total P to water, cooperto water and cadmium to arable soil whereas the source-separatedsystem was better in regard to total N to water, fresh water

Drinking water production

Households

Incineration

Landfill

WWTP

Stormwater

Spreading and soil

(a)

Stormwater

Drinking water production

Households

Incineration

Landfill

WWTP

Spreading and soil

Urine collection and storage

Anaerobic digestion

(b)

Fig. 9. Principle material flows (e.g., N, P, K and Cd) through the combined system(a) and source-separated system (b).Adapted from Jeppsson and Hellstrom (2002).

consumption and net energy consumption. Similarly, ORWARE hasalso been applied to urban water systems in Chile by Ramírez et al.(2002). To compare 8 different management scenarios which aredifferent combinations of wastewater treatment (biological vs. chem-ical), sludge treatment (digestion vs. composting) and disposal (landfillvs. agriculture reuse), five environmental impact categories wereconsidered, including GWP, EP, pathogenic organisms, emissions ofheavy metals and toxic organic substances and energy use orproduction. The results indicated that the biological/composting/agriculture reuse scenario achieved an overall lowest environmentalimpact except for EP. Regarding the pathogen reduction in recycledwater, both chemical precipitation and biological treatment were goodenough for treated effluent discharging into the sea while additionaldisinfection was needed in the case of effluent to rivers. The effluentcould not be reused in agriculture as concentrations of toxic substanceswere too high.

3.2. LCA coupled with ERA

There has been effort to partially or fully incorporate ERA within aLCA approach in some of research areas, such as electronics,nanomaterials and mineral waste reuse scenarios (Montangero et al.,2006a; Socolof and Geibig, 2006; Sweet and Strohm, 2006). Overall, bylinking LCA and ERAwithin the same toolbox, thewhole of amaterial'slife cycle risk can be considered in an integrated manner, therebypromoting continuous improvement as well as proactive riskreduction and adaptive approaches under current situations. Howev-er, Udo de Haes et al. (2006) pointed out that the implementation of acombined approach required a careful study on similarities, differ-ences and synergisms between LCA and ERA. Although the fullintegration was recommended, it was not achieved in practice dueto the fundamentally different model structure (the use of thefunctional unit concept in LCA versus the use of flows of actual sizein ERA). The use of the two tools in a combined form has not beenreported in water recycling field. Yet it is possible to use ERA as amoredetailed and site-specific analysis after an LCA has been carried out.

3.3. Multi-criteria decision making

To achieve a more systematic and holistic analysis, the outputsfrom MFA, LCA and ERA models could be further combined withadditional qualitative or quantitative measures (e.g., economic andsocial assessments). Fig. 10 shows a multi-criteria decision makingframework for recycled water schemes based on the characteristics ofindividual and integrated assessment tools. When multiple conflict-ing criteria (e.g., technical, environmental, economic and socialissues) are taken into account, the following procedure in Multi-criteria Analysis (MCA) is to investigate the trade-off among theseselected criteria and then arrive at the best solution under certainmathematical algorithms (e.g., multi attribute utility theory, goalprogramming, compromise programming, analytical hierarchy pro-cess, ELECTRE I-III, PROMETHEE and cooperative game theory). Withan overall result in terms of feasibility, reliability and affordability ofthe recycled water scheme, MCA can provide a powerful guidance forsustainable water recycling management in the long term. It is alsopossible to suggest how much a successful strategy could benefit thedecision maker in water reuse planning and expansion. With thesehighly persuasive data, the public acceptability on recycled water

Integrated water reuse planning and expansion

Use of Material Flow Analysis (MFA) as an initial screening in understanding the environmental

sanitation conditions at local region

Environment sustainable Actions not

required

Environment not sustainable

Proposal of improved management strategies and verify the effectiveness of water recycling and reuse

using MFA

Use of Life Cycle Assessment (LCA) to analyse different wastewater treatment technologies with

respect to their environmental effects

Optimal treatment combinations

Application of Environmental Risk Assessment (ERA) on different grades of recycled water for

various end uses

Multi-criteria decision making of recycled water schemes

Social considerations

Economic considerations

Human health risk considerations

Fig. 10. Outline of the multi-criteria decision making framework for recycled water schemes.


applications can also be greatly improved and the potential recycledwater markets can be further expanded.

3.3.1. Economic assessmentMore specifically, the economic assessment of recycling schemes

is often carried out by a cost-benefit analysis, which should not onlyinclude the internal impacts (capital, operating and maintenancecosts and the price of recycled water) but also external impacts of anenvironmental or social nature. The prime objective is to maximisethe total benefits, which is the difference between income and costs(Eq. (13)).

MaxBT ¼ BI þ BE−OC ð13Þ

where BT=total benefit (total income−total costs), BI=internalbenefit (internal income− internal costs), BE=external benefit (posi-tive externalities−negative externalities), and OC=opportunity cost.Internal income can be earned by multiplying the selling price ofrecycled water and the volume obtained. In Australia, to encourage theusage of recycledwater, its price is currently set at approximately 75% ofthe drinking water price, which is much lower than the real cost ofproviding recycled water (Hurlimann and McKay, 2007; MacDonaldand Dyack, 2004). However, the interaction of willingness to adoptrecycled water and pricing strategies still has not reached a conclusiveresult so far (Dolničar and Saunders, 2006). Future pricing strategiesshould be based on costs as well as include the value of recycled wateritself, its environmental effects and its own opportunity cost.

The internal costs consist of:

• Investment costs. Investment costs account for 45–75% of the totalcost of a recycling project, which include land, civil works,machinery and equipment, distribution facilities and connectionworks (Hernández et al., 2006).

• Financial costs. Financial costs result from financing the investment.Some projects have been state-financed or funded by private

initiatives while others have received public participation in theform of investment subsidies, long-term loans or interest rebates.

• Operating and maintenance costs. The costs include water treatment,storage systems and pressure maintenance, water quality monitor-ing and life cycle costs (Urkiaga et al., 2008).

• Taxes. Taxes should also be considered if the scheme attracts tax(Godfrey et al., 2009).

To ensure the internal benefit, Urkiaga et al. (2006) have specifieda minimum capacity of the agricultural scheme, which is to serve10,000–20,000 inhabitant equivalents, or to irrigate a golf course and/or a crop extension of 3,500,000 m2. They also indicated that two ormore different types of treatments and end uses are more econom-ically suitable than a single option. While the internal impacts can becalculated directly in terms of monetary units, there are a series ofexternal influences where no explicit market exists. Hence, externalbenefit is to capture the most tangible and measurable ones andquantify the available aspects based on hypothetical scenarios orpatterns observed in related markets. For instance, Godfrey et al.(2009) identified the health benefits of greywater reuse from reducednumber of diarrhoeal cases, work/school absenteeism avoided, etc.Besides, the opportunity cost is normally estimated by an alternativeuse of the land with certain profitability (Hernández et al., 2006).

3.3.2. Social assessmentThe considerations in social assessment include aboriginal and

heritage, aesthetics, traffic disruption, public attitude and politicalimpacts (Muthukaruppan et al., 2011). As the principal target is toensure the smooth develop of recycling schemes with maximumbenefits, public acceptance is highly addressed, which is being re-cognised as the main ingredient of success for any recycling project.Dolničar and Saunders (2005) concluded that the acceptance ofrecycled water was correlated with a high level of education, followedby being in the younger age category, while income and genderappeared significant in only one third of the studies. Hurlimann (2007)


indicated that males had more knowledge about recycled water andwere more supportive than females. McKay and Hurlimann (2003)predicted that the greatest opposition to water recycling schemeswould be people aged 50 years and over. Po et al. (2003) suggested thatthere are many other factors influencing the acceptance, including thedisgust emotion, perceptions of risk from recycled water, sources ofrecycled water, specific uses of recycled water, trust of authorities andknowledge, attitudes towards the environment, environmental justiceissues, the cost of recycled water and socio-demographic factors. Theperception of risks from the use of recycled water are related to health,foremost among people's worries are the safety of their children.

The higher water quality such as low salt, colourless and odourlesscould contribute to the increased acceptability (Hurlimann andMcKay, 2007). Respondents were most willing to pay for qualityincreases of recycled water when used for clothes washing. Non-potable reuse carries the least public health risk and the publicsupports for agriculture, golf courses, parks and industries aregenerally high. Recycledwater use inside the homewas less preferred,where more than 70% of the respondents agreed to use it for toiletflushing, gardens and car washing but only 60% and 13% supported forwashing clothes and filling swimming pools respectively (Pham et al.,2011). There are greater concerns on Indirect Potable Reuse (IPR)and Direct Potable Reuse (DPR) projects due to health issues.

Table 7Management approaches in sustainability improvement of particular end uses.

Possible end uses Solutions

Agricultural irrigation(pastures and foddercrops)

● Selection of advanced irrigation methods (e.g., d● Extension of the withholding period● Increase of capital costs

■ Additional on-site water treatment (e.g., floccu■ Signage

● Increase of maintenance costs■ Modified working hours■ Soil amendments (e.g., acid or gypsum injectio

Landscape and gardenirrigation

● Selection of advanced irrigation methods● Increase of droplet size of the water● Extension of the withholding period● Increase of capital costs

■ Additional on-site water treatment (e.g., floccu■ Signage and fencing■ Staff access protection

● Increase of maintenance costs■ Small absorbers for odour control■ Modified working hours and public access con■ Soil amendments (e.g., acid or gypsum injectio

Environmental flows andrecreational impoundments

● Increase of capital costs■ Additional on-site water treatment (e.g., floccu■ Clarifier

● Increase of maintenance costs■ Public access protection■ Algae control■ Small absorbers for odour control

Industry ● Increase of capital costs■ Additional on-site water treatment (e.g., coagu

acid, reverse osmosis, water softening and dem■ De-oiling and de-gasification pre-treatment■ Worker safety protection

Toilet flushing ● Increase of capital costs■ Spray controllers on toilet bowls■ Additional on-site water treatment (e.g., colou

and pathogen removal by membrane)■ Determination of residual chlorine concentrati

● Increase of maintenance costs■ Small absorbers for odour control■ Regular filter cleaning (2.5 min of labour per m

Car washing ● Increase of droplet size when spraying water● Increase of capital costs

■ Additional on-site water treatment (e.g., filtratcoagulation and membrane filtration)

■ Determination of residual chlorine concentrati● Increase of maintenance costs

■ Small absorbers for odour control

Encouragingly, Lampard et al. (2010) reported that 74% of 3000respondents across five Australian capital cities expressed theirwillingness to drink recycled water if they could be assured of itssafety.

Apart from water quality improvement, public participation mayalso play an important role in enhancing the public accountability andthus acceptability of recycled water schemes. It can occur througheducation, information dissemination, advisory or reviewboards, publicadvocacy, public hearings and submissions and even litigation(Richardson and Razzaque, 2006). Pham et al. (2011) suggestedinforming people about the different benefits and terms of recycledwater as well as continuously seeking feedback from the communitywhereas Hurlimann (2007) believed that authorities should focus ongaining the community's trust. A willingness to use model constructedbyMenegaki et al. (2007) also indicated that information and educationmight be useful tools in making people realise the benefits of recycledwater, allowing them to pass fromnegative to positive attitudes. On thecontrary, based on Ajzen's theory of planned behaviour, Po et al. (2005)developed a model to examine participant's behavioural responses torecycled water. The model on an IPR case study at the ManagedAquifer Recharge Scheme in Perth, Western Australia, showed thatknowledge and risk perceptions were not dominant in influencingbehavioural intentions to drink the recycled water. Nevertheless, the

References

rip or sub-surface irrigation) Al-Hamaiedeh and Bino (2010)and Shanahan and Boland (2008)

lation and disinfection)

n)Stevens et al. (2008) and Suzukiet al. (2002)

lation and disinfection)

troln)

Suzuki et al. (2002)lation and disinfection)

Brissaud (2008) and Chiou et al.(2007)lation, precipitation, ion exchange,

ineralization)

March et al. (2004) and Nolde(1999)

r removal by flocculation and coagulation;

on

3)Asano et al. (2007)

ion, sedimentation; flocculation,

on


relationships between people's level of knowledge, perceptions andacceptance of recycled water are fairly complex, broad conclusions onthese factors from small amount of research are still not enough(Marks, 2003). Although the difficulties and complexities in the fastimprovement of public acceptance have been witnessed, greaterpublic input via participation may still promote environmental justiceand assist decision makers to understand and identify public interestconcerns while formulating recycled water implementation policiesto some extent (Richardson and Razzaque, 2006). While most of thecurrent social perception studies are qualitative descriptions, moreadvanced quantitative models as well as indexes should be developedin the future, which could provide highly persuasive informationto decision makers, local water authorities, stakeholders and thecommunity.

3.3.3. MCA implementationFinally, when assessment results from the multi-criteria decision

making are proved to be unsustainable, several management orcontrol approaches should be carried out (Table 7) and the recyclingscheme can be re-evaluated in this integrated approach. However, themodel structure and boundary differences of different tools are stillneeded to be carefully considered, which are likely to introducemisinterpreted conclusions at the end. For instance, when applyingLCA to previous MCA results, the extended system boundaries arenormally hard to define, which include other parts of urbaninfrastructure (e.g., incineration plant, landfill, arable and receivingwaters). Besides, the outcomes from LCA and ERA may suggestopposite solutions for the environmental preferred choice as some-times the wastewater treatment technology with lowest environ-mental impacts could not arrive at the highest recycled water qualityfor environmental protection. These outcomes, together with resultsfrom economic and social assessments should be weighted to ensurethe final evaluation results are consistent with the preferences ofdecision makers.

4. Conclusion

Recycled water has received great attention over the recentdecades due to its numerous advantages, such as low energyconsumption, availability, low effluent and sludge discharge andreduced freshwater consumption. As the sustainability of waterrecycling system directly influences the introduction of new enduses and the expansion of the current scheme, several conventionalenvironmental assessment tools have been increasingly applied inthis field. MFA was found to be an effective initial screening inunderstanding the environmental sanitation conditions at localregion. LCA has been widely used in selecting the optimal wastewatertreatment technologies while ERA is mainly used in evaluating thepotential effects of particular chemical hazards in recycled water onsoil, surface water and groundwater. However, several limitations stillexist in these individual tools in terms of boundary conditions, spatialspecification, the interventions and impact types that are considered.The application of integrated tools together with qualitative orquantitative economic and social assessments have addressed theseweaknesses and provided a more systematic and holistic frameworkin recycledwater decisionmaking. Nevertheless, the complexities andproblems in model integration have also been observed, which arelikely to impede the further development of combined models. Assuch, future work should test the performance of integrated approachunder different water reuse scenarios and refine the system bound-aries and assessment equations over time. Besides, further improve-ments such as additional onsite wastewater treatment, advancedirrigation or other recycledwater usemethods, access barriers are alsoconsidered essential.

Acknowledgement

This work was funded by Australian Research Council (ARC)Industry Linkage Grant (LP100100494). The authors acknowledge thesupport from Allconnex Water, City West Water, Port MacquarieHasting Council and Sydney Olympic Park Authority.

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http://www.pwri.go.jp/eng/activity/pdf/reports/Suzuki-yutaka020327.pdf

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