sustainability assessment of wastewater and sludge
TRANSCRIPT
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Sustainability Assessment of Wastewater and
Sludge Treatment Techniques for Removal of
Compounds from Pharmaceuticals and Personal
Care Products (PPCPs)
A thesis submitted to the University of Manchester for the degree of Doctor of
Philosophy in the Faculty of Engineering and Physical Sciences
2016
Raphael Ricardo Zepon Tarpani
Faculty of Science & Engineering
School of Chemical Engineering and Analytical Science
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Table of contents
Acknowledgements ........................................................................................................ 19
1. INTRODUCTION ..................................................................................................... 20
1.1. BACKGROUND ............................................................................................. 20
1.2. RESEARCH AIMS AND OBJECTIVES ........................................................ 21
1.3 STRUCTURE OF THE THESIS ...................................................................... 22
2. LITERATURE REVIEW ......................................................................................... 24
2.1. PHARMACEUTICALS AND PERSONAL CARE PRODUCTS .................. 24
2.2. PPCP COMPOUNDS IN NATURE ................................................................ 26
2.3. PPCP COMPOUNDS IN WASTEWATER TREATMENT PLANTS ........... 35
2.4. EUROPEAN REGULATIONS RELATED TO PPCP COMPOUNDS ......... 45
2.5. ADVANCED WASTEWATER AND SLUDGE TREATMENT
TECHNIQUES ........................................................................................................ 48
2.6. WASTEWATER TREATMENT AND SUSTAINABLE DEVELOPMENT 68
3. METHODOLOGY FOR SUSTAINABILITY ASSESSMENT ........................... 75
3.1. METHODOLOGY FOR ESTIMATING CONCENTRATION OF PPCP
COMPOUNDS IN WWTPS ................................................................................... 76
3.2. OPERATING PARAMETERS, RESOURCE RECOVERY AND REMOVAL
OF PPCP COMPOUNDS ....................................................................................... 83
3.3. SUSTAINABILITY ASSESSMENT .............................................................. 88
4. METHODOLOGY FOR ESTIMATING CONCENTRATIONS OF PPCP
COMPOUNDS IN WWTPS ....................................................................................... 106
4.1. ESTIMATION OF INFLUENT FLOW IN WWTPS .................................... 106
4.2. ESTIMATION OF INFLUX OF PPCP COMPOUNDS INTO WWTPS AND
REMOVAL RATES ............................................................................................. 107
4.3. ESTIMATION OF CONCENTRATION RANGES OF PPCP COMPOUNDS
IN WWTPS ........................................................................................................... 111
4.4. ESTIMATION OF FRESHWATER CONCENTRATIONS OF PPCP
COMPOUNDS ...................................................................................................... 117
4.5. CHAPTER CONCLUSIONS ........................................................................ 119
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5. LIFE CYCLE ASSESSMENT OF WATERWATER TREATMENT
TECHNIQUES ............................................................................................................ 121
5.1. GOAL AND SCOPE ..................................................................................... 121
5.2. INVENTORY ANALYSIS ............................................................................ 122
5.3. LIFE CYCLE IMPACTS RESULTS AND DISCUSSION .......................... 132
5.4. PARAMETRIC ANALYSIS ......................................................................... 139
5.5. FRESHWATER ECOTOXICITY POTENTIAL OF PPCP COMPOUNDS 139
5.6. WASTEWATER REUSE .............................................................................. 142
5.7. CHAPTER CONCLUSIONS ........................................................................ 143
6. LIFE CYCLE ASSESSMENT OF SLUDGE TREATMENT TECHNIQUES 145
6.1. GOAL AND SCOPE ..................................................................................... 145
6.2. INVENTORY ANALYSIS ............................................................................ 145
6.3. LIFE CYCLE IMPACTS RESULTS AND DISCUSSION .......................... 150
6.4. SENSITIVITY ANALYSIS .......................................................................... 157
6.5. FRESHWATER ECOTOXICITY OF PPCP COMPOUND AND HEAVY
METALS ............................................................................................................... 160
6.6. CHAPTER CONCLUSIONS ........................................................................ 165
7. LIFE CYCLE COSTING OF ADVANCED WASTEWATER AND SLUDGE
TREATMENT TECHNIQUES ................................................................................. 167
7.1. GOAL AND SCOPE ..................................................................................... 167
7.2. COSTS ESTIMATION AND DATA SOURCES ......................................... 168
7.3. RESULTS AND DISCUSSION .................................................................... 172
7.4. SENSITIVITY ANALYSIS .......................................................................... 176
7.5. ECONOMIC FEASIBILITY OF WASTEWATER REUSE AND
RESOURCE RECOVERY FROM SLUDGE ...................................................... 179
7.6. CHAPTER CONCLUSIONS ........................................................................ 180
8. INTEGRATED SUSTAINABILITY ASSESSMENT OF WASTEWATER AND
SLUDGE TREATMENT METHODS ...................................................................... 182
8.1 SUMMARY OF LIFE CYCLE ENVIRONMENTAL IMPACTS ................ 182
8.2 SUMMARY OF LIFE CYCLE COSTS ......................................................... 184
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8.3. SOCIAL LIFE CYCLE IMPACT ASSESSMENT ....................................... 186
8.4. INTEGRATED SUSTAINABILITY ASSESSMENT .................................. 200
8.5. CHAPTER CONCLUSIONS ........................................................................ 208
9. CONCLUSIONS, RECCOMENDATIONS AND FUTURE WORK ................ 209
9.1. PPCP COMPOUNDS IN WWTPS ................................................................ 209
9.2. SUSTAINABILITY ASSESSMENT ............................................................ 214
9.3. RECOMMENDATIONS ............................................................................... 218
9.4. FUTURE WORK ........................................................................................... 221
10. REFERENCES ...................................................................................................... 224
11. SUPPLEMENTARY INFORMATION .............................................................. 261
11.1. CHAPTER 4 SUPPLEMENTARY INFORMATION ................................ 261
11.2. CHAPTER 5 SUPPLEMENTARY INFORMATION ................................ 270
11.3. CHAPTER 7 SUPPLEMENTARY INFORMATION ................................ 272
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List of Figures
Figure 1 - Molecular transformations of parent PPCP compounds in living organisms and
the environment ............................................................................................................... 25
Figure 2 – Main source and transport routes of PPCP compounds during their life cycle
......................................................................................................................................... 29
Figure 3 - Main transport and degradation mechanisms of PPCP compounds in the
environment..................................................................................................................... 32
Figure 4 – Representation of standardized PNEC values (adapted from Roman et al. 1999)
......................................................................................................................................... 33
Figure 5 – Typical solid matter content in sewage sludge (adapted from Jordão & Pessôa
1995) ............................................................................................................................... 36
Figure 6 – Main removal process of PPCP compounds during conventional wastewater
treatment plants ............................................................................................................... 44
Figure 7 – Selected options for advanced wastewater and sludge treatment and their
respective products .......................................................................................................... 52
Figure 8 – Scheme of granular activated carbon treatment and main removal mechanism
of micro-contaminants in granular activated carbon particles ........................................ 54
Figure 9 – Scheme of nanofiltration and main removal mechanisms of micro-
contaminants in nanofiltration membranes ..................................................................... 56
Figure 10 – Scheme of a solar-photo Fenton treatment panel for wastewater treatment 59
Figure 11 - Scheme of ozonation process for wastewater treatment .............................. 61
Figure 12 – Scheme of anaerobic digestion of thickened sludge for agricultural
application ....................................................................................................................... 63
Figure 13 - Scheme of composting of thickened sludge for agricultural application ..... 65
Figure 14 – Scheme of thickened sludge incineration for electricity and heat recovery 66
Figure 15 – Scheme of thickened sludge pyrolysis for recovery of heat, bio-oil and
biochar ............................................................................................................................. 67
Figure 16 – Scheme of thickened sludge wet air oxidation for recovery of methanol (for
denitrification) ................................................................................................................. 68
Figure 17 – Tripartite interception approach defining the sustainable development goals
(cross hatched area) ......................................................................................................... 69
Figure 18 – Role of advanced wastewater and sludge treatment techniques in integrated
wastewater reuse of resource recovery management ...................................................... 74
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Figure 19 – Methodology for sustainability assessment of advanced wastewater and
sludge treatment techniques for the removal of PPCP compounds in wastewater treatment
plants ............................................................................................................................... 75
Figure 20 - Life cycle assessment methodology according to ISO 14044 (2006) .......... 89
Figure 21 – System boundaries for life cycle assessment of the advanced wastewater and
sludge treatment techniques ............................................................................................ 90
Figure 22 – The methodology for creation of life cycle inventories considered in this work
......................................................................................................................................... 91
Figure 23 – Sensitivity analysis assuming variations in the quality of the secondary
effluent and thickened sludge.......................................................................................... 93
Figure 24 – Axis configuration for the integration of nexus indicators (nexus triangle) 99
Figure 25 - Nexus influence (Anexus) according different vk values .............................. 100
Figure 26 – Correlation between daily water influent Q and population p served by
WWTPs based on the data in Table 1 ........................................................................... 107
Figure 27 – Annual per-capita discharge IMinf,i of target PPCP compounds estimated
using eqn. (5) and data from Table 1. Each point on the graph represents IMinf for one
target compound i .......................................................................................................... 108
Figure 28 – Removal rates Rrate,i for target PPCP compounds estimated using eqn. (6) and
data from Table 1. Each point in the graph represents Rrate,i for one target compound i
....................................................................................................................................... 109
Figure 29 – Normalized and weighted results for the number of data points for IMinf,i
(dataset A) by world region........................................................................................... 110
Figure 30 – Minimum (a) and maximum (b) daily influx of target PPCPs estimated
according to eqn. (7) for different size of the population served by WWTPs .............. 112
Figure 31 – Estimated range of WWTP removal rates (Rrange,i) for the target compounds
(q = 428 L/inhab.day) .................................................................................................... 114
Figure 32 – Predicted environmental concentration (PEC) of target PPCP compounds in
freshwaters for the mean expected effluent concentration (γmean,i) and for different
freshwater flows: F1 = 5bn L/day; F2 = 500 M L/day; F3 = 100 M L/day; F4 = 50 ML/day
....................................................................................................................................... 118
Figure 33 – System boundaries and life cycle stages of the advanced wastewater treatment
techniques considered in the study (*Excluded for ozonation due to a lack of data) ... 122
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Figure 34 – Estimated amounts of fresh and regenerated granular activated carbon for
1,000 m3 of wastewater treated for different bed service times and empty-bed contact
times (EBCT) (nmax:10, mloss:10%, GAC density: 564 kg/m3) ..................................... 124
Figure 35 – Amount of fresh and regenerated granular activated carbon in contactors
according the number of bed regenerations (mloss:10%) ............................................... 124
Figure 36 – Estimated electricity consumption per 1,000 m3 of wastewater for different
ozone transfer efficiencies and applied ozone dosage .................................................. 126
Figure 37 – Graphical illustration of the advanced treatment methods considered in the
study .............................................................................................................................. 127
Figure 38 – Best-fit curves for the estimation of the removal rates of the target PPCP
compounds by the advanced treatment techniques based on experimental data in the
literature. Data points include some non-target compounds to improve the reliability of
the estimates .................................................................................................................. 129
Figure 39 – Life cycle impact of the advanced wastewater treatment techniques for PPCP
compounds (error bar represents minimum and maximum values for the parameters as
specified in Table 16). All impacts are expressed per 1,000 m3 of wastewater ............ 137
Figure 40 – Contribution of different life cycle stages to the impacts of advanced
treatment options ........................................................................................................... 138
Figure 41 – Freshwater ecotoxicity potential of effluents released from advanced
wastewater treatments compared to the impact from effluent with no advanced treatment
(estimated using USEtox methodology). The impact for “No treatment” in figure b) has
been multiplied by a factor of 10 to show on the scale ................................................. 140
Figure 42 – Fuel sources used in the electricity grid supply between 2000 and 2015 in the
UK ................................................................................................................................. 144
Figure 43 - Overview of the sludge treatment methods considered in the study showing
the recovery of resources (fertilizer, heat, electricity, fuels and methanol) .................. 149
Figure 44 - Life cycle impacts of sludge treatment techniques expressed per 1,000 kg DM
(The error bars represent the minimum values for the recovery of resources specified in
Table 1. DB: dichlorobenzene; PM10: particulate matter, 10µm; NMVOC: non-methane
volatile ........................................................................................................................... 155
Figure 45 - Contribution of different life cycle stages to the impacts of advanced treatment
options (The values refer to the maximum recovery of resources. ADG: anaerobic
digestion; COM: composting: INC: incineration; PYR: pyrolysis; WAO: wet air
oxidation) ...................................................................................................................... 156
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Figure 46 – (cont.) The effect of different resource recovery rates on the environmental
impacts of different sludge treatment techniques (100%, 50% and 0% refer on the x-axis
represent the maximum, mean and minimum values, respectively, for the recovery of
resources from different treatment options. ADG: agricultural application of
anaerobically digested sludge; COM: composting; INC: incineration; PYR: pyrolysis;
WAO: wet air oxidation) ............................................................................................... 160
Figure 47 – Total freshwater ecotoxicity potential (including PPCPs and heavy metals)
of the sludge treatment techniques according to the USEtox methodology (ADG:
anaerobic digestion; COM: composting: INC: incineration; PYR: pyrolysis; WAO: wet
air oxidation) ................................................................................................................. 163
Figure 48 – Freshwater ecotoxicity potential estimated according to the USEtox
methodology and based on the legislative limits for heavy metals in sludge applied to
agricultural land in some European countries and in the US in relation to the range of
impact from heavy metals estimated in this work for different sludge treatment methods
(horizontal red lines). The impact takes into account only direct emissions from the
application of the sludge (i.e. it is not on a life cycle basis) ......................................... 164
Figure 49 - Life cycle costs of the advanced wastewater treatment techniques showing
the contribution of different stages (The data labels represent the costs for the average
and the error bars for the minimum and maximum values of the parameters in Table 25)
....................................................................................................................................... 173
Figure 50 - Contribution of different life cycle stages to the costs advanced of advanced
wastewater treatment techniques for the average operating parameters (For the latter, see
Table 25. NF: nanofiltration; EDTA: ethylenediaminetetraacetic acid) ....................... 174
Figure 51 - Life cycle cost of sludge treatment techniques showing the contribution of
different stages (the data labels represent the costs for the average and the error bars for
the minimum and maximum values of the parameters in Table 26.). Values for transport
in pyrolysis includes waste management of non-recovery resources ........................... 175
Figure 52 - Contribution of different life cycle stages to the costs of sludge treatment
techniques for the mean resource recovery (for the latter, see Table 26) ..................... 176
Figure 53 – Influence of energy costs on the life cycle costs of advanced wastewater (a)
and sludge (b) treatment techniques (The vertical bars show the average LCC costs and
the error bars the minimum and maximum costs of energy given in Table 29) ........... 177
Figure 54 – Influence of the costs of chemicals on the life cycle costs of advanced
wastewater (a) and sludge (b) treatment techniques (The vertical bars show the average
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LCC costs and the error bars the minimum and maximum costs of chemicals given in
Table 27) ....................................................................................................................... 178
Figure 55 – Influence of the costs of activated carbon and membranes on the life cycle
costs of granular activated carbon and nanofiltration (The vertical bars show the average
LCC costs and the error bars the minimum and maximum costs of these materials given
in Table 28) ................................................................................................................... 178
Figure 56 – Comparison of costs estimated in this work for the production of potable
water from wastewater with water and sewage costs in the UK and costs of desalination
worldwide (*Membrane bioreactor coupled with one of the advanced wastewater
treatment techniques operating at the average operating requirements; distribution of the
reclaimed wastewater to the end user not included) ..................................................... 180
Figure 57 –Potential environmental life cycle impacts of the advanced wastewater
treatment techniques for the mean operating conditions. Results per 1,000 m3 of
secondary effluent ......................................................................................................... 183
Figure 58 - Potential environmental life cycle impacts of the sludge treatment techniques
at the mean operating conditions. Results per 1,000 kg of dry matter .......................... 184
Figure 59 – Life cycle costs of the advanced wastewater treatment techniques showing
the contribution of different stages ............................................................................... 185
Figure 60 – Life cycle cost of sludge treatment techniques showing the contribution of
different stages .............................................................................................................. 186
Figure 61 – Impact of the advanced wastewater treatment techniques on the energy-water-
food nexus (integration of nexus indicators) for the minimum, mean and maximum
operating requirements .................................................................................................. 191
Figure 62 - Impact of the sludge treatment techniques on the energy-water-food nexus
(integration of nexus indicators) for the maximum, mean and no recovery of resources
....................................................................................................................................... 197
Figure 63 – MCDA results for the advanced wastewater treatment techniques with the
equal weights for the sustainability indicators and environmental, economic and social
dimensions of sustainability: (a) minimum operating requirements; (b) mean operating
requirements; and (c) maximum operating requirements ............................................. 201
Figure 64 – MCDA results for the advanced wastewater treatment techniques with the
environmental dimension of sustainability five times more important: (a) minimum
operating requirements; (b) mean operating requirements; and (c) maximum operating
requirements .................................................................................................................. 202
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Figure 65 - MCDA results for the advanced wastewater treatment techniques with the
economic dimension of sustainability five times more important: (a) minimum operating
requirements; (b) mean operating requirements; and (c) maximum operating requirements
....................................................................................................................................... 202
Figure 66 - MCDA results for the advanced wastewater treatment techniques with the
social dimension of sustainability five times more important: (a) minimum operating
requirements; (b) mean operating requirements; and (c) maximum operating requirements
....................................................................................................................................... 203
Figure 67 – Sensitivity analysis for the advanced wastewater treatment techniques for
different weights of importance for the sustainability dimensions ............................... 204
Figure 68 - MCDA results for the sludge treatment techniques with the equal weights for
the sustainability indicators and environmental, economic and social dimensions of
sustainability: (a) maximum resource recovery; (b) mean resource recovery; and (c) no
product recovery............................................................................................................ 205
Figure 69 - MCDA results for the sludge treatment techniques with the environmental
dimension of sustainability five times more important: (a) maximum resource recovery;
(b) mean resource recovery; and (c) no products recovery ........................................... 206
Figure 70 - MCDA results for the sludge treatment techniques with the economic
dimension of sustainability five times more important: (a) maximum resource recovery;
(b) mean resource recovery; and (c) no products recovery ........................................... 206
Figure 71 - MCDA results for the sludge treatment techniques with the social dimension
of sustainability five times more important: (a) maximum resource recovery; (b) mean
resource recovery; and (c) no products recovery .......................................................... 207
Figure 72 - Sensitivity analysis for the sludge treatment techniques for different weights
of importance for the sustainability dimensions ........................................................... 207
Figure 73 – Concept of the ultimate role of advanced wastewater and sludge treatment
techniques in the rational use of resources in the EWF nexus ...................................... 223
Figure 74 - Box plots for IMinf,i values showing the ranges of data found in the literature
for different world regions. White dots represent estimated mean values, horizontal lines
median values and small red dots the outliers ............................................................... 266
Figure 75 - The exchange rate of British Pounds (£) in the period 2006 - 2015 against the
US dollar (US$) and Euro (€), taking into account the inflation in the UK in the same
period............................................................................................................................. 272
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List of Tables
Table 1 - Target PPCP compounds in wastewater treatment plants in different countries
......................................................................................................................................... 77
Table 2 – Products that can be recovered from advanced sludge treatment techniques and
products that they potential displace ............................................................................... 87
Table 3 – Recipe 2008 midpoint impact categories considered in this work.................. 92
Table 4 – Social issues and indicators for social sustainability assessment of advanced
wastewater and sludge treatment techniques .................................................................. 95
Table 5 - Weights of importance for the environmental, economic and social indicators
considered in the MCDA .............................................................................................. 105
Table 6 – Outliers for the influx of PPCP compounds (A dataset) and removal rates
(dataset B) in WWTPs (data points in SI Table 36 and Table 37) ................................ 110
Table 7 – Distribution of data for IMinf,i (dataset A) in different world regions ........... 110
Table 8 - Estimated influent concentration ranges for the target PPCPa ...................... 113
Table 9 – Effect of acid dissociation constant (pKa) on estimated removal of PPCP
compounds by conventional WWTPs ........................................................................... 115
Table 10 – Estimated effluent concentration ranges for the target PPCP compoundsa 116
Table 11 – Estimated sludge concentration ranges for the target PPCP compoundsa .. 117
Table 12 – Operating parameters for GAC (eqns. (12)-(15)) and ozonation (eqns. (16)-
(17)) per 1,000 m3 of wastewater .................................................................................. 127
Table 13 – Original data of the advanced wastewater treatment techniques operation 128
Table 14 – Estimated efficiencies for the removal of the target PPCP compounds in the
advanced treatment plants (%) ...................................................................................... 130
Table 15 – Estimated concentrations of target PPCP compounds in effluents after the
advanced wastewater treatment (µg/L) ......................................................................... 130
Table 16 – Inventory data for the advanced wastewater treatment techniques (per 1,000
m3 of secondary effluent) .............................................................................................. 131
Table 17 – USEtox characterization factors for freshwater ecotoxicity of target PPCP
compounds .................................................................................................................... 140
Table 18 – Inventory data for the sludge treatment techniques (per 1,000 kg of dry matter)
....................................................................................................................................... 149
Table 19 - Heavy metals in sludge applied on agricultural land and emitted by incineration
....................................................................................................................................... 161
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Table 20 – USEtox characterization factors for freshwater ecotoxicity potential of PPCP
compounds and heavy metals........................................................................................ 161
Table 21 – Freshwater ecotoxicity potential of PPCP compounds and heavy metals
contained in sludge using the USEtox methodology .................................................... 162
Table 22 – Legislative limits for some heavy metals in the sludge applied on agricultural
land in some European countries and the US ............................................................... 164
Table 23 – Construction costs for the advanced wastewater and sludge treatment
techniques ...................................................................................................................... 169
Table 24 – Infrastructure replacement costs for the advanced wastewater treatment
techniques over the lifespan of the plant (60 years) ...................................................... 170
Table 25 – Operating, waste management and transport data for the advanced wastewater
treatment techniques (per 1,000 m3 of secondary effluent) .......................................... 170
Table 26 – Operating, waste management and transport data for the sludge treatment
plants (per 1,000 kg of dry matter) ............................................................................... 170
Table 27 – Prices of chemicals in the UK and imported from Chinaa .......................... 171
Table 28 – Prices of granular activated carbon and nanofiltration membranesa ........... 171
Table 29 – Energy prices in the UKa ............................................................................. 171
Table 30 – Costs of waste disposal and transport ......................................................... 172
Table 31 – Market prices of products replaced by the equivalent resources recovered by
sludge treatment ............................................................................................................ 172
Table 32 – Social sustainability assessment of the advanced wastewater treatment
techniques (per 1,000 m3 wastewater) .......................................................................... 187
Table 33 – Social sustainability assessment of the sludge treatment techniques (results
per 1,000 kg of dry matter) ........................................................................................... 188
Table 34 – Results for energy-water-food nexus impacts of the advanced wastewater
treatment techniques...................................................................................................... 190
Table 35 – Results for the energy-water-food nexus impacts of the sludge treatment
techniques ...................................................................................................................... 196
Table 36 - Estimated annual per-capita influx into WWTPs of target PPCP compounds
(dataset A)a .................................................................................................................... 267
Table 37 - Estimated removal rates for the target PPCP compounds (dataset B)a. ....... 268
Table 38 - Estimated daily influx for the target PPCP compounds for a WWTP serving a
population “p” ............................................................................................................... 269
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Table 39 - Estimated ranges for the removal of the target PPCP compounds in WWTPs
....................................................................................................................................... 269
Table 40 - Operating data for GAC, NF and SPF considered in the study ................... 270
Table 41 - Spiral wound modules inventory modules (Bonton et al. 2012) ................. 270
Table 42 - Freshwater ecotoxicity potential of effluents discharged to freshwaters
estimated according to the USEtox methodology ......................................................... 271
Table 43 - Freshwater ecotoxicity potential of effluents discharged to agricultural soils
estimated according to the USEtox methodology ......................................................... 271
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List of abbreviations
ADG anaerobic digested sludge
AF assessment factor
AOP advanced oxidation process
AS activated sludge
BCR European Community Bureau of Reference
DM dry matter
DOC dissolved organic carbon
DOM dissolved organic matter
DWTP drinking water treatment plant
EBCT empty-bed contact time
EC emerging contaminant
EDTA ethylenediaminetetraacetic acid
EEC Environmental European Commission
EMEA European Medicines Agency
ERA environmental risk assessment
EU European Union
EWF energy-water-food
GAC granular activated carbon
GC gas chromatography
HDPE high-density polyethylene
HPLC high-performance liquid chromatography
HRT hydraulic retention time
ISO International Standard Organization
IUWM Integrated Urban Wastewater Management
LCA life cycle assessment
LCC life cycle costing
LC-MS/MS liquid chromatography-tandem mass spectrometry
LOD limit of detection
LOEC lowest observable effect concentration
LOQ limit of quantification
MBR membrane bioreactor
MCDA multi-criteria decision analysis
MEC measured concentrations
MW molecular weight
MWCO molecular weight cut off
NF nanofiltration
NOEC no observable effect concentration
OECD Organization for Economic Co-operation
OM organic matter
OZO ozonation treatment
PEC predicted environmental concentration
PNEC predicted no-effect concentration
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POP persistent organic pollutants
PP polypropylene
PPCP pharmaceuticals and personal care product
SFP solar photo-Fenton
SLCA social life cycle assessment
SPE solid phase extraction
SRT sludge retention time
SS suspended solids
TE transfer efficiency
TP transformation product
UASB flow anaerobic sludge blanket digestion
UNEP United Nations Environment Programme
UNICEF United Nations Children's Fund
UK United Kingdom
US United States
USA United States of America
UV ultra-violet
WHO World Health Organization
WWTP wastewater treatment plant
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SUSTAINABILITY ASSESSMENTS OF WASTEWATER AND SLUDGE
TREATMENT TECHNIQUES FOR REMOVAL OF COMPOUNDS FROM
PHARMACEUTICAL AND PERSONAL CARE PRODUCTS (PPCPs)
Raphael Ricardo Tarpani, The University of Manchester, 2016
Submitted for the degree of Doctor of Philosophy
ABSTRACT
Environmental releases of chemical compounds from Pharmaceuticals and
Personal Care Products (PPCPs) are receiving growing attention in the scientific
community. Most research suggests that the main pathway for these substances to reach
the environment is via Wastewater Treatment Plants (WWTPs) due to the effluents from
households, industry and hospitals, which can contain substantial amounts of these
compounds. Many of these contaminants are poorly treated in conventional WWTPs and
are often discharged into the environment with the effluent and sludge, posing
ecotoxicological risks to the wildlife and humans. Therefore, it is necessary to limit their
release into the environment by controlling their discharge from WWTPs. This can be
achieved by adopting advanced wastewater treatment techniques, currently not used as
there are no legislative limits on PPCP compounds. However, as the scientific evidence
is growing on their adverse impacts, it is only a matter of time before their advanced
treatment becomes compulsory.
To help guide future developments and inform policy in this area, this work
considered a range of advanced treatment techniques with the aim of identifying the most
sustainable options. Adopting a life cycle approach and considering all three dimensions
of sustainability (economic, environmental and social), nine technologies were assessed
on sustainability: four for WWTP effluent and five for sludge treatment. The advanced
wastewater treatment methods considered are: (i) granular activated carbon, (ii)
nanofiltration, (iii) solar photo-Fenton, and (iv) ozonation. The sludge treatment
techniques comprise: (i) anaerobic digestion of sludge for agricultural application; (ii)
sludge composting, also for agricultural application; (iii) incineration; (iv) pyrolysis; and
(v) wet air oxidation. They were assessed on sustainability using over 28 indicators, some
of which were also used to evaluate the implication of different treatment techniques for
the energy-water-food (EWF) nexus. Multi-Criteria Decision Analysis (MCDA) was
applied to aggregate the sustainability indicators into an overall sustainability index for
each alternative and identify the most sustainable option(s).
The results suggest that, among the four techniques considered for advanced
effluent treatment, nanofiltration and granular activated carbon have the lowest life cycle
environmental impacts. Although not preferable at all operating ranges, they have the
lowest burdens and are, overall, most sustainable. The latter also has the lowest impact
on the EWF nexus at mean operating parameter, and is the preferred option as the treated
effluent can be used for potable water due lower concerns over the presence of PPCPs.
However, the results also suggest that, from the ecotoxicological point of view, there is
little benefit in using any of the advanced wastewater treatment techniques assessed. This
is due to the life cycle ecotoxicological impacts from the treatment itself being similar or
even higher than for the effluent released into the environment untreated. For sludge
treatments, anaerobic digestion and pyrolysis are environmentally and economically
preferable techniques. The former is the best with respect to the EWF nexus due to the
recovery of energy and agricultural fertilizers. In relation to social aspects, wet air
oxidation is amongst the most desirable for high resource recovery, together with the two
former techniques. The heavy metals content in the sludge applied on agricultural soils is
a major concern for freshwater ecotoxicity potential, posing risks orders of magnitude
higher than PPCP compounds.
18
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institute of learning.
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19
Acknowledgements
To my parents, José Ricardo and Rosa. I thank you for everything in my life and for the
help to accomplish this thesis. I have no words to describe the encouragement in these
last four years.
A special thanks to my supervisor, Adisa Azapagic, for the support and guidance during
these years of academic experience.
I would also like to thank the SIS group for the help and companionship along my PhD,
I'm grateful for all.
Finally, I would like to thank the Conselho Nacional de Desenvolvimento Científico e
Tecnológico (CNPq - Brazil) for the financial support.
20
1. INTRODUCTION
1.1. Background
Pharmaceuticals and Personal Care Products (PPCPs) comprise a diverse group
of substances for human and veterinary use with distinct characteristics and numerous
applications. Their presence in the environment has been receiving growing attention of
the scientific community over the last decades due their potential ecotoxicology and
unknown consumption patterns. The first evidence of these substances in nature was
found during the 1970’s and since then, especially during the past 15 years, several studies
confirmed their presence in many locations worldwide at significant concentrations in the
aquatic environment (Daughton 2016; Daughton 2004; Daughton & Ternes 1999). At
present, most studies on the topic originate from developed nations and to a lesser extent
from developing countries (Liu & Wong 2013; Kolpin et al. 2002; Hughes et al. 2013).
Further studies identified Wastewater Treatment Plants (WWTPs) as the main
pathway for releasing PPCP pollutants to the environment (Heberer et al. 2002; Ratola et
al. 2012). This is due their high concentration in urban effluents and poor degradation
during conventional wastewater treatment, leading to their release with the effluent and
sludge. Since no regulations for these substances exist to date, the amounts of PPCP
compounds originating from WWTPs remain largely unknown (Topp et al. 2008; Ellis
2006; Loos et al. 2013). In addition, as a result of the ageing human population, increasing
urbanization and other factors, it is expected that the consumption of PPCPs will continue
to grow in the future, resulting in increasing discharges of their compounds from WWTPs
(Lyons 2014; WHO 2004).
Although at present the risks posed by the presence of PPCP chemicals in the
environment are considered minor, such claims are often based on premises far from
being representative of field conditions (Fent et al. 2006; Cleuvers 2003; Ortiz de Garcia
et al. 2014). However, it has been found that some of the chemical substances contained
in PPCPs are highly persistent in nature and cause severe toxicological damage to wildlife
(Jobling et al. 1998; Oaks et al. 2004; Kidd et al. 2007). Thereby, their presence in the
environment should be considered taking a precautionary principle to limit their release
and, as a consequence, the risks they pose (Jjemba 2006; Khetan & Collins 2007; Li &
Randak 2009).
21
Stringent environmental regulations, climate change threats, delicate geopolitics
issues, food security and public awareness, among other concerns, demand more rational
approaches related to urban wastewater. Among them, advanced wastewater treatments
could not only improve the quality of effluents discharged to the environment, hence
controlling the presence of emerging contaminants, but also enabling safe reuse of
wastewater (Bixio et al. 2008; Salgot et al. 2006; Tchobanoglous et al. 2011). Similarly,
sludge generated during conventional biological treatment of wastewater also contains
PPCP compounds and, if released to the environment, could pose ecotoxicological and
human risks. These risks could be minimized or avoided altogether if certain methods
were used for their treatment, while at the same time enabling recovery of their nutrient
and energy content (Healy et al. 2008; Stehlík 2009; Fytili & Zabaniotou 2008).
There are many advanced methods for treatment of wastewater and sludge to
remove PPCP compounds. However, at present, there is scant knowledge on which of
these options is most sustainable and should be implemented in practice. This is the topic
of this research, with the aims and objectives outlined below.
1.2. Research aims and objectives
The main aim of this research is to evaluate life cycle sustainability of different
techniques for advanced wastewater and sludge treatment for removal of PPCP
compounds, and identify the most sustainable options considering environmental,
economic and social aspects. The techniques considered for advanced wastewater
treatment are: (i) granular activated carbon, (ii) nanofiltration, (iii) solar photo-Fenton,
and (iv) ozonation. For sludge treatment, the following methods are evaluated: (i)
agricultural application of anaerobic digested sludge, (ii) agricultural application of
composted sludge, (iii) incineration, (iv) pyrolysis, and (v) wet air oxidation. The
application of these technologies is assumed to be in the UK.
The main objectives of the work are:
to evaluate environmental sustainability using life cycle assessment (LCA);
to assess economic sustainability through life cycle costing (LCC);
to explore social issues through social LCA (SLCA);
to consider their impact on the energy-water-food (EWF) nexus;
22
to identify most sustainable options based on the findings of LCA, LCC and SLCA,
using Multi-Criteria Decision Analysis (MCDA) and assuming different preferences
for different sustainability aspects; and
to make recommendations to the wastewater industry, policy makers and consumers.
As far as the author of the study is aware, this is the first study of its kind
internationally. In addition to that, the following specific parts of the study represent
novel contributions to knowledge:
a new methodology for estimating the amount and concentration of PPCP compounds
in WWTPs (see Chapter 4);
estimation of life cycle environmental (including potential ecotoxicological effects of
PPCP compounds released from WWTPs into the environment), economic and social
impacts of different advanced wastewater and sludge treatment techniques (Chapters
5-7);
consideration of the impacts on the EWF nexus of the advanced wastewater and
sludge treatment techniques for the removal of PPCP compounds (Chapter 8); and
identification of most sustainable options considering differing preferences for
sustainability impacts (Chapter 8).
1.3 Structure of the thesis
Following this introduction, Chapter 2 presents a literature review related to the
presence and pathways of PPCP compounds in the environment, ecotoxicological
evaluation methodologies and their physicochemical behavior during conventional
wastewater treatment. The advanced wastewater treatment techniques are also discussed,
together with current European policies on PPCP compounds.
In Chapter 3, the methodology applied for the sustainability assessment is
presented, including the method developed for the estimation of the amount and
concentration of PPCP compounds in WWTPs. The methodologies for LCA, LCC and
SLCA and MCDA are also outlined. An approach developed for the consideration of the
impact on the EWF nexus is also described.
23
The results of the research are presented and discussed in Chapters 4-8. Chapter
4 details the methodology developed for estimating the concentration of PPCP
compounds in WWTPs. The results of environmental sustainability assessment of the
wastewater and sludge treatment techniques are given in Chapters 5 and 6, respectively.
The economic sustainability of both types of method is discussed in Chapter 7. The results
are summarized in Chapter 8 to identify the most sustainable options through MCDA and
determine their effect on the EWF nexus. Finally, the conclusions, recommendations and
future work are provided in Chapter 9.
24
2. LITERATURE REVIEW
This literature review first defines PPCPs by briefly describing the target
compounds investigated in this research. Next, an overview of analytical methods for
detecting and measuring these substances is presented, followed by the current state of
knowledge about their origin, degradation, and ecotoxicological potential in the
environment. Afterwards, a literature search related to the presence and removal of these
compounds using conventional wastewater treatments, the current regulations related to
their industrial production and consumption, and the directives on the control of the
presence of these compounds in the environment are discussed. Lastly, a description of
the chosen treatment techniques is provided, and their role in sustainable development
was conducted.
2.1. PHARMACEUTICALS AND PERSONAL CARE PRODUCTS
Pharmaceuticals and personal care products (PPCPs) are substances used by
humans for personal health and cosmetic reasons or by the agricultural industry to
maintain the health or enhance the growth of livestock. The term was coined by Daughton
& Ternes (1999) and includes thousands of chemical compounds, varying from
prescribed and non-prescribed pharmaceuticals (human and veterinary) to active
ingredients in skin and dental care products, soaps, sun screen agents, fragrances,
cosmetics and many other products. Furthermore, the term also includes their metabolites
and transformation products that are discussed below.
Metabolite refers to molecules resulting from structural changes in the parent
PPCP compound within living organisms. On the other hand, molecules resulting from
structural changes when these compounds reach the environment can undergo biotic and
non-biotic processes forming the so-called “transformation products” (TPs) (Kagle et al.
2009; Daughton 2001; Kümmerer et al. 2000). These molecular changes are described in
Figure 1. Alternatively, PPCPs can remain indefinitely unchanged in both situations
(Deblonde & Hartemann 2013; Ingerslev et al. 2003; Kümmerer 2009b; Farré et al. 2008).
25
Figure 1 - Molecular transformations of parent PPCP compounds in living organisms and the environment
Most PPCP compounds are small nonpolar molecules with molecular weights
(MW) varying between 100 and 1,000 Daltons (Da) and a broad range of physicochemical
properties (Barron et al. 2009; Daughton & Ternes 1999; Ratola et al. 2012). Many of
these substances are high-volume production chemicals, whereas others are produced in
smaller amounts by chemical industries, such as the chemicals found in shampoos and
antibiotics, respectively. In addition, many compounds are often used in more than one
product and in different proportions (Ellis 2006; Kot-Wasik et al. 2007).
Global production of these compounds has increased in the last decades. The
consumption and variety of pharmaceutical compounds, for example, are expected to
expand, especially in developing countries such as China, India, Brazil, and Mexico (IMS
2011). This is mostly due to population ageing, per capita income growth, urbanization,
transformations in disease treatment, and escalation and improvement of health care
systems among other factors related to economic progress (Hill & Chu 2009; Sherer 2006;
WHO 2004). Moreover, the constant development of new chemical compounds has led
to the commercialization of a wider range of these chemicals increasing their presence in
the environment (Daughton 2004; Sarmah et al. 2006).
Metabolite
phase 1
Parent PPCP compound
Metabolite
phase 2
Conjugation with:
Glucuronic acid
Sulphate
Amino acid
Environment
Organism
Ex
cret
ed u
nch
ang
ed
Dir
ect
to t
he
env
iro
nm
ent
26
Substances such as PPCP compounds are included in a broader category named
“emerging contaminants” (ECs) (Petrie et al. 2014; Gavrilescu et al. 2014). The term is
used to designate natural and synthetic substances that have been increasingly detected in
different environments with suspected ecotoxicological effects. Nowadays there are no
major regulations for control of their presence in nature (Farré et al. 2008; Richardson
2009). Concerns and early reports about the presence and harm of these emerging
contaminants to the environment can be traced back to 1962, with the publication of the
book “Silent Spring” by Carson (1962). The book addresses the effects of unregulated
pesticides on bird’s eggs in some regions of the United States of America (USA),
describing other environmental issues in detail. It had considerable repercussions that
culminated in the banishment of most pesticides used at that time.
2.2. PPCP COMPOUNDS IN NATURE
This section presents the current state of knowledge regarding the presence and
ecotoxicological evaluation of PPCP compounds in the environment, with a special focus
on the aquatic environment. Firstly, the most common analytical methods for monitoring
these compounds are presented, followed by the current knowledge about their origin,
fate, and occurrence in different environmental compartments. The most frequent
transport and degradation pathways of these substances in nature are then discussed.
Lastly, it presents a discussion about the current ecotoxicological evaluations describing
the limitations when assessing the risks posed by these compounds to nature.
2.2.1. Sampling and analytical techniques
When evaluating the presence of PPCP compounds in freshwaters, the first step
is determining the sample locations that would be representative of their actual
concentrations in time and space. This is done by defining sampling locations at sites
likely to show greater variability (near urban areas and industries) and considering
variations among seasons, weather conditions and volume flows. This enables a better
balance for appraisal of their occurrence and fate in the studied area (Dębska et al. 2004;
Hilton & Thomas 2003; Kot-Wasik et al. 2007; McArdell et al. 2003). It is recommended,
for the sake of preservation, that samples are kept at low temperatures (2-5°C) and not
exposed to light during transportation to the laboratory, and oftentimes the addition of
reagents for ensuring low reactivity (Fedorova et al. 2014).
27
Once in the laboratory, the separation of suspended matter by centrifugation
produces filtered water and solids, which are analysed separately (Kot-Wasik et al. 2007).
Many methods are available to detect PPCP compounds in liquid and solid phases;
however, due to the great number of compounds, procedures are particularly focused on
the most relevant ones (Bolong et al. 2009; Richardson & Ternes 2005; Snyder et al.
2003; Ternes 2001). Furthermore, occasionally the extraction of the sample is performed
for achieving or enhancing detection. Currently, the most frequent option for this intent
is the solid phase extraction (SPE), notwithstanding many novel techniques are under
research for future utilization (Chenxi et al. 2008; Yu & Wu 2012). The matrix influence
at this stage is usually the impairment of the final results, although magnification could
also occur (Koutsouba et al. 2003; Richardson & Ternes 2005).
The most common analytical techniques for the detection of PPCP compounds
are gas chromatography (GC) and high-performance liquid chromatography (HPLC).
Nowadays the combination of HPLC and mass spectrometry (MS) is becoming the most
ordinary option for this purpose (Dębska et al. 2004; Rodrıguez et al. 2003; Zorita et al.
2008). Another promising technique is the advanced liquid chromatography-tandem mass
spectrometry (LC-MS/MS), an improvement over the traditional GC-MS analysis, in as
much as derivatization is avoided and measurements at lower concentrations can be
reliably accomplished, e.g. limit of detection (LOD) and limit of quantification (LOQ)
(Brooks et al. 2012; Richardson & Ternes 2005; Oliveira et al. 2015).
2.2.2. Presence and sources
Initially detected in surface waters of the USA in the 1970’s and worldwide since
then, chemical substances originated from PPCPs are usually present at low
concentrations in the environment (ng/L) (Jones-Lepp & Stevens 2007). However,
studies have confirmed high concentrations (µg/L) of some of these compounds (Cooper
et al. 2008; Roig 2010; Carmona et al. 2014). Information provided in reports and
scientific articles regarding measurements carried out in several different locations in the
USA and Europe confirmed that many PPCP compounds are expected to be found in
impacted freshwaters in concentrations ranging from 0.001 to 0.01 µg/L in these regions
(Bendz et al. 2005; Boyd et al. 2003; Jones et al. 2002; Lyons 2014; Spongberg & Witter
2008; Zuccato et al. 2000; Kümmerer 2009b; Blair et al. 2013; Park & Park 2015; Hughes
et al. 2013; Kolpin et al. 2002; Huber et al. 2016).
28
More recent studies have demonstrated that a subset of these substances is also
frequently found in ground waters in Germany, France, United Kingdom, and Spain
(Stuart et al. 2012; Vulliet & Cren-Olivé 2011; Jones et al. 2002; López-Serna et al.
2013). Furthermore, in more extreme situations, these compounds have also been
detected at significant concentrations in treated drinking water of some cities, indicating
low efficiency of current traditional drinking water treatment plants (DWTPs) to remove
many of these compounds resulting in their continuous ingestion by human populations
(Benotti et al. 2009; Stackelberg et al. 2004; Ternes et al. 2002; Webb et al. 2003; Xu et
al. 2009; WHO 2011; Azzouz & Ballesteros 2013; Gaffney et al. 2015; Pal et al. 2014).
More recently, the presence of these contaminants has been assessed in Asian
mainland oftentimes showing higher concentrations than those in Europe and North
America. According to some authors, this is mainly due to the lack of proper wastewater
treatment, greater volume of effluent released by urban conurbations as a result of higher
urban density or larger number of inhabitants, and the supposed indiscriminate use of
these substances (Chang et al. 2010; Kim et al. 2009; Kolpin et al. 2002; Lin & Tsai 2009;
Liu & Wong 2013; Minh et al. 2009; Richardson et al. 2005). Similar findings have been
reported in studies conducted in South America, more specifically in Brazil (Ghiselli
2006; Kuster et al. 2009; Stumpf et al. 1999).
Among the several sources that release these substances into the environment, the
most common, though diffuse, are runoff from livestock manure and agricultural
irrigation using wastewater (Kemper 2008; Love et al. 2012; Yu et al. 2013; Siemens et
al. 2008). Additional routes include manufacturing plants, often associated with the
release of high amounts of these substances into their effluents and surroundings (Butters
et al. 2006; Farré et al. 2008; Fick et al. 2009; Larsson et al. 2007). Nevertheless, WWTPs
effluents are considered nowadays to be the major route of PPCP release in the
environment (Celle-Jeanton et al. 2014; Heberer et al. 2002; Roig 2010; Siemens et al.
2008; Michael et al. 2013; Li 2014). To illustrate what will be discussed next, Figure 2
summarizes the main routes for transport of these substances in the environment, from
their manufacturing to the complete degradation of a single PPCP compound.
29
Figure 2 – Main source and transport routes of PPCP compounds during their life cycle
Treated urban effluents originating from households, commercial establishments,
industries, slaughterhouses, and hospitals and other similar urban establishments, if
treated, are directed to sewage networks, and consequently their PPCP compounds reach
WWTPs. If the sewage is not treated, these compounds are directly released into the
environment together with raw sewage thus leading to other issues that are not relevant
to the objectives of the present study. Accordingly, it is well known that the concentration
of PPCPs in urban sewage is highly variable (Ratola et al. 2012; Verlicchi, Al Aukidy &
Zambello 2012; Deblonde et al. 2011). Although consumption data is usually adopted as
a straightforward method to estimate amounts reaching WWTPs, sewage composition
variations, local precipitation, temperature, industrial activities, hygiene habits, to cite a
few, are also important inputs to assess the amount of these chemicals likely to be found
in WWTP influents. These variations and uncertainties will be further discussed in topic
2.3 of this literature review.
Non-biological transformation
Conventional treatment
Advanced treatment
Biological transformation
Manufacturing Prescribed sales
Directed to wastewater treatment plant
Directed to wastewater treatment plant
Landfill
FarmlandRunoff / infiltration/ erosion
Leaching
Release / leaking
Release
Dow
n t
he
dra
in
To t
he
envir
on
men
t
Was
te m
anag
emen
t
Rura
l w
aste
Urb
an e
fflu
ent
Manure / release
Ru
noff
/ l
each
ing
Use
Uptake by biotaFauna / Flora
To landfill
Biosolids
Reu
se
Dir
ect or
indir
ect urb
an r
euse
Food c
onsu
mpti
on
Was
te m
anag
emen
t
Use
Marketing
Non-biological metabolism
Metabolism
Microbial metabolism See
Fig
ure
1
Release
Improper disposal
Wastes
30
After reaching the WWTP, the influent undergoes several stages of treatment to
remove suspended solids (SS), dissolved organic matter (DOM), nutrients, and other
pollutants from sewage. However, many PPCP compounds are poorly removed because
their physicochemical properties are significantly different from those of pollutants that
must be removed from wastewaters (Verlicchi, Al Aukidy & Zambello 2012; Ratola et
al. 2012; Luo et al. 2014). This topic will be further discussed in detail in section 2.3 of
this literature review. Therefore, a substantial fraction of these compounds is likely to be
found not only in the final treated effluent released into freshwaters but also in the sludge
removed from the liquid phase during biological treatments, which is commonly applied
to land after treatment (i.e. biosolids) (Verlicchi & Zambello 2015; Fytili & Zabaniotou
2008). Thus, WWTPs may act simultaneously as the punctual and diffuse source of
contamination of PPCP compounds in the environment.
While the presence of PPCP compounds in freshwaters is most often associated
with WWTP effluent releases, their presence in overland soils and groundwater is mainly
due to the agricultural use of sludge, irrigation with reclaimed wastewater, leachate from
manure, and poor sanitation services (Kemper 2008; Mantovi et al. 2005; Sarmah et al.
2006; Siemens et al. 2008; Topp et al. 2008; Sorensen et al. 2015). The presence of these
compounds in agricultural soils could lead to their uptake by plants and animals, potential
accumulation through the food chain, and ingestion by humans through food (Jensen et
al. 2001; Karnjanapiboonwong et al. 2011; Love et al. 2012; Sablayrolles et al. 2010; Wu
et al. 2013; Zenker et al. 2014).
2.2.3. Transportation and degradation
Since WWTPs are the major point sources for release of PPCP compounds into
the environment and water (their main destination), their compounds’ sorption properties
constitute an important factor affecting the transport and (bio)availability of
pharmaceuticals in aquatic environments (Li 2014; Heberer 2002). While hydrophilic
compounds are likely to be readily found in freshwaters (due to their tendency to resist
biological treatments, discussed later in section 2.3), antibiotics are known for their
tendency to bind to soil particles or to form complexes with different ions (Bowman et
al. 2002; Kibbey et al. 2007; Kummerer 2003). This behaviour is especially affected by
the amount and nature of suspended matter in the aquatic compartment, and the formation
of complexes may cause the loss in their detectability in collected samples devised to
laboratorial analysis (as discussed in topic 2.2.1).
31
With regard to terrestrial compartments, the sorption behaviour of PPCP
compounds is known to be particularly complex in soils and sediments (Barron et al.
2009; Dodgen et al. 2014; Stevens-Garmon et al. 2011; Tolls 2001). Nevertheless,
psychiatric drugs, antiseptics and hormones seem to have low sorption potential and are
therefore more likely to infiltrate or leach into surface waters (e.g. runoff), while others
seems to be easily degraded or prone to remain sorbed (Y. Fang et al. 2012;
Karnjanapiboonwong et al. 2010; Lapen et al. 2008; Katz et al. 2013; Yu et al. 2013).
Furthermore, the presence of organic matter (OM) and media pH may have significant
influence on the sorption behaviour of many compounds (Calisto & Esteves 2012;
Karnjanapiboonwong et al. 2011; Katz et al. 2013; Pan et al. 2009).
After entering the environment, each PPCP compound is (bio) degraded according
to its physicochemical properties and environmental conditions (Farré et al. 2008; Khetan
& Collins 2007; Kümmerer 2009b). In terms of biological degradation, bacteria and fungi
are the most likely to biodegrade organic compounds, but fungi are considered
uncommon in aquatic environments (Kagle et al. 2009; Kümmerer 2009b; Kümmerer et
al. 2000). An effective biodegradation of some PPCP compounds in freshwaters may
require adaptation of the microbial community, whose previous presence / addition to the
media enables faster degradation rates. Moreover, incomplete biodegradation frequently
results in the generation of TPs (Kagle et al. 2009; Kümmerer 2009a; Kümmerer 2004).
Photolysis and temperature can play key roles in the degradation of many PPCPs
in surface waters, especially for compounds that are more difficult to remove during
biological wastewater treatment. Photolysis is strongly dependent on latitude. Regions
with higher solar irradiation are expected to greatly assist PPCP overall degradation
(Aranami & Readman 2007; Fono et al. 2006; Nikolaou et al. 2007; Packer et al. 2003).
On the other hand, significant reduction of PPCP degradation rates has been observed in
Finnish freshwaters (Finland) during the winter (Vieno et al. 2005), as well as in seasonal
climate transitions (although possibly heavily dependent on variations in consumption
and rainfall amounts) (Papageorgiou et al. 2016). A general scheme summarizing the
transport and the importance of degradation mechanisms of PPCP compounds released
by WWTPs is shown in Figure 3.
32
Figure 3 - Main transport and degradation mechanisms of PPCP compounds in the environment
2.2.4. Ecotoxicological evaluations
The primary purpose of ecotoxicological evaluations is to predict potential effects
in the biota of certain substance (and its stressors) released to the environment (Fent et al.
2006; Li & Randak 2009; Roig 2010), and they usually include: (i) exposure and (ii)
hazard assessments. Due to animal welfare and other reasons, most studies concerning
PPCP exposure are focused on acute tests although effects of PPCP compounds are
expected to show greater risks of chronic exposure. The criteria for effects or endpoints
(i.e. hazard) include, among others, the: lowest observable effect concentration (LOEC),
no observable effect concentration (NOEC), median effective concentration (EC50), and
median lethal concentration (LC50).
The general procedure in ecotoxicological evaluations (e.g. environmental risk
assessments - ERAs) of a substance usually begins with the use of its predicted no-effect
concentration (PNEC) to estimate its potential adverse effects. It takes in account
exposure and hazard assessments simultaneously to determine tolerable levels for the
presence of a substance in the environment. The PNEC calculation is based on two main
assumptions: (i) the overall vulnerability of the ecosystem is dependent on its most
sensitive species; and (ii) shielding the ecosystem guarantees the correct function of the
environment (Joint Research Centre 2003).
Parent PPCP compound
Surface water
+ acclimation
+ temperature
+ latitude
- biodegradation
- hydrolysis
- sorption
Groundwater
+ sorption
- biodegradation
- photolysis
- temperature
- oxidation
Soil /
Sediment+ sorption
+ pH
+ photolysis
+ organic matter
- hydrolysis
Infiltration
Runoff Upwelling
Interactions
Wastewater treatment plant
Eff
luen
t(t
o s
oil
s)
33
The PNECs are calculated either in a deterministic or probabilistic manner (Joint
Research Centre 2003). In the former, confidence values are given through assessment
factors (AF) to account for uncertainties related to species sensitivity, species trophic
level, compartment, exposure period, and test type (Hickey 2010; Joint Research Centre
2003). In the latter, statistical methods define accepted protection levels by experts or
organizations and it can be performed using statistical extrapolation (see Figure 4 for a
general descriptive example) (Roman et al. 1999). Important to note that more specific
guidelines for ecotoxicological tests has been specifically created for pharmaceutical
substances developed after 2006 (Länge & Dietrich 2002; Grung et al. 2008; Roig 2010).
Figure 4 – Representation of standardized PNEC values (adapted from Roman et al. 1999)
Mixture effects (i.e. synergism) are an important topic of concern associated with
the presence of PPCP compounds in the environment. The joint action of several different
compounds is critical when assessing the potential harms of these substances since
thousands of them act simultaneously in nature at variable concentrations. However there
are few ecotoxicological studies available addressing synergetic effects (Claessens et al.
2013; Cleuvers 2003; Cleuvers 2004; Kümmerer 2009a). Similarly, the effects of their
metabolites and transformation products have rarely been considered in the literature,
which is also a matter of great concern (Dann & Hontela 2011; Farré et al. 2008).
\\\\\\\\\\\\\\\\\\\\\\\\\
Per
cen
t sp
ecie
s
PNEC methodologies
Mean NOEC
+1 / -1 Standard deviation (SD)
NOEC histogram
34
Another issue, associated solely with antibiotics and antiseptics, is microbial
resistance. The prevalence of microorganisms with acquired resistance to the action of
these substances poses a continuous and increasing threat to all living organisms on the
planet and thus to the environment (Dann & Hontela 2011; Kümmerer 2004; Yazdankhah
et al. 2006; Kostich et al. 2014; Michael et al. 2013; Rizzo et al. 2013). Accordingly, it
also represents a direct risk to human health since resistant microorganisms are associated
with higher rate of infections in hospitals, among other inconvenient matters.
Nonetheless, only recently they started to receive the attention of the scientific
community (Bound & Voulvoulis 2004; Ohlsen et al. 2003; Schwartz et al. 2003;
Spellberg et al. 2008).
Given the aforementioned concerns and other inherent uncertainty matters, a more
realistic scenario for assessments should claim continuous exposure to a multitude of
stressors and holistic appraisals (Nash et al. 2004; Escher et al. 2008; Sanderson et al.
2003). One approach is the toxicant totality tolerance trajectory (4Ts) of exposure, a term
that intends to embrace the complete context of an organism exposure to chemical
stressors in a particular environment during their entire life cycle (Daughton 2004).
Another approach is ecotoxigenomics, which aims to include gene expression profiles in
target microorganisms and create scenarios for better evaluation of a pollutant in living
organisms (Poynton & Vulpe 2009; Fedorenkova et al. 2010; Snape et al. 2004). Recent
studies and regulations have been developing a list of priority PPCP compounds that pose
greater risks to the environment, but many obstacles have yet to be removed to overcome
uncertainties and deal with the issue in a straightforward manner (Roos et al. 2012; Roig
2010).
Despite hesitation involving harmful effects of PPCP compounds, some studies
have confirmed their undesired effects in regions around the globe. For instance, the steep
decline in vultures (Gyps bengalensis) in the Indian subcontinent was directly linked to
diclofenac intake present in the carcasses of livestock fed with high amounts of this
substance, leading to an abnormal death rate of these animals due to acute liver
intoxication (Oaks et al. 2004; Risebrough 2004). Similarly, in Canada and the United
Kingdom, a decline in a controlled fish populations (Pimephales promelas and Rutilus
rutilus) was directly connected to the chronic exposure to hormones that ultimately
induced an abnormal fish reproduction, driving population size downward to extinction
(Jobling et al. 2002; Kidd et al. 2007).
35
In other words, nowadays the risk posed by the presence of PPCP compounds in
the environment seems to concern environmental hygiene rather than ecotoxicology
(Boxall 2004; Carlsson et al. 2006; Jones-Lepp & Stevens 2007; Kümmerer 2009a).
Besides, effects and side effects of lifelong intake, consequences over subpopulations,
synergistic effects, among other sensitive issues, have not still been properly scrutinized
and well established in regulations and legislations (Blasco & Delvalls 2008; Fent et al.
2006). Therefore, the presence of PPCP in the environment should be considered, if
nothing else, deleterious (Khetan & Collins 2007; Jjemba 2006; Gavrilescu et al. 2014).
2.3. PPCP COMPOUNDS IN WASTEWATER TREATMENT PLANTS
Urban wastewaters and their management say a lot about society customs,
including nutrition, sexual habits, and attitudes towards fashion trends. Moreover,
evolution of human societies through the ages further enlightens civilizations’ pursuit of
technological, economic, and social advances. Unfortunately, nowadays many
developing nations still lack access to adequate sanitation services, and several important
issues arise from this fact (WHO/UNICEF 2015). In this context, the insightful paper by
Lofrano & Brown (2010) is an excellent source regarding wastewater management, from
earliest human communities to modern development trends.
Efficient methods for wastewater treatment were initiated around 1850 in the
margins of the river Thames in London, United Kingdom. At that time, the so-called
“sewage farms” were established to receive urban wastewaters to contain cholera
outbreaks that plagued the city. Back then, treatment plants performed only basic
biological treatment before discharging their effluents into the river, partially removing
the pollutants. From then onwards, wastewater treatment continuously evolved to its
current technological status (Apedaile 2001; Shannon et al. 2008). The next sections
provide relevant information related to the presence and removal of PPCP compounds in
contemporary WWTPs.
36
2.3.1. Conventional wastewater treatment methods
Wastewater treatment plants applying solely biological (e.g. secondary) treatment
after preliminary (or primary) effluent treatment and using typical sludge conditioning
methods are called “conventional treatment plants”. Primary treatment involves
clarification, sedimentation, and settling of raw sewage. Secondary treatment includes
three main categories: activated sludge (AS), membrane bioreactor (MBR), and up flow
anaerobic sludge blanket digestion (UASB) (Wang et al. 2009; Sperling 2007). With
respect to the sludge generated during the wastewater treatment, there are many different
conditioning methods currently available (L. Wang et al. 2008; Kelessidis & Stasinakis
2012). However, thickening, stabilization, and dehydration or dewatering are the most
frequently adopted methods in Europe (Fytili & Zabaniotou 2008).
To better describe the behaviour of PPCP compounds during wastewater
treatment, it is necessary to characterize the solid matter content in the sewage sludge
since it plays a major role in the design and operation of the treatment itself (Jordão &
Pessôa 1995; Sperling 2007). Figure 5 depicts a balance flowchart for sewage sludge’s
solid matter, and the measurement methods employed in this task are as follows:
• Particle dimensions: suspended solids; colloidal solids, dissolved solids;
• Settleability: settable solids; floating solids, non-settable solids;
• High drying temperature (550-600 ˚C): fixed solids, volatile solids; and
• Average drying temperatures (103-105 ˚C): total solids, total suspended solids,
and total dissolved solids.
Figure 5 – Typical solid matter content in sewage sludge (adapted from Jordão & Pessôa 1995)
Total solids
100%
Suspended solids
Settable solids
60%
Dissolved solids
40%
Volatile solids
50%
Fixed solids
10%
Volatile solids
20%
Fixed solids
20%
Volatile solids
70%
Total solids
100%
Fixed solids
30%
37
2.3.2. Concentration of PPCP compounds in influents
The concentration of several PPCP compounds in influents in many Western
European WWTPs has been reported in the literature; many articles show measurement
results for specific plants, and few reviews summarize their findings. Examples of these
results can be found in the studies by Deblonde et al. (2011), Ratola et al. (2012) and
Verlicchi et al. (2012). These authors suggest that most these compounds should not be
considered ubiquitous in urban effluents although many can be frequently found;
therefore, they have been referred to as ‘‘pseudo-persistent” pollutants (e.g. the
compounds targeted in the present study).
Insofar as WWTPs are the major point sources of PPCP contaminants release into
the environment, the contribution of the corresponding served populations in terms of
prescribed/sold/consumed PPCP amounts is a variable of utmost importance (Roig 2010;
Ortiz de García et al. 2013). Another one is hospital effluent discharges; however, their
relative contribution to the final PPCP concentration at the municipal scale wastewater is
generally so small that they can be considered negligible, except for a minority of
compounds and specific scenarios in which their influence appears to be significant
(Chang et al. 2010; Kosma et al. 2010; Langford & Thomas 2009; Ort et al. 2010;
Schuster et al. 2008; Sim et al. 2010; Frédéric & Yves 2014; Verlicchi, Al Aukidy,
Galletti, et al. 2012).
The interpretation of the above-mentioned data can be open to doubt due to its
complex and frequently unknown dependence on people’s hygienic habits, cultural
aspects, economic situation, climate conditions, and other variables that directly affect
PPCP compounds concentration in WWTP influents (Alexy et al. 2006; Carballa, Omil,
et al. 2008; Oosterhuis et al. 2013; Zhang & Geißen 2010; Kosma et al. 2014). In spite of
that, more than often this is the predominant approach to estimate the concentration of
these chemicals reaching WWTPs, notwithstanding the widely recognized scarcity of
data and large data scattering among districts and even neighbouring cities (Boyd et al.
2003; Diener et al. 2008; Ellis 2006; Göbel et al. 2005; Li & Zhang 2011; Zhou et al.
2009; Roig 2010). Nonetheless, some studies adopted this approach and reasonably
accurate results were found for some PPCP compounds (Coetsier et al. 2009; Ortiz de
García et al. 2013; Khan & Ongerth 2004; Oosterhuis et al. 2013; Roig 2010).
38
2.3.3. Removal of PPCP compounds by conventional effluent treatments
In this section, the term “removal” denotes that the compound of interest is no
longer detectable in the compartment of analysis or sampling. Therefore, the compound
could have switched compartments or could have been partially degraded or mineralized.
The reason is that very few studies have addressed these topics simultaneously, indicating
a knowledge gap that requires a closer look by the scientific community (Kosma et al.
2014; Evgenidou et al. 2014). For example, the identification and quantification of the
ratio between PPCP degradation rate and compartment change from parameters such as
dissolved organic carbon (DOC) can aid in the extension of PPCP conversion into
inorganic salts (Quintana et al. 2005; Weigel et al. 2004; Petrie et al. 2014).
2.3.3.1. Removal in primary treatments
Studies regarding primary treatments and removal of PPCP compounds have
suggested that this stage is unable to efficiently remove most of these chemicals since the
primary objective of this stage is to withdraw large solid particles. It has been estimated
that traditional primary treatments can account for the removal of up to 15% of PPCP
compounds if compared to secondary treatments (Suárez et al. 2008; Lee et al. 2009).
However, during the primary stage, PPCP extraction efficiency can be improved through
coagulation-flocculation techniques, which could lead to removal rates as high as 60%
for some compounds (Carballa, 2005, Carballa, Omil, et al., 2005).
2.3.3.2. Removal in secondary treatments
As previously stated, several studies have suggested that conventional WWTPs
(i.e. primary treatment followed by secondary treatment) are inefficient in removing
many PPCP compounds. The reason is that they are typically designed for removing
pollutants such as organic matter, nutrients, and microorganisms, which are non-polar
compounds that are easily biodegradable and are large enough to be removed, using this
type of treatment. Therefore, since most PPCP compounds are small molecular weight
(MW) compounds with a polar tendency, their elimination is, indeed, hindered (Ratola et
al. 2012; Verlicchi, Al Aukidy & Zambello 2012).
39
Likewise, in the natural environment, the removal of PPCP compounds during
activated sludge treatments (AS) occurs according to the compounds’ physicochemical
properties and the surrounding conditions. However, degradation or mineralization by
photolysis, hydrolysis, and air stripping are known to be of less relevance, and
microbiological degradation (i.e. biodegradation) is considered as the main mechanism
for extracting most of these compounds (Onesios et al. 2009; Suárez et al. 2008). The
biodegradation of PPCP compounds encompasses: (i) the use of organic compound as an
energy source (catabolism) and (ii) coincidental transformation of the compound without
use as an energy source (cometabolism). The outcomes of these complex activities are
transformations towards complete degradation and formation of various TPs or even
minor chemical modifications, which vary according to operating parameters and many
other factors (Kagle et al. 2009; A. Y. C. Lin et al. 2009; Suárez et al. 2008; Collado et
al. 2012).
Operating parameters mainly influencing the (bio)degradation of PPCP
compounds during AS treatments are hydraulic retention time (HRT) and sludge retention
time (SRT) (Jelic et al. 2011; Joss et al. 2005; Koh et al. 2008; Verlicchi, Al Aukidy &
Zambello 2012; Stasinakis et al. 2007). With regard to the HRT, Reif et al. (2008)
suggested that variations in HRT in the range commonly used in WWTPs have little effect
on the removal of most PPCP compounds they evaluated. On the other hand, increased
SRT ( ≥ 10 days) has been associated with the improved biological degradation of PPCPs
since there is more time for the growth of bacteria and other microbes (i.e. microbial
acclimation). However, SRT ≥ 25 days seems to not affect the removal of these
compounds (Clara, Kreuzinger, et al. 2005; Suarez et al. 2010; Batt et al. 2007; Blair et
al. 2015; Roig 2010).
Microbial acclimation consists of many processes including genetic processes and
population diversity of microorganisms. Even if the appropriate organisms are present in
sufficient numbers, in many instances the genes necessary are not constitutive for proper
growth due to the low concentration of PPCP compounds. In this case, to enforce the
entry of convenient microorganisms may be a good alternative (Kagle et al. 2009; Clara,
Kreuzinger, et al. 2005). Moreover, if molecules with minor transformations are kept
embedded in the reactor, they can act as a reservoir and be occasionally released as the
original compound by cleavage (Gao et al. 2012; Li & Zhang 2011; Blair et al. 2015).
Furthermore, the transformation of one compound into another could also occur, such as
the case of the hormone oestrone, which under oxidizing conditions can be converted into
40
its more potent for, 17-oestradiol (Carballa et al. 2004; Schlüsener & Bester 2008;
Grover et al. 2011; Xu et al. 2012; Esperanza et al. 2007; Baronti et al. 2000).
As discussed above, the extent of PPCP compounds biodegradation during AS
can vary greatly. A rough indicator of biodegradation rate is calculated based on the SS
concentration (solid content retained when the solution is filtered through a fibreglass
filter with 1.2-millimeter pores), described as a pseudo-first order reaction (Joss et al.
2006):
dCeff,i
dt= kbiol,i XSS Cinf,i (1)
Where:
Ceff,i concentration of a compound in the effluent (µg/L)
Cinf,i concentration of a compound "i" in the influent (µg/L)
t time (d)
XSS suspended solids concentration (gss/L)
kbiol,i biological rate degradation constant of a compound (L/gss d)
Relying solely on kbiol, the removal of PPCP compounds during AS treatments
has been classified by some authors as: low removal (kbiol < 0.10 L/gss × d, with < 20%
predicted removal), moderate removal (0.10 L/gss × d < kbiol < 10 L/g × d, with 20 - 90%
predicted removal), and high removal (kbiol > 10 L/gss × d, with > 90% predicted removal)
(Suárez et al. 2008; Joss et al. 2005; Joss et al. 2006).
An alternative to increase and/or more effectively control the SRT during AS
treatments is the use of membrane filtration systems (MBR). In an MBR, the use of a
microfiltration or ultrafiltration membrane in the AS bioreactor enhances SS retention by
increasing their concentration during the secondary treatment, in addition to filtrating
many micro-pollutants and pathogens (Sipma et al. 2010; Urase et al. 2005).
Consequently, this system configuration increases the overall treatment efficiency,
decreases sludge production in comparison to that of AS treatment, and it enables
wastewater reuse due to better overall effluent quality (Wintgens et al. 2005).
41
It has been suggested that MBR can promote extra biological transformation and
thus greater degradation of PPCP compounds. However, most studies comparing AS and
MBR treatments have concluded that the SRT itself and other factors are actually more
important (Bernhard et al. 2006; Cases et al. 2011; Fernandez-Fontaina et al. 2013;
Tadkaew et al. 2011; Hai et al. 2011; Petrovic et al. 2009; Sipma et al. 2010; Kimura et
al. 2005; Tambosi et al. 2010; Reif et al. 2008; Clara, Strenn, et al. 2005). Nevertheless,
as previously stated effluents from MBRs have better quality compared to those of AS
(e.g. less organic matter, suspended solids, and pathogens content) and are oftentimes
compatible to freshwaters. Therefore, they are also more suitable for advanced
wastewater treatment techniques (Hai et al. 2014; Wang et al. 2009), which will be further
discussed in section 2.5.3.1. of this literature review.
More recent studies have evaluated the removal of PPCP compounds using UASB
treatments. Comparing AS and UASB treatments, it has been suggested that, in general,
antibiotics are more easily removed under anaerobic conditions, while other fourteen
compounds had higher biological degradation rates during AS treatments. Moreover, it
has been found that during UASB treatments some PPCP eradication was positively
correlated with methane generation, whereas during AS treatments, it has been associated
with nitrifying conditions (Alvarino et al. 2014; Suarez et al. 2010).
2.3.4. Concentration of PPCP compounds in sludge
The PPCP compounds sorbed in the solid phase during primary treatments and
secondary treatments (i.e. mixed sludge) ire separated from the aqueous phase and
directed to further unit operations. Their composition is strongly dependent on the
influent composition, reactor type, SRT, HRT, and other operating parameters (Wang et
al. 2005; Rulkens 2007). Nevertheless, the volume of sludge (Vsludge) produced during
conventional wastewater treatment can be approximately estimated using its fixed solids
content, usually on a dry matter basis (DM), as reported by Jordão & Pessôa (1995):
Vsludge =100
100−DMx
Msludge
σ (2)
Where:
Vsludge sludge volume (m3)
Msludge dry solids mass (kg)
42
DM dry matter content (%)
σ water density (kg/m3)
A number of physical and pH-dependent mechanisms influence the sorption
potential of PPCP compounds onto sludge, hampering a reliable estimation (Barron et al.
2009; Carballa, Fink, et al. 2008; Carballa et al. 2005; Hörsing et al. 2011; Verlicchi &
Zambello 2015). Humic substances, for instance, may alter the surface properties of solid
particles resulting in inconsistent sorption behaviour (Alvarino et al. 2014; Jones-Lepp &
Stevens 2007; Kümmerer 2009b).
Many PPCP compounds are expected to be found in the sludge supernatant, being
recycled as inlet during sludge thickening. This is especially the case of some hormones
that have high sorption potential at low pH values but high desorption at high pH values
(Carballa et al. 2004; Andersen et al. 2005). This possibility should be considered
separately for each specific treatment since the supernatant recycled back to influent
treatment line may contain relevant loads of these compounds affecting mass balances
(Clara, Kreuzinger, et al. 2005; Sim et al. 2011).
Furthermore, the total amount of PPCP compounds in the sludge also depends on
their influent concentration. Nonetheless, the high lipid content in the sludge allows
inferring that less polar compounds with higher sorption potential properties are likely to
have higher concentrations (Carballa, Fink, et al. 2008; Joss et al. 2005; Ternes, Joss, et
al. 2004). Thus, a suitable approach to a primary assessment of the concentration of these
substances in the sludge is to determine the solid-water distribution coefficient (Kd), an
indication of the compound’s affinity to sludge solids, considering the adsorption and
absorption processes simultaneously (Carballa, Fink, et al. 2008; Göbel et al. 2005;
Stevens-Garmon et al. 2011; Ternes, Herrmann, et al. 2004). This is described in Jones et
al. (2002) and Verlicchi & Zambello (2015) as:
Kd,i = Csorbed,/ Csoluble,i (3)
Csludge,i =Minf,i
(VWWTP Kd,i⁄ )+Msludge (4)
Where:
Kd,i solid–water distribution coefficient of a compound (L/kgss)
43
Csorbed,i concentration in sludge of a compound in suspended solids (kgi/kgss)
Csoluble,i concentration in the aqueous phase of a compound (kgi/L)
Csludge,i concentration of a compound "i" in the sludge (kgi/kgss)
Minf,i discharge of a compound "i" in the wastewater treatment plant (kg)
VWWTP influent volume of wastewater (L)
Msludge dry solids mass of sludge produced during treatment (kg)
2.3.5. Removal of PPCP compounds by conventional sludge treatments
Conventional sludge treatment refers to methods for volume reduction and
stabilization of the mixed sludge. In most developed countries, the treatments used for
this intent are mainly physicochemical process, such as chemical conditioning and
mechanical drying/dewatering, in combination with stabilization methods such as
digestion and composting, after gravity thickening (Wang et al. 2005; Verlicchi &
Zambello 2015). The degree and final combination in which these processes are applied
to the sludge varies from the extension of the required final sludge volume,
physicochemical characteristics, final pathogens and metal content, as well as the size of
the WWTP. The final destination of the treated sludge are usually land application (forests
or agriculture) (Kelessidis & Stasinakis 2012).
The influence of such treatments in the removal of PPCP compounds is unknown
since, so far, studies addressing this topic have reported inconstant removal results due to
the number of variables involved (especially Kd, sludge retention time, sludge
composition and treatment temperature). Furthermore, often the untreated liquid phase is
returned to wastewater treatment line, which difficult the mass balance of these
substances during wastewater treatment (especially true in the case of hormones)
(Carballa, Fink, et al. 2008; Tunçal et al. 2011; Carballa et al. 2007; Blair et al. 2015;
Suárez et al. 2008). Nevertheless, anaerobic digestion seems the only process removing
considerably these substances in comparison to other conventional methods (Verlicchi &
Zambello 2015).
44
2.3.6. Summary of literature
The objective of this literature review is to demonstrate that conventional WWTPs
are likely to have variations in influent loads and in the behaviour of PPCP compounds
and describe the difficulties faced when making estimates. Therefore, it can be said that
a considerable number of the more than 3,000 PPCP compounds currently marketed are
very poorly degraded during conventional wastewater treatments. The main topics that
corroborate this assertion are summarized graphically in Figure 6.
Figure 6 – Main removal process of PPCP compounds during conventional wastewater treatment plants
Additionally, taking into account the aspects supposed to have the most influence
on the presence of these compounds in the effluent and sludge of conventional wastewater
treatment plants, kbiol and Kd, a rough degradation estimation was defined by Suárez et al.
(2008) demonstrating the following:
• High kbiol / low Kd : well removed independently of SRT and HRT;
• Low kbiol / high Kd: efficiently removed at long enough SRT;
• High kbiol / medium Kd: moderately removed slightly dependent on SRT; and
• Low kbiol / low Kd: not well removed nor biodegraded regardless of the SRT.
Degradation products
Pa
ren
t co
mp
ou
nd
co
nce
ntr
ati
on
in
th
e co
mp
art
men
t
Ceffluent,iDegradation products
Sorbed (~Kd,i )
Transformation products
Conversion
Mineralization products
Infl
uen
t tr
eatm
ent
rem
oval
cap
aci
ty
(plant operation, reactor design, temperature …)
(thickening, dewatering, stabilization…)
No
t re
mo
ved
Parental compound
Degradation products
Non sorbed
Su
per
nata
nt
Transformation products
ineralization products
Parental compound
Rem
oved
Non sorbed
Conversion
Eff
luen
tS
lud
ge
Stripping
Biological impairment
iological impairment
Stripping
Sludge retention time
Acclimation
Transformation products
Desorption
45
2.4. EUROPEAN REGULATIONS RELATED TO PPCP COMPOUNDS
Until the beginning of 1990's, persistent organic pollutants (POPs) and heavy
metals were the focus of environmental monitoring programs in Europe. These
substances were subjected to strict regulations and control measures which successfully
reduced their emissions and consequently the environmental concerns related to them.
Since then, attention has been directed towards potential environmental impacts caused
by new classes of the so-called ECs (emerging contaminants), such as those originated
from PPCPs, surfactants, plasticizers, endocrine disrupting, and others (Petrovic et al.
2004).
Nowadays, only few countries have a regulatory framework for the presence of
PPCP compounds, although directives adjusting the manufacturing and production of
these substances, more specifically pharmaceuticals, have been adopted in Europe since
2004. Closer monitoring of the presence of PPCP compounds in WWTPs has also been
considered in some European and North American regions, especially those facing
freshwater shortage. Furthermore, the monitoring of PPCP compounds is expected to take
place soon, mainly due to the increasingly application of sewage sludge in agricultural
practices. These topics are expanded next.
2.4.1. Production, consumption and disposal of pharmaceuticals
The production of pharmaceutical ingredients involves several multistage
processes. It frequently generates small quantities of the desired final product and large
amounts of associated waste. As for the generated waste, different manufacturing
methods for more economically interesting and environmentally friendly pharmaceutical
goods are currently being researched (Butters et al. 2006; Kampa et al. 2008; Khetan &
Collins 2007; Liu & Wong 2013; Castensson et al. 2009; Cardoso et al. 2014). On the
other hand, the Directives 2004/27/EC and 2001/82/EC set out requirements for
collection of pharmaceutical products, preventing improper disposal and unexpected
sources of contamination (European Parliament 2004; European Parliament 2000).
However, the efficiency of these strategies for reducing the presence of PPCP compounds
in the environment is still unknown at the moment (Roig 2010; Lubick 2010).
46
Due to safety reasons related to increasing bacterial resistance (Yazdankhah et al.
2006; Sarmah et al. 2006; Dann & Hontela 2011), the use of antibiotic as growth
promoters in the agricultural sector was banned by the European Union in 2006. In that
same year, the European Medicines Agency (EMEA) issued guidelines on
pharmaceutical compounds regulating the launch of new substances in order to verify
their environmental safety (Kemper 2008; Roig 2010; Grung et al. 2008). More precise
attempts to assess and classify the environmental safety of pharmaceutical products
during their life cycle are currently being studied to promote the expansion of regulatory
practices aiming to penalize industries for the environmental impacts of their products
(Khetan & Collins 2007; Roig 2010).
2.4.2. Presence in water, wastewater and sludge
Currently, substances considered to pose significant risks to or through aquatic
environments in Europe (water framework directives 2000/60/EC and 2008/105/EC)
include mainly pesticides and toxic metals. However, recent amendments, such as article
8b and 8c of directive 2013/39/EU, have initiated regulations on pharmaceuticals
(diclofenac, 17β-oestradiol, and 17-ethinylestradiol) at more specific conditions for
possible further inclusion in the aforementioned list (European Parliament 2000;
European Parliament 2013; European Parliament 2008).
As for urban wastewater treatment in Europe (directive 91/271/EEC), efforts have
been made towards full compliance with wastewater collection requirements and more
stringent standards concerning its conventional treatment in new European Union
member states (EEC Council Directive 1991; European Commission 2013).
Nevertheless, as the European Union support the precautionary principle, discussions
have focused on some pharmaceuticals as candidates for further monitoring and control
in wastewaters, resulting from the previously mentioned directive 2013/39/EU (Kampa
et al. 2008).
Regarding the sewage sludge, the European Union has highlighted the benefits of
its use in agricultural practices due to promotion of nutrient recycling. Many directives
balancing positive and negative effects have been adopted (directive 86/78/EEC).
Nonetheless, public opinion and stakeholders’ resistance are still a barrier in some regions
(e.g. heavy metals and pathogens content concerns), and WWTPs have faced significant
resistance to market sewage sludge as fertilizers (Iranpour et al. 2004; Fytili & Zabaniotou
2008; Milieu et al. 2010; European Commission 1986).
47
Furthermore, more recently concerns related to the presence of PPCP compounds
have been added and future regulations on agricultural application, derived from article
8c of directive 2013/39/EU, are likely to be even more restrictive for this practice. Actions
in this direction have already been carried out in some European regions (e.g. southern
Germany) setting limit values for some of these compounds (Roig 2010; Jones-Lepp &
Stevens 2007; Martín et al. 2012; Dann & Hontela 2011).
2.4.3. Environmental risk assessment
A primary assessment to estimate the concentration of PPCP compounds in
surface waters is through the calculation of predicted environmental concentrations
(PECs), assuming spatially and temporally evenly distributed usage of the target
compounds (presuming the absence of metabolism or degradation products) and national
prescription/sales/consumption data. However, estimation of PECs involves the same
difficulties as those encountered in estimating the concentration of these compounds in
WWTPs influents, in addition to the challenges regarding their removal potential (see
topic 2.3.3) and dilution in freshwaters.
Comparisons between non-refined PECs values and actual measured
concentrations (MECs) indicate that predicted values are often overestimated, but
underestimations might occur (as in the case of hormones) (Boxall et al. 2014; Celle-
Jeanton et al. 2014; Liebig et al. 2006; Ortiz de García et al. 2013). Nonetheless, PECs
and PNECs (see topic 2.2.4) are the main requirements for providing a first assessment
of potential risks pose by medical substances in Europe (i.e. ERAs), assessed through risk
quotient, the ratio between PEC and PNEC (Länge & Dietrich 2002; Roig 2010).
There are many studies conducted in North America (Atkinson et al. 2012; Cooper
et al. 2008), Europe (Gros et al. 2010; Leung et al. 2012; Sebastine & Wakeman 2003;
Stuart et al. 2012; Andersen et al. 2005; Tauxe-Wuersch et al. 2005), and Asia (T. H.
Fang et al. 2012; Wang et al. 2010) regarding ERAs of freshwater and groundwater in the
surroundings of WWTPs. Most of them have indicate that compounds such as ibuprofen,
sulfamethoxazole, carbamazepine, and 17β-oestradiol potentially pose a considerable risk
(i.e. risk quotient greater than 1).
48
Regarding terrestrial environments, studies carried out in Spain after the
application of sewage sludge to agricultural soils have shown that these same compounds
and gemfibrozil can potentially pose risks if application rates exceed the limit defined in
the guidelines. Moreover, the risks can be reduced if the sludge is pre-treated at higher
temperatures for stabilization (Martín et al. 2012; González et al. 2012). Thus, it can be
argued that previous attempts of prioritization of PPCP compounds indicated that
analgesics/anti-inflammatory drugs, antibiotics, psychiatric drugs, and hormones are
substance groups often posing the highest risks to the environment (Verlicchi & Zambello
2015; Cooper et al. 2008; Roos et al. 2012).
2.5. ADVANCED WASTEWATER AND SLUDGE TREATMENT TECHNIQUES
As outlined in the previous section, although initiatives exist to mitigate
freshwater and soil pollution originated from WWTPs, they do not significantly
contribute to the reduction of the risks associated with the presence of PPCP compounds
in the environment. Thus, the adoption of so-called advanced wastewater treatment
techniques is necessary to effectively diminish the presence of these environmental
pollutants. The term “advanced” refers to processes capable of significantly improve the
overall quality of secondary effluents and in the present study also includes further
removal of ECs, more particularly PPCP compounds (Wang et al. 2007; Barceló &
Petrović 2008; Shannon et al. 2008).
Concerning sludge treatment techniques, here the most common and promising
technologies for this intent and for resource recovery in Europe were considered for
assessment (Rulkens 2007). It includes conventional methods designed to produce high
quality biosolids for agricultural application, and thermal processes undergoing constant
technological development that simultaneously promoting the recovery of diverse
resources from sewage sludge (Kelessidis & Stasinakis 2012; Fytili & Zabaniotou 2008).
Nonetheless, the adoption of such treatments is not to be expected exclusively for
the purpose of removing PPCP compounds, but first and foremost intending wastewater
reuse and resource recovery from sludge (González et al. 2015; Tyagi & Lo 2013; Hall
2014). The next section elucidates some of these aspects discussing these techniques’
usefulness concerning freshwater availability and sludge handling situation in Europe.
Afterwards, the selected wastewater treatment techniques are described in detail.
49
2.5.1. Water scarcity and integrated urban water management
Water is an indispensable and irreplaceable resource for virtually all modern
activities, from human consumption to industrial practices; it is also essential for
maintaining functional and sustainable ecosystems. However, pollution, wastage, and
water misuse, still frequent, present many challenges in the near future worldwide (OECD
2012). Furthermore, climate change has caused erratic rainfall and stream-flow patterns,
further worsening the problem (Barnett et al. 2005; Schewe et al. 2014). Therefore, in
order to satisfy current human requirements and guarantee functional environments, the
availability of good quality freshwater will be one of the major concerns in the following
decades, and actions should be undertaken towards this goal (Oki & Kanae 2006; Postel
2000; Cook & Bakker 2012).
Many indices to evaluate water scarcity have been proposed in the last decades
aiming to investigate the situation of many water basins and countries around the globe.
The Falkenmark indicator is one of the most commonly used indices that considers water
availability and respective human population. Ohlsson (1999) modified the Falkenmark
indicator to include social water scarcity considering the adaptive capacity to deal with
freshwater shortage, while other authors also incorporated specific concerns related to
usage by the agriculture sector (the most water demanding activity). Other models such
as that proposed by Alcamo et al. (2000) went even further by creating scenarios for future
global water scarcity, foreseeing in some cases that nearly half of the human population
will be living in water-stressed areas by 2025. However, since the water cycle has great
spatial and temporal variations, strong criticism of their calculations are common
(Jeswani & Azapagic 2011; Mekonnen & Hoekstra 2011; Smakhtin et al. 2004; Savenije
2000; Brown & Matlock 2011).
Nonetheless, due to the nature of the previously discussed evaluations, they still
do not encompass water scarcity at the local scale. However, many urban centres all over
the world have periodically faced acute freshwater shortage (Vairavamoorthy et al. 2008;
Yi et al. 2011). Additionally, there are increasing concerns about urban density and
economic growth. To cope with this subject, traditional approaches include the
construction of water dams and reservoirs. However, nowadays these practices are
becoming saturated or a somewhat delicate social-political issue, at least in most
developed regions, together with groundwater sources exploitation (Jury & Vaux 2007;
Rijsberman 2006).
50
As a more modern approach to the above cited issues, the integrated urban water
management (IUWM) deals with the topic at narrow frames, following a stricter
engineering-oriented approach regarding urban water infrastructure and availability at
water-basin level (e.g. flood management, storm water management, aquifer recharge,
increase distribution efficiency, grey water systems, etc.). Among them, wastewater reuse
is suggested as a practicable alternative for coping with freshwater scarcity, especially in
favourable conditions near the urban centres of developed regions (Miller 2006; Bixio et
al. 2006; Niemczynowicz 1999; NRC 2012). However, there are several issues for this
alternative to be broadly applied and considered as a sustainable and viable source of
freshwater. Advanced wastewater treatments have a key role in this context (Bogardi et
al. 2012; Tchobanoglous et al. 2011; Rodriguez et al. 2009), which is further explored in
the last section of this literature review.
Although water is regarded as abundant in the UK, several regions in the country
have already experienced water shortages, mainly associated with severe droughts, profit
seeking, and flawed socio-technical considerations, which have ultimately demonstrated
that water management in the UK is oftentimes ineffective or unprepared to deal with
more severe water crisis in the country (Marsh & Turton 1996; Marsh 2004; Taylor et al.
2009; Bakker 2000; Kowalski et al. 2011). Unlike southern European and other dry
regions around the world, in the UK wastewater reuse is not considered a crucial issue.
However, issues related to climate change, lack of further freshwater sources, and the
aforementioned discussed issues demand revaluation of the UK's future position
regarding wastewater reuse and similar actions, such as reduction of water pipe leakage
and flood resilience, in order to increase water security (Angelakis et al. 1999; Angelakis
& Bontoux 2001; Wilby & Perry 2006).
51
2.5.2. Present situation and future of sludge handling in Europe
Sludge generation increased by nearly 50% in Europe between 1992 and 2005 due
the enforcement of regulations for improving the quality of household effluents in the
region. Estimates of the total amount yielded in the European Union (EU-27) is
approximately 12 million tonnes of dry solids in 2010, and it is projected to increase
reaching over 13 million by 2020 (Milieu et al. 2010; Kelessidis & Stasinakis 2012).
Although estimates are inaccurate, the main disposal and recycling routes for sewage
sludge in most European countries are application on soil and incineration. Landfilling is
still a fairly common practice in many countries although it has already been restricted or
banned, previously occurred with sea disposal routes (European Commission 2001b;
Kelessidis & Stasinakis 2012).
According to projections, agricultural application of the sludge is expected to
slightly increase in the EU-27 to approximately 45% of the above cited tonnage. This
includes different types of pre-treatment to ensure compliance with application
requirements, including basic conditioning, anaerobic digestion and composting. These
last two treatments are the ones that have been considered by old member states to replace
other disposal routes such as landfilling and to cope with the increasing sludge generation.
Incineration is also expected to increase, stimulated by old member states, potentially
reaching around 35% of the total sludge produced by 2020 (Fytili & Zabaniotou 2008).
All routes are subjected to stricter regulations to avoid their negative aspects such as
volume occupied in landfills, wastefulness of nutrients and energy, air pollution, and
heavy metals release (Milieu et al. 2010; Fytili & Zabaniotou 2008).
The sewage sludge generated during wastewater treatment in the UK is largely
used in natural and agricultural soils (around 80% of the total, 20% and 60% each,
respectively), and to lesser extent it is directed to incineration units (18%) and landfill
sites (0.6%), totalling 1,413.103 tonnes of dry matter in the year 2010. The reuse of sludge
in agriculture is highly regulated, and anaerobic digestion is the most adopted route to
ensure compliance with directives. Furthermore, biogas from sludge can be used to
generate electricity, heat, and fuels, and this potential is currently being researched.
Although the reuse of sludge on agricultural land has faced considerable resistance from
stakeholders, mostly due to costs, stricter regulations, and concerns over heavy metals
and pathogen content, it appears to be the preferential disposal route for sludge in the UK
(DEFRA 2012; DEFRA/DECC 2014; DECC 2015; Appels et al. 2011).
52
2.5.3. Selected treatment techniques
In the present study, four advanced wastewater treatment techniques were
assessed: (i) granular activated carbon, (ii) nanofiltration, (iii) solar photo-Fenton, and
(iv) ozonation. They are preferably carried out along with MBR treatments because of its
overall superior quality compared to conventional secondary effluents and, therefore,
more beneficial for this research purposes due to lower interference in the advanced
wastewater treatment operation (e.g. lower dissolved and suspended solids concentration)
and high nutrients removal (there are not concerns associated with eutrophication
potential) (Sipma et al. 2010; Tambosi et al. 2010; Cases et al. 2011). Moreover, they are
capable to generate effluent compatible to potable water standards (e.g. high disinfection
rates, removal of metals, corrosion and pH control) (NRC 2012; Kazner 2011; Malato et
al. 2009).
In regards sludge treatment techniques, five treatments were selected: (i)
agricultural application of anaerobic digested sludge with electricity and fertilizer
recovery; (ii) agricultural application of composted sludge with fertilizer recovery; (iii)
incineration with electricity and heat recovery; (iv) pyrolysis with biochar and bio-oil
recovery; and (v) wet air oxidation with methanol recovery. They are coupled after an
ordinary thickening process to comply with the basic requirements of the selected
techniques (Andreoli & Von 1997; L. Wang et al. 2008). An overview of these techniques
coupled to WWTP operating MBR treatments is shown in Figure 7.
Figure 7 – Selected options for advanced wastewater and sludge treatment and their respective products
Preliminary +
primary treatment
Influent
Cinfluent,i
Effluent
Ceffluent,i
A. digestion
Membrane
filtration
Aeration basin
(see Figure 6)
Granular activated carbon
Recycled / Activated sludge
Su
per
nat
ant
Secondary sludge
Pri
mar
y s
lud
ge
Thickening
Crops
Centrifugation Waste
Filter pressing Pyrolysis
Nanofiltration
Solar photo-Fenton
Ozonation
Ad
va
nce
d e
fflu
ent
trea
tmen
t
Sludge
Csludge,i
Polymer addition
Composting
Waste
Mixing Crops
Filter bed
Wet oxidation
Incineration
T. drying
Potable effluent
Electricity / Fertilizers
Fertilizers
Electricity / Heat
Biochar / Bio-oil
Methanol
Slu
dg
e tr
eatm
ent
53
2.5.3.1. Selected options for advanced wastewater treatment
In this thesis, four options for advanced wastewater treatment were considered
due to two main reasons: (i) a substantial number of studies and enough information in
the literature regarding their operating requirements and removal of the target PPCP
compounds from wastewaters; (ii) to be among the most traditional or promising options
commonly considered for advanced wastewater treatment and are also effective in
reducing pollution originated from PPCP compounds and other ECs. The term
“traditional” refers to techniques already in use at large scale over the last decades known
to remove micro-contaminants at high rates (Wang et al. 2005; Yoon et al. 2007; Barceló
& Petrović 2008; ui et al. 2016; ousel et al. 2016). Granular activated carbon and
nanofiltration satisfy this criterion. y “promising” it can be understood methods being
increasingly adopted at small and industrial scale for the removal of micro-contaminants,
however still under development regarding their operating requirements and commercial
application in larger scales (Gogate & Pandit 2004b; Lofrano 2012; Liu et al. 2009;
Malato et al. 2009). These two are solar photo-Fenton and ozonation.
2.5.3.1.1. Granular activated carbon
The granular activated carbon (GAC) treatment operates by removing
contaminants through physical adsorption and biodegradation processes during passage
of the flux through single or several bed columns (contactors or mass transfer zone) at
predetermined time intervals (Macova et al. 2010; Simpson 2008). This technique is often
used due to its robustness, reliability, and relatively modest building requirements, often
showing low electricity demand (Grassi et al. 2012; Ng et al. 2011; Lee et al. 2009). The
GAC is a well-established treatment due to its great ability to remove a wide variety of
macro and micro-pollutants at varied concentrations, notwithstanding the removal in
wastewaters is less investigated (Delgado et al. 2012; Grover et al. 2011; Clements 2002;
Cabrita et al. 2010).
54
Some advantages of this type of treatment are providing a barrier against harmful
by-products (TPs) and the highly homogeneous removal (over 90%) of several organic
compounds. Disadvantages include the pathogens and heavy metals that occasionally
pass-through under malfunctioning conditions (Fuerhacker et al. 2001; Silva et al. 2012;
Yu et al. 2009; Zhang & Zhou 2005; González et al. 2015; Wang et al. 2005). Figure 8
shows the operation of this type of treatment and the main removal mechanisms of micro-
pollutants in the activated carbon particles.
Figure 8 – Scheme of granular activated carbon treatment and main removal mechanism of micro-
contaminants in granular activated carbon particles
The configuration of GAC treatment unit is based on previously acquired
knowledge. This is due to the multitude of important variables to be considered
simultaneously in terms of the overall removal of desired target contaminants and
economical aspects (Wang et al. 2005). However, the main criteria to determine its
configuration and feasibility are based on the definition of two variables (Reed et al. 1996;
Clements 2002):
• Empty bed contact time; and
• Bed service bed time.
GAC particle
Adsorbent radius
Mass transfer zone
Film diffusion
Entrapment
Influent
Regeneration
Effluent
Fresh GAC
Contactor Contactor
Maximum number
of regenerationsCoagulation tank
Chemicals
55
Empty bed contact time (EBCT) is a parameter used to assess the necessary
amount of granular activated carbon for influent treatment, which is determined by
estimating the required GAC volume to achieve a certain removal rate of a target
contaminant under predetermined operating conditions. Studies conducted in water
treatment plants have reported optimum EBCTs for drinking water treatment revolving
around 20 minutes with hydraulic loading rates of 2.0-18 m3/m2.h. and bed column depths
of 3.0-6.0 meters (Wang et al. 2005).
The mentioned above implies in different periods of time that the granular
activated carbon inside the contactors can maintain the desired contaminants removal rate
(i.e. breakthrough time), leading to an “exhausted” bed (Clements 2002; Lee et al. 2009;
San Miguel et al. 2001). This variable is called bed service time (tGAC), and it defines the
time when the carbon bed should be removed, replaced and regenerated due decreased
efficiency in removing contaminants. Furthermore, it is also influenced by the creation of
biofilm between the granules, often leading to extended tGAC. It has been found that tGAC
at drinking water treatment plants varies from 300 to 600 days (Wang et al. 2005;
Simpson 2008).
Therefore, the total amount of fresh and respective regenerated granular activated
carbon during the treatment life cycle can be estimated by variations in the EBCT, tGAC,
and the maximum number of regenerations (nmax). This last factor depends on the influent
composition and the method used for regeneration since they can significantly change the
initial activated carbon properties. Moreover, the number of regenerations should account
for losses in the carbon mass (mloss), commonly ranging from 10 up to 20% (Yu et al.
2008; Clements 2002; Creek & Davidson 2000; Bayer et al. 2005).
The removal of PPCP compounds by GAC treatment depend on the target
compounds characteristics, e.g. acid dissociation constant (pKa) or octanol-water partition
coefficient (Kow), operating parameters of carbon bed columns (hydraulic loading rate,
temperature), and wastewater composition (suspended solids, dissolved organic carbon,
and natural organic matter) (Zhang et al. 2013; Wang et al. 2005; Yu et al. 2008). High
removal rates ( > 90%) have been reported in the literature for many PPCP pollutants,
according to Delgado et al. (2012). Sewage and wastewater facilities applying MBR
and/or physicochemical post-treatment processes such as coagulation are recommended
when operating GAC since its effluents are often of higher quality and low in total organic
carbon, hence less interference in the activated carbon adsorption sites is likely to occur.
Moreover, such combination can also provide complementary action to remove organic
compounds (Snyder et al. 2007; Nguyen et al. 2012; Wang et al. 2005).
56
2.5.3.1.2. Nanofiltration
The nanofiltration (NF) treatment operates through high pressurized water fluxes
directed firstly to pre-filters devised to wider particles retention; thereafter the influx is
headed to filtration membranes with pore sizes from 0.1 up to 1.0 nm, giving rise to the
permeate (treated effluent) and the concentrate (Lee et al. 2009; Schrader 2006). The
contaminants are removed primarily through physical sieving, followed by adsorption
and electrostatic repulsion (Xu et al. 2004; Nghiem et al. 2005; Bellona et al. 2004). The
concentrate (typically less than 15% of the permeate) is often directed back to the influent
starting point (Nederlof et al. 2005; Bozkaya-Schrotter et al. 2009).
The nature and strength of removal forces are strongly dependent on the
physicochemical properties of the solute, e.g. molecular weight (MW), hydrophobicity,
wastewater composition and membrane properties; high removal efficiency of heavy
metals can be achieved, and formation of by-products (TPs) is avoided during the process
(Bellona et al. 2004; Xu et al. 2004; Qu et al. 2013). Furthermore, the overall removal
efficiency is related to applied pressure and flow rate (Ozaki et al. 2008; Yoon et al. 2006;
Comerton et al. 2008). The most suitable commercial material for NF treatment is
polyamide (Bolong et al. 2009; Drewes et al. 2005; Le-Minh et al. 2010). Figure 9 shows
the NF treatment operation and the main removal process in nanofiltration membranes.
Figure 9 – Scheme of nanofiltration and main removal mechanisms of micro-contaminants in nanofiltration
membranes
Exclusion by compound size
Pressurized feed
Permeate
Concentrate
Permeation
Adsorption
Membrane
fouling
Concentrate
Permeate
(effluent)
Pressurized influent
Nanofiltration membrane
Releasing/concentrate treatment
Chemicals
(effluent balancing)
Chemicals
(fouling control)
57
An important drawback of NF treatments is membrane fouling, which is the loss
of a membrane due to the deposition of suspended or dissolved solids and microorganisms
on its surface, at its pores openings, or within its pores. The process is characterized by
reduction of specific flux at constant pressure (and hence increased electricity
consumption), and it is particularly complex when treating wastewaters due the often high
content of dissolved organic matter (Bruggen et al. 2008; Nghiem & Hawkes 2007). Due
to this reason, plants operating MBRs are also beneficial for this type of treatment since
it decreases membrane cleaning periods, which is reflected in extended membranes
lifetime (potentially reaching 10 years of lifetime) (Chon et al. 2012; Qin et al. 2006;
Bonton et al. 2012; González et al. 2015). Therefore, the main aspects that should be
considered during NF treatments operation can be summarized as:
• Cleaning agent usage; and
• Electricity consumption.
To minimize undesirable formation of inorganic and organic deposits in the
membrane surface, chemicals substances (e.g. phosphates and acid anti-scalant) are
applied during the process (Al-Amoudi & Lovitt 2007; Shirazi et al. 2010). However,
when the original permeate flux is compromised, membrane cleaning is required (Wei et
al. 2010; Botton et al. 2012). To avoid membrane damage and effective treatment
operation, cleaning should be performed using an optimal combination of chemicals. The
choice of chemical cleaning reagents depends on the type of fouling to be removed and
the cleaning strategy (Nghiem & Hawkes 2007; Simon et al. 2013; Al-Amoudi & Lovitt
2007). Ethylenediaminetetraacetic acid (EDTA) and sodium hydroxide (NaOH) has been
suggested as an efficient combination to treat MBR effluents (Mo et al. 2010; Wei et al.
2010). There are no studies addressing the amount of cleaning chemicals for membrane
cleaning in the literature. Drinking water treating plants use approximately 4.2 g of
cleaning solution for every treated cubic meter (Bonton et al. 2012).
The electricity consumption in NF treatments is mainly due to pressurization
requirements for filtration (often taking place at 500 up to 1,000 kPa) and water heating,
both usually accounting for over 35% of the total operating costs (Bruggen et al. 2001).
Electricity consumptions in full-scale plants range from 0.27 to 0.53 kWh / m3 of treated
effluent, varying during the winter and summer seasons (influent ranging from 1 to 25ºC)
and desired feed flow (Cyna et al. 2002; Bonton et al. 2012).
58
Although nanofiltration membranes have been mainly used for desalination (i.e.
reverse osmosis), there are several studies addressing their capacity to remove PPCP
compounds. Analgesics and most antibiotics are known to have good potential to
excellent removal (usually higher than 90%) for different membranes and operating
conditions, while other compounds, especially hormones, have reported values often
considerably lower ( ≤ 75%) (Al-Rifai et al. 2011; Bodzek & Dudziak 2006; Sahar et al.
2011; Zazouli et al. 2009; Yoon et al. 2006; Nghiem & Schäfer 2004; Yoon et al. 2007).
According to Gur-Reznik et al. (2011), the influent composition variation is an important
factor for the removal of some compounds. In NF treatments, there is no generation of
by-products, and the removal of pathogens is usually high, especially if combined with
post-disinfection process, being therefore suitable for potable water reuse (Margot et al.
2013; Snyder et al. 2007; Alturki et al. 2010).
2.5.3.1.3. Solar photo-Fenton
Fenton treatments are advanced oxidation processes (AOPs) used in many
industries for wastewater treatment due their high efficiency in degrading most organic
contaminants and simple operation (Andreozzi 1999; Lofrano 2012; Oller et al. 2011).
The process consists of adding a catalyst and hydrogen peroxide to the influent that is
directed to special reactors irradiated by ultra-violet (UV) light, generating OH radicals,
which in turn oxidize contaminants (Gernjak et al. 2006). An approach to decrease the
operating requirements is the use of solar light for irradiation, which is called solar photo-
Fenton (SFP) treatment (Bauer & Fallmann 1997; Santiago-Morales et al. 2013; Gogate
& Pandit 2004a; Robert & Malato 2002). To ensure effective performance, a homogenous
influent distribution and strong acid environment (pH < 3.5) are required in the reactors
(Gernjak et al. 2006; Chong et al. 2010; Malato et al. 2009; Lofrano 2012; Klamerth
2011).
The infrastructure for the SPF comprises the assemblage of several panels in
number necessary to reach the required treatment capacity. The main part of a solar photo-
Fenton panel consists of the photo catalytic reactor, in which most of the reactions take
place (Klamerth 2011). Due the addition of many different chemicals (acids, catalyst, and
hydrogen peroxide), pipes and pumps should be made of resistant materials. They are
usually composed of high-density polyethylene (HDPE) or polypropylene (PP); the
materials used should also be inert to UV degradation (Klamerth 2011; Malato et al.
2009). A scheme of a solar photo-Fenton panel operation is shown in Figure 10.
59
Figure 10 – Scheme of a solar-photo Fenton treatment panel for wastewater treatment
Today there are only pilot-scale plants of SFP treatment for research purposes.
However, there has been increased research on this type of treatment aiming at making it
commercially feasible (Gernjak et al. 2006; Malato et al. 2009). Variations in the
following two parameters account for uncertainties due to operation, reactor geometry,
solar irradiation, temperature and the influence of other substances on the influent:
• Hydrogen peroxide dosage; and
• Catalyst dosage.
Hydrogen peroxide (H2O2) dosages commonly used in experiments utilizing
wastewater as matrix to reach high removal percentage (> 80%) of PPCP compounds
commonly vary between 50 and 150 mg/L. With regard to the catalyst, iron salts (usually
ferrous sulphate) dosages ranging from 5.0 up to 20 mg/L have been reported as near
optimum for many wastewaters (Klamerth 2011; Ortiz 2006; Trovó et al. 2013; Vogna et
al. 2004; Feng et al. 2005). The proper choice of iron salt is also crucial for efficient
removal, and it should be based on the effluent composition, chemical structure, and
initial concentration of the target compounds (Nogueira et al. 2005). Furthermore, higher
temperature may play an important role in the consumption of chemicals an process
efficiency (Gernjak et al. 2006; Malato et al. 2007).
Solar
irradiation
Fe2+ + H2O2 Fe3+ + OH˙ + OH-
θa
O
H O
Fe(OH)2+ Fe2+ + OH-UV light
Sola
r p
an
el
Glass tubes
H O
t0
Precipitates removal
Landfill
60
As a not well-establish technique yet, few treatments currently use SPF treatment
for the removal of micro-contaminants, and the content of DOC can hinder the efficient
removal of these compounds (Gernjak et al. 2004; Oller et al. 2011). Pilot-scale
experiments with secondary wastewater effluents showed removal efficiency varying
from as low as 20 % to near complete degradation of all target compounds, depending on
irradiation time, catalyst/H2O2 ratio proportions and contact time, as reported in the thesis
of Klamerth (2011). The production of harmful by-products (TPs) is a concern in terms
of the applicability of this treatment (Gogate & Pandit 2004b; Sirtori et al. 2009; Vogna
et al. 2004; Fernández-Alba et al. 2002; Malato et al. 2009).
2.5.3.1.4. Ozonation
The ozonation (OZO) treatment was first used in small drinking water treatment
plants in the beginning of the last century. After the 1970’s, this method started to be
increasingly used to obtain lower pathogens content in conventional wastewater treatment
effluents and a concomitantly effective removal of algae, colour, taste, odour, and several
organic compounds (Wang et al. 2005; Esplugas et al. 2007). Ozonation occurs with
direct and indirect reactions of pollutants with hydroxyl radicals (OH) generated by ozone
(O3) decomposition in the contactors (Huber 2004; González et al. 2015). This method
can achieve good metals removal but can also generate harmful by-products (Tripathi &
Tripathi 2011; Westerhoff et al. 2005). Figure 11 show a scheme of common OZO
treatment operation.
In OZO treatments, the ozone is generated from liquid oxygen or atmospheric air,
depending on the volume to be treated, consuming large amount of electricity (usually
corona discharges operating with over 10,000 V and up to 2,000 Hz). The overall
treatment efficiency is directly linked to the influent pH, alkalinity, and organic matter
content (Tripathi & Tripathi 2011; Broséus et al. 2009; Wang et al. 2005). There are
several variables to be considered when designing efficient ozonation systems, including
ozone contact time and diffuser shape. Two main variables are of concern:
• Transferred ozone dosage; and
• Electricity consumption.
61
Figure 11 - Scheme of ozonation process for wastewater treatment
Transferred ozone dosage (T) is the residual O3 from the applied ozone dosage
(DOzone), thereby active in achieving the required disinfection rate. This is due the fact
that DOC and other substances often react with ozone radical beforehand (i.e. initial
ozone demand). T is measured as transfer efficiency (TE), and it is strongly dependent on
the contactor design, diffuser type, and applied pressure. Thus, T and DOzone variables
should be studied simultaneously in water and wastewater treatments (Gogate & Pandit
2004a; Wang et al. 2005). Studies defining DOzone are usually doubled to attend field
conditions due to unpredicted hydraulic behaviour in large reactors. Nonetheless, DOzone
in wastewaters range from 4.0 up to 42 mg/L achieving high to very high pathogens
disinfection at common contact periods, i.e. CT99.9 ( > 99.9% inactivation of Giardia
cysts) in secondary effluents (Wang et al. 2005; Xu et al. 2002).
The electricity consumption is originated almost entirely from ozone generation
and the remaining is employed basically in water pumping and ozone destruction. The
electricity for generation is dependent on air particle filtration, ambient temperature, heat
loss, contactor design among others aspects (Wang et al. 2005). The choice of using
oxygen instead of ambient air for ozone production depends on the size of the treatment
plant (in larger plants on-site ozone production is advised). It has been reported that the
electricity consumption for ozone generation from oxygen is 9.92, and from ambient air
it is 16.53 kWh per kg of ozone generated (Kim & Tanaka 2011; Wang et al. 2005).
Influent
Effluent
Contactor
to
t1
O3
Transferred ozone
O3 + H2O O2 + 2OH ̇
Applied ozone dosage
Diffuser
Diffuser
Diffuser
Hea
tO
zon
e g
ener
ato
r
Ambient air /
Oxygen
Hea
t
Ozone
destruction
62
Estimations of PPCP removal by OZO treatments (as for other oxidation process)
are scarce and often doomed to fail without pilot-scale for a closer evaluation in each
wastewater and operating requirements (Huber et al. 2003). However, removal rates of
over 80% for many PPCP compounds have been reported by some authors when DOzone
is higher than 10 mg/L although some antibiotics and hormones showed lower removal
potential (Esplugas et al. 2007; Kim & Tanaka 2011; Ternes et al. 2003; Margot et al.
2013; Broséus et al. 2009; Y. Lin et al. 2009). The production of TPs (or disinfection by-
products) during OZO treatments is a problem, which should be closely evaluated when
considering wastewater reuse or release in sensitive areas. As for earlier described
techniques, the facilities operating MBR are beneficial since their effluents enable the
ozone to function with less interference from other substances in removing pollutants
(Laera et al. 2012; Huber 2004; Lee et al. 2012).
2.5.3.2. Selected options for sludge treatment
The present study discusses five different alternatives for the treatment of sewage
sludge. Three of them are the most commonly used sludge treatment techniques in most
old European Union member states (see topic 2.5.2): (i) agricultural application of
anaerobic digested sludge; (ii) agricultural application of composted sludge; and (iii)
incineration. The other two techniques have not been applied to treat significant amounts
of sludge although their viability has improved due to technological advances (Tyagi &
Lo 2013; Fytili & Zabaniotou 2008): (i) pyrolysis; and (ii) wet air oxidation.
2.5.3.2.1. Agricultural application of anaerobic digested sludge
The agricultural application of anaerobic digested sludge (ADG) begins with
anaerobic digestion itself. This is a process comprising hydrolysis followed by complex
microbial activities (acidogenesis, acetogenesis and methanogenesis) in ambient lacking
oxygen. The process generates carbon dioxide (CO2) and methane (CH4) as by-products,
primarily used to maintain the digestion reactor at suitable temperatures; the surplus is
occasionally used for electricity generation (Appels et al. 2008; Chen et al. 2008;
DECC/DEFRA 2011; Houdková et al. 2008; Yu & Schanbacher 2010). In Figure 12 is
shown a scheme of agricultural application of anaerobic digested sludge.
63
Figure 12 – Scheme of anaerobic digestion of thickened sludge for agricultural application
The digestion process can take place under mesophilic and thermophilic
conditions. The latter has been more frequently implemented because it produces higher
quality sludge, which can be used without restrictions in terms of pathogen content and
vector attraction potential (Iranpour et al. 2004; Fytili & Zabaniotou 2008). After
digestion, the sludge is mixed with polymers to facilitate dewatering until reaching a dry
matter content of around 25%. The product is then distributed to farmers as a substitute
for synthetic fertilizers due their high nutrient content and to improve soil characteristics
(Singh & Agrawal 2008; Hospido et al. 2005; X. Wang et al. 2008). Oftentimes, storage
is required to accumulate enough amounts of the product or to wait for more appropriate
application periods. Land application of the anaerobic digested sludge can be performed
in many different ways, as for example infiltration trenches, soil incorporation, or simple
surface spreading, which varies according to the property requirements and other factors
(L. Wang et al. 2008; EPA 1995).
One of the negative aspects regarding this technique is methane emissions. Part
of it is minimized by burning the biogas during anaerobic digestion. However, after land
application, the sludge decay can potentially generate methane that will eventually be
released into the atmosphere (Appels et al. 2011; Yu & Schanbacher 2010). Still
regarding land application, another issue is the excess of nutrients (e.g. nitrogen and
phosphorus) in the soil, which could leach to surface water and groundwater causing
eutrophication (Singh & Agrawal 2008). Furthermore, the heavy metal content also poses
risks to farmland soils (Udom et al. 2004). The minimization of these problems is usually
Thickened
sludge
Heat
generationBurning
Filter bed
Farmland application Storage
H
Mixing zone
Sludge zone
Fluid zone
Biogas
Direct to influent line
CO2
Heavy metals
64
possible with the use of conservative application rates (often ranging from 0.5 to 10 dry
tons per acre) and adoption of buffer zones (X. Wang et al. 2008; L. Wang et al. 2008;
EPA 1995).
There are few studies available regarding the removal of PPCP compounds by
anaerobic digestion. Attempts to predict their removal (mostly hormones) under different
operating conditions have shown great variability, and some produced contradictory
results (Carballa et al. 2007; Barret et al. 2010; Verlicchi & Zambello 2015).
Nevertheless, it has been suggested that their removal by anaerobic digestion is dependent
less on retention time or temperature and more on the sludge characteristics (Carballa et
al. 2005; Hospido et al. 2010; Ifelebuegu et al. 2010; Sim et al. 2011).
2.5.3.2.2. Agricultural application of composted sludge
The agricultural application of composted sludge is a process of natural
degradation of organic matter under controlled aerobic environments, generating
important products for agricultural use (Kosobucki et al. 2000). Sewage sludge can be
composted by mixing the sludge cake with bulking agents (e.g. bark or straw), followed
by the fermentation/maturation phase (internal temperatures ranging from 50 to 70C) for
stabilization, being turned from time to time until reaching the required composition for
farmland use (Roca-Pérez et al. 2009; Ponsa et al. 2009). The compost is forced to go
through mesophilic and thermophilic phases for long periods of time, which can vary
from days to months (Andreoli & Von 1997; L. Wang et al. 2008; Hernández et al. 2006).
Two different commercial-scale sludge composting processes have been
commonly used: windrow turner process and forced-aeration process (Hung et al. 2013).
The first one consists of covered windrows or elongated piles that are turned at a
decreasing frequency for control of moisture content and oxygenation. The second is a
more automated and controlled indoor process for greater optimization (Farrell & Jones
2009). The composted sludge produced has often better quality for soil amendment than
those from anaerobic digestion and lower methane emissions. However often lower
fertilizer potential (which provides better control regarding eutrophication in the applied
area) and bioavailability of heavy metals (Mantovi et al. 2005; Hernández et al. 2006;
Kosobucki et al. 2000; Singh & Kalamdhad 2012; Zigmontiene & Zuokaite 2010).
65
Figure 13 shows a scheme of agricultural application of composted sludge. In
regards removal of PPCP compounds, there studies in the literature suggest to that
digestion and composting have similar removal for many compounds since in both
processes the removal of organic pollutants is due to microorganisms (Poulsen & Bester
2010; Verlicchi & Zambello 2015).
Figure 13 - Scheme of composting of thickened sludge for agricultural application
2.5.3.2.3. Incineration
Incineration refers to the thermal degradation of materials in ambient with excess
of oxygen (i.e. combustion). Previously to the process itself, it is important to take into
account efficient drying technologies to reduce the water content in order to minimize
overall costs and increase its heat value (and consequently energy recovery potential)
(Rulkens 2007; Tyagi & Lo 2013; Werther & Ogada 1999). An advantage of this method
is the considerable volume reduction of sludge (often over 80%) (Kelessidis & Stasinakis
2012; Houdková et al. 2008; Stasta et al. 2006) but invariably producing combustion
wastes, e.g. ashes that are landfilled (bottom ashes) or deposited underground (fly ashes).
These ashes should be properly disposed of to avoid the release of heavy metals (Marani
et al. 2003; Hwang et al. 2007).
Thickened
sludge
Farmland application
Controlled environment
CH4
Bulk agent
Periodic turning
Windrows
Mixing
Heavy metals
66
Although the high temperature of the process guarantees the destruction of PPCP
compounds, an important drawback of this process is the emission of hazardous gases
(e.g. dioxins, CO2, SO2, NOx). Therefore, it has been consistently restricted by regulations
around the world, but the introduction of new technologies in the last years for the control
of gaseous emissions has contributed to the compliance with the ever restricted limits
(Kim & Parker 2008; Sänger et al. 2000; Fytili & Zabaniotou 2008) Moreover,
technologies for sludge incineration have greatly improved lately in terms of the process
engineering, energy recovery efficiency, and compactness (Fullana et al. 2004; Murakami
et al. 2009; Lundin et al. 2004; Cao & Pawłowski 2013; Andreoli & Von 1997). Figure
14 shows a simplified scheme for sludge incineration with generation of both electricity
(excess power directed back to the grid) and heat (distributed through district networks)
(Rulkens 2007; Murakami et al. 2009).
Figure 14 – Scheme of thickened sludge incineration for electricity and heat recovery
2.5.3.2.4. Pyrolysis
Pyrolysis refers to the thermal decomposition of organic material under oxygen-
free conditions; it is usually carried out in reactors at temperatures ranging between 300
and 900°C. It involves the generation of organic matter with recycling potential, including
several organic liquids (oils, acetic acid, acetone, and methanol), gases (hydrogen,
methane, carbon monoxide, carbon dioxide) and carbonaceous solids (Malkow 2004;
Werle & Wilk 2010; Luque et al. 2011).
Centrifuge
Hot air inletBottom ashes
(to landfill)
Heavy fuel
Thickened
sludge
Fly ashes
(to underground deposit)
Water
Electricity
District heating
Sand zone
Polymer
67
The proportion of the products generated during the process depends on
temperature, reactor residence time, pressure, turbulence, and characteristics of the
effluent (Fonts et al. 2009; Fonts et al. 2012; Werle & Wilk 2010). This type of advanced
treatment is an interesting alternative for sludge handling since it concentrates heavy
metals in the solid residue (landfilling) thus reducing their release to the environment
(Agrafioti et al. 2013) while promoting the complete destruction of organic pollutants
(e.g. PPCP compounds) due to the high temperature applied during the process.
Additionally, products of high energy content and with several potential uses are
recovered, with lower emissions of toxic gases when compared to incineration (Baggio
et al. 2008; Kim & Parker 2008; Thipkhunthod et al. 2006). Figure 15 shows a simplified
scheme of this process for the recovery of biomaterials and syngas.
Figure 15 – Scheme of thickened sludge pyrolysis for recovery of heat, bio-oil and biochar
2.5.3.2.5. Wet air oxidation
Wet air oxidation consists in the addition of polymer and often catalysts to the
sludge, which is thereafter pressurized and mixed with oxygen at high temperatures
(usually between 20-200 bar and 200-350ºC) during period varying from 15 to 120
minutes (Zou et al. 2007; Levec & Pintar 2007; Luck 1999). The results resulting in
carbon dioxide (released to the atmosphere), effluent containing high concentration of
carbon (redirected to the wastewater treatment line), and inert solid residue (disposed of
in sanitary landfills) (Tungler et al. 2015; Zou et al. 2007).
Thickened
sludge
Biochar / Bio-oil
Gas
clean-up
Sy
ngas
District
heating
Heat
generation
Polymer
ConditionerThermal
dryingFilter press
68
The carbon load in the liquid phase contains several low molecular weight
carboxylic acids, methanol, ethanol, acetone, etc., and their concentration and proportion
are dependent of the sludge composition and catalyst used during the process (Luck
1999). Nonetheless, they can potentially substitute considerable amounts of methanol
used in denitrification in wastewater treatment (i.e. biological process promoting
transformation of nitrogen compounds in nitrogen gas and further carbon removal)
(Kolaczkowski et al. 1999; Houillon & Jolliet 2005; Foglar & riški 2003). As for other
thermal processes, this alternative also promotes the complete destruction of PPCP
compounds (Zou et al. 2007). Despite its evident advantages, this alternative involves a
delicate operation and demanding maintenance, which impairs its broader use for sewage
sludge treatment (Wang et al. 2007; Chauzy 2010). An operational scheme of the wet air
oxidation method is shown in Figure 16.
Figure 16 – Scheme of thickened sludge wet air oxidation for recovery of methanol (for denitrification)
2.6. WASTEWATER TREATMENT AND SUSTAINABLE DEVELOPMENT
The next sections initially present the concept of sustainability and sustainable
development, followed by a brief discussion about wastewater treatment impacts
encompassing its benefits and negative impacts on the environment based on the literature
relevant to this topic. Lastly, the role of wastewater treatment in the rational management
of urban wastewaters and sustainable development is discussed.
Ox
idat
ion
to
wer
Thickened
sludge
Heat
exchanger
Inert waste
(to landfill)
Air
compressor
Recycled
Polymer
Denitrification
(as methanol)
Exce
ss
High-pressure
pumping
Gases
Catalyst
69
2.6.1. Sustainable development and sustainability assessment
Due to increasing concerns related to negative social and environmental
implications caused by the unbridled pursuit of economic growth in the post-war era in
the last century, there has been an increasing movement towards discussion and
minimization of these issues in many western world nations. The notions behind these
ideas and possible solutions could be interpreted as "sustainable development" (Pope et
al. 2004). One of the first attempts to define this concept can be dated from the early
1980’s, when the United Nations Environment Programme (UNEP) published the “world
conservation strategy”, afterwards broaden and settled in the Brundtland Commission
(Brundtland 1987). From then onwards, this concept has been continuously discussed and
interpreted from different points of view.
Nowadays, commitment to sustainable development is more than ever necessary
to solve modern “wicked” problems, maintain current society status, and ensure future
generation needs. Although mixed point of views about this concept prevails, it is
generally acknowledged that the triple-bottom life cycle assessment, considering the
environmental, financial, and social impacts, is the most appropriate (Sneddon et al. 2006;
Azapagic & Perdan 2014; Finkbeiner et al. 2010; Ness et al. 2007). The former is
nowadays the focus of discussion among the scientific community, while the inclusion of
the latter two are oftentimes paltry, creating gaps towards decision making among
stakeholders (Weidema 2006; Kloepffer 2008; Adams 2006). Figure 17 shows a Venn-
type diagram illustrating the basic conditions for achieving the sustainable development
goals.
Figure 17 – Tripartite interception approach defining the sustainable development goals (cross hatched
area)
Economic sphere
Social sphere
Environmental
sphere
70
2.6.2. Impacts of wastewater treatment and resource allocation
Although modern conventional wastewater treatments provide essential services
for improving environmental aspects and human health (e.g. controlling water bodies
eutrophication and spreading of pathogens) (WHO/UNICEF 2015; Montgomery &
Elimelech 2007), they also impose several environmental burdens (besides ecotoxicity
generated by PPCP compounds). In other words, although evident residual pollution
directly associated with effluent and sludge releasing, several environmental impacts are
generated from wastewater treatments. These are mostly derived from electricity
requirements and, to a lesser extent, use of chemical products. Examples of the potential
environmental impacts of conventional wastewater treatment on a life cycle basis are
available in the studies of Lemos et al. (2013), Pasqualino et al. (2009), Gallego et al.
(2008), Hospido et al. (2008), Rodriguez (2013), Lundin et al. (2004) and Suh &
Rousseaux (2002); reviews on this topic can be found in the study carried out by
Corominas et al. (2013) and Yoshida et al. (2013) for conventional wastewater and sludge
treatment techniques, respectively. Hence, WWTPs provide minimization of punctual
pollution, but marginal transfer of the pollution to dissipated sources (directed to water,
air and soil) occurs inevitably (Garrido-Baserba et al. 2014). Economic and social impacts
are not here review due to lack of information in literature.
2.6.2.1. Wastewater reuse feasibility
Improvements in urban infrastructure to enable WWTPs effluents to be further
utilized for diverse purposes (i.e. wastewater reuse) mean that the eventual release of
treated effluents to the environment can be minimized, as outlined in topic 2.5.1.
Although the use of lower quality wastewaters for reuse (e.g. for irrigation) can be
achieved with any secondary treatments combined to more simple post-treatments
(MBRs, UV irradiation, sand filtration), more stringent applications requires compliance
to potable water standards, even though this topic is still an open-ended subject in most
of the world (CDPH 2009; arceló & Petrović 2011; EPA 2012).
71
Studies comparing environmental profiles of DWTPs and some advanced
wastewater treatments suggested they present similar potential impacts. However, other
research concluded that, although significantly increasing the total impact of conventional
wastewater treatments, the addition of advanced treatments could lead to environmental
profiles lower than drinking water treatment facilities and expressively inferior than
desalination facilities for the production of potable water (Friedrich et al. 2010; Muñoz
& Fernández-Alba 2008; Stokes & Horvath 2006; Amores et al. 2013; Meneses et al.
2010; Pasqualino et al. 2011). This opens a door for promoting wastewater reuse as a
viable and sustainable option for coping with freshwater scarcity at urban level (as
discussed in topic 2.5.1).
A feasible practice of wastewater reuse is highly dependent on infrastructure for
the above-mentioned intent. For instance, it often requires that the effluents are
transported from WWTPs to strategic locations, for example back to DWTPs or other
effluent distribution locations (Tchobanoglous et al. 2011; Lemos et al. 2013; Stokes &
Horvath 2006; Zarghami et al. 2008; Rodriguez et al. 2009; Yi et al. 2011). This demands
large energy for water pumping, and other more complex issues involving piping
networks implementation in urban areas, increasing the environmental and economic
burdens of this practice. Not only that, issues concerning social acceptance have not yet
been elucidated (Urkiaga et al. 2006; Balkema et al. 1998; Salgot et al. 2006; Sala & Serra
2004; Hartley 2006; Chen et al. 2015).
2.6.2.2. Electricity, heat and fuel recovery from sludge
Among the different sources expected to contribute to the achievement of
renewable energy targets in Europe, sludge incineration is one of the least explored,
possibly because of the most promising results obtained using other alternatives in this
regard (Umbach 2010; Bagliani et al. 2010). On the other hand, some studies have
indicated incineration as a promising route for the sludge generated in Europe with similar
or better environmental profiles when compared to agricultural application of sludge.
This is due to the fact that agricultural application requires high electricity use,
transportation of large amounts of sludge and occasional storage (Suh & Rousseaux 2002;
Lundin et al. 2004; Kelessidis & Stasinakis 2012).
72
Therefore, since that sludge landfill disposal is to be banned or strongly
discouraged in a near future in Europe, the adoption of incineration with or without
recovery of electricity and heat could contribute to the responsible handling of previously
landfilled amounts (Fytili & Zabaniotou 2008; DEFRA/DECC 2014; DEFRA 2011).
Besides, incineration could grant over generation in regions not prone to or saturated by
biosolids application; therefore, even not contributing to renewable energy goals,
incineration could at least prevent sludge handling from being a significant environmental
burden (Rulkens 2007).
However, similarly to the case of wastewater reuse, the adoption of incineration
to recover most of the sludge energy content requires a great deal of urban infrastructure
for its heating distribution, i.e. district heating networks, which is economically
demanding and impairs its implementation in less densely populated areas. This is
especially true when competing against traditional heating sources (Pöyry Energy 2009;
Lund et al. 2010; Which? 2015). As for fuel recovery from sludge, novel techniques such
as pyrolysis are not yet available for broader use, which is mostly due to technical
problems and concerns, such as the estimation of marketing potential of its products in a
near future, therefore discouraging its immediate commercial adoption (Bridgwater &
Watkinson 2015; Werle & Wilk 2010; Ryu et al. 2007).
2.6.2.3. Nutrients recovery from sludge
Other relevant issues in sustainable development are food security, which refers
to the continuous provision of nutritious food to a growing human population without
compromising agricultural soil and natural environment, and lessening dependence on
fossil fuels and synthetic fertilizers (e.g. nutrients) (Godfray et al. 2012). With regard to
the above mentioned dependence, adverse effects derived from the use and natural cycle
of phosphorus and nitrogen were already observed worldwide (Canfield et al. 2010;
Childers et al. 2011; Elser & Bennett 2011; Gruber & Galloway 2008). Therefore, nutrient
recovery from sludge is important to ensure a healthier nutrient cycle (Galloway et al.
2008; Cordell et al. 2009; Tyagi & Lo 2013). Moreover, the toxicity impacts from heavy
metals is a major concern during nutrient recycle (e.g. agricultural application of digested
and composted sludge), and it should be evaluated accordingly (Singh & Agrawal 2008;
Udom et al. 2004).
73
2.6.2.4. Rational management of urban wastewaters
Effluents released into freshwaters or sludge landfilled without further recovery
of resources means that assets are being wasted or unavailable to further utilization in
urban centres after their treatment. Thus, actions towards increasing the suitability of
wastewater treatment is necessary to avoid such wastes of resources (Tyagi & Lo 2013;
Kärrman 2001; Muga & Mihelcic 2008; Lundie et al. 2004). As previously discussed,
conventional WWTPs have often been employed solely as a punctual source to minimize
the pollution originated from urban wastewaters, only recently have been integrated as
part of the broader concept proposed by the IUWM. According to the concept, the
adoption of specific wastewater treatment techniques is necessary to effectively extract
the resources contained in wastewaters in an efficient manner (i.e. wastewater reuse and
resource recovery from sludge), thus moving towards contributing to a more sustainable
development practices concerning wastewater treatment (Balkema et al. 2002; Balkema
et al. 1998; Murray et al. 2009; Amores et al. 2013).
This certainly will be a widely researched topic due to many forthcoming concerns
associated with climate change, resources scarcity, increasing urbanization and economic
growth (Devesa et al. 2009; Lim et al. 2008; Lim et al. 2010; OECD 2012). As expected,
there is a lack of knowledge about the topic in literature, especially in terms of the
adoption of wastewater treatment techniques and their role in promoting sustainable
management of urban wastewaters (Miller 2006; Bixio et al. 2008; Urkiaga et al. 2006;
Savenije & Van der Zaag 2008; Tyagi & Lo 2013; Rulkens 2007; Fytili & Zabaniotou
2008; Makropoulos et al. 2008).
Thenceforth, the wastewater reuse through the adoption of advanced wastewater
treatment techniques is expected to play a key role in decreasing dependence on natural
resources. These tasks will certainly deal with the presence of PPPC compounds and other
ECs in wastewaters, topic this discussed previously during this literature review.
Consequently, advanced wastewater treatment techniques are expected in a near future to
aid sustainable development practices by:
Fostering wastewater reuse for different applications at water-basin level;
Increasing energy security through recovery of electricity, heat and fuels; and
Promoting nutrients recycling in agriculture.
74
The topics discussed in this section were demonstrated to be closely related to the
interdependencies across the water, energy and food sectors, which can be more properly
understood, handled and managed through the energy-water-food (EWF) nexus approach.
This is also important because improvements in one of these sectors can potentially have
unwanted and unpredictable negative influence on the other two, thereby impairing their
sustainable development. Thus, the concomitant study of these topics has been
increasingly recognized as a requirement for a better understanding of the issues involved
on these topics (Bazilian et al. 2011; Beck & Villarroel Walker 2013; Mo & Zhang 2013).
Research regarding the role of WWTPs in the EWF nexus is at its very beginning, and
due to the intrinsic complexity of the subject a framework that would allow an upstanding
estimation of their potential influence in the nexus is highly desired, as acknowledged
earlier by previous authors (Mo & Zhang 2013; McCarty et al. 2011; Verstraete et al.
2009).
To sum up, considering the possibilities that advanced wastewater treatment
techniques could provide for a better management of urban wastewater, a scheme
portraying the issues discussed in this section is shown in Figure 18.
Figure 18 – Role of advanced wastewater and sludge treatment techniques in integrated wastewater reuse
of resource recovery management
Groundwater
recharge
Drinking water
treatment plant
City Factory
Eff
luen
t
Farm
Ind
ust
rial
reu
seA
gri
cult
ura
l re
use
Surface water
Infl
uen
t
Urb
an r
euse
Direct potable reuse
Non-potable uses
Po
tab
le w
ate
r d
istr
ibu
tio
n s
yst
em
Nu
trie
nts
Buffer
Blending
En
erg
y
Indirect potable reuse
Ad
va
nce
d
trea
tmen
t
Ad
va
nce
d
trea
tmen
t
Rec
har
ge
Con
ven
tion
al
trea
tmen
t
Con
ven
tion
al
trea
tmen
t En
erg
y
Landfill
75
3. METHODOLOGY FOR SUSTAINABILITY ASSESSMENT
The methodology for sustainability assessment of the advanced wastewater
treatment techniques applied in this work is divided in three main steps, as outlined in
Figure 19. It comprises: (i) estimation of concentration of PPCP compounds in WWTPs;
(ii) estimation and definition of operating parameters for the removal of PPCP
compounds by the advanced wastewater treatments; (iii) study of their potential
environmental, economic and social impacts on a life cycle basis, followed by an
integrated sustainability assessment and impact on the EWF nexus. These steps are
described in the sections below.
Figure 19 – Methodology for sustainability assessment of advanced wastewater and sludge treatment
techniques for the removal of PPCP compounds in wastewater treatment plants
3.2. Operating parameters, resource recovery and removal of PPCP compounds
3.3.2. Economic life
cycle assessment
3.3.1. Environmental life
cycle assessment
3.3.3. Social
life cycle assessment &
energy-water-food nexus
3.1. Selection of PPCP compounds and their concentrations in WWTPs
3.3.4. ulti criteria decision analysis
76
3.1. METHODOLOGY FOR ESTIMATING CONCENTRATION OF PPCP
COMPOUNDS IN WWTPS
The following sections describe the methodology for the estimation of
concentrations of PPCP compounds in wastewater treatment plants. The results can be
found in Chapter 4.
3.1.1. Selection of target PPCP compounds and data collection
The first step of the methodology involves selection of target PPCP compounds
from over 3,000 currently used in Europe alone (Daughton & Ternes 1999; Roig 2010;
World Health Organization 2004). In this work, the target compounds have been selected
based on data availability, environmental risks they pose and their different
physicochemical properties (and hence different behaviour during wastewater treatment).
(Jelena Radjenović et al. 2009; Petrie et al. 2014; Jelic et al. 2011). Thus, the following
14 PPCP substances ubiquitous in WWTPs are considered (see supplementary
information – SI - for their description):
analgesics: acetaminophen, diclofenac and ibuprofen;
antibiotics: erythromycin, trimethoprim and sulfamethoxazole;
cardiovascular beta-blocker: metoprolol;
lipid regulators: gemfibrozil and bezafibrate;
psychiatric drugs: carbamazepine;
hormones: oestrone and 17β-oestradiol;
antiseptics: triclosan; and
stimulants: caffeine.
To enable estimation of the parameters in steps 2-4 of the methodology, the
following data needed to be collected based on actual WWTPs and measurements:
influent and effluent concentrations of the target PPCP compounds;
wastewater influent into WWTPs; and
population served by WWTPs.
77
As part of this research, the above data were collected from the literature for 81
full-scale WWTPs based in different countries. These data are summarised in Table 1,
sorted by the region, starting with the countries in North America and followed by those
in Asia, Europe and, finally, Australia.
Table 1 - Target PPCP compounds in wastewater treatment plants in different countries
Location of WWTP
(Source of data)
Compounds and their average
concentrations in WWTP (influent;
effluent)
(µg/L)
Influent water
flow (m3/d)
Population served
(no. of inhabitants)
Treatment
type
US (Gao et al. 2012)
Sulfamethoxazole (1.10; 0.10)
Carbamazepine (0.10; 0.20)
Caffeine (42.0; 0.70)
45,400 Not stated Activated sludge
US (Conkle et al.
2008)
Acetaminophen (39.3; 0.02)
Ibuprofen (9.92; 0.08)
Sulfamethoxazole (4.09; 0.31)
Metoprolol (0.21; 0.02)
Gemfibrozil (1.65; 1.82)
Carbamazepine (0.06; 0.09)
Caffeine (25.6; 0.03)
7,200 Not stated Constructed
wetland
US (Thomas &
Foster 2005)
Diclofenac (0.47; ~0.0)
Ibuprofen (9.50; 0.02)
Triclosan (3.00; 0.08)
Caffeine (43.8; 0.04)
11,300 194,000 Activated sludge
US (Batt et al. 2007) Trimethoprim (7.90; 0.26)
Sulfamethoxazole (2.80; 0.63) 113,562 Not stated
Activated sludge
+ sand filter
US (Yang et al. 2011)
Acetaminophen (80.0; 0.05)
Diclofenac (0.22; 0.01)
Ibuprofen (11.0; 0.06)
Trimethoprim (0.61; 0.28)
Erythromycin (0.34; 0.27)
Sulfamethoxazole (2.60; 0.42)
Carbamazepine (0.23; 0.25)
Triclosan (0.47; 0.02)
Caffeine (80.0; 0.07)
227,000 Not stated Membrane bioreactor
Canada (Lishman et
al. 2006)
Diclofenac (0.20; 0.19)
Ibuprofen (8.45; 0.38)
Gemfibrozil (0.45; 0.25)
Oestrone (0.03; 0.01)
Triclosan (1.93; 0.11)
202,133 Not stated Severala
Canada (Atkinson et
al. 2012)
Oestrone (0.05; 0.10)
17β-oestradiol (0.05; 0.003) 422,000 786,130 Activated sludge
South Korea (Behera
et al. 2011)
Acetaminophen (7.50; 0.01 )
Diclofenac (0.15; 0.02)
Ibuprofen (2.20; 0.15 )
Trimethoprim (0.20; 0.04)
Sulfamethoxazole (0.09; 0.09)
Metoprolol (0.005; 0.004)
Gemfibrozil (0.20; 0.02)
Carbamazepine (0.10; 0.08)
Oestrone (0.05; 0.02)
17β-oestradiol (0.004; ~0.00)
Triclosan (0.55; 0.10)
451,000 1,100,000 Severalb
78
Caffeine (2.50; 0.02)
South Korea (Sim et
al. 2010)
Acetaminophen (8.00; ~0.00)
Diclofenac (0.01; 0.01)
Ibuprofen (1.00; ~0.00)
Erythromycin (0.75; 0.15)
Gemfibrozil (0.02; ~0.00)
Carbamazepine (0.30; 0.20)
Caffeine (6.00; 0.02)
1,302,100 3,600,000 Severalc
South Korea (Choi et
al. 2008)
Acetaminophen (31.9; 0.01)
Trimethoprim (0.22; 0.05)
Sulfamethoxazole (0.52; 0.16)
Carbamazepine (0.23; 0.09)
Caffeine (27.4; 0.32)
1,710,000 Not stated Activated sludge
Japan (Nakada et al. 2006)
Ibuprofen (0.80; 0.01)
Carbamazepine (0.08; 0.05)
Oestrone (0.04; 0.05)
17β-oestradiol (0.02; 0.01)
Triclosan (0.60; 0.10)
2,785,000 4,688,000 Severald
Japan (Nakada et al.
2007)
Ibuprofen (0.40; 0.01)
Carbamazepine (0.08; 0.03)
Oestrone (0.04; 0.02)
17β-oestradiol (0.02; 002)
Triclosan (0.55; 0.12)
170,000 460,000 Activated sludge
+ sand filter
Japan (Hashimoto et
al. 2007)
Oestrone (0.03; 0.04)
17β-oestradiol (0.012; 0.002) 158,012 Not stated Severale
Hong Kong (Leung et al. 2012)
Trimethoprim (0.20; 0.19)
Erythromycin (1.00; 1.00)
Sulfamethoxazole (0.10; 0.07)
2,081,000 5,381,900 Severalf
Hong Kong (Xu et al.
2007)
Erythromycin (0.86; 0.74)
Sulfamethoxazole (0.05; 0.03) 1,725,000 3,500,000
Modified
activated sludgeg
Hong Kong
(Gulkowska et al. 2008)
Trimethoprim (0.21; 0.23)
Erythromycin (0.55; 0.51) 1,377,000 3,500,000
Modified
activated sludgeg
China (Zhou et al. 2012)
Oestrone (0.08; 0.012)
17β-oestradiol (0.04; 0.002) 1,000,000 2,400,000 Activated sludge
China (Sui et al.
2010)
Diclofenac (0.35; 0.20)
Trimethoprim (0.30; 0.10)
Metoprolol (0.10; 0.09)
Gemfibrozil (0.04; 0.03)
Bezafibrate (0.04; 0.01)
Carbamazepine (0.15; 0.12)
Caffeine (6.00; 0.01)
2,200,000 6,109,000 Severalh
Spain (Gracia-Lor et
al. 2012)
Acetaminophen (55.1; ~0.00)
Diclofenac (0.53; 0.34)
Ibuprofen (14.6; ~0.00)
Trimethoprim (0.10; 0.09)
Sulfamethoxazole (0.45; 0.05)
Gemfibrozil (0.21; 0.49)
Bezafibrate (0.08; 0.06)
36,000 Not stated Activated sludge
Spain (Carballa et al.
2004)
Ibuprofen (3.70; 1.33)
Sulfamethoxazole (0.58; 0.25)
Oestrone (0.002; 0.004)
n.a. 100,000 Activated sludge
Spain (Radjenović et
al. 2009)
Acetaminophen (9.90; 0.11)
Diclofenac (1.32; 1.05)
Ibuprofen (21.7; 0.41)
Trimethoprim (0.20; 0.12)
Erythromycin (0.82; 0.54)
42,000 Not stated Activated sludge
79
Sulfamethoxazole (0.09; 0.02)
Metoprolol (0.04; 0.03)
Gemfibrozil (3.08; 3.08)
Bezafibrate (14.9; 3.01)
Carbamazepine (0.16; 0.16)
Spain (Santos et al. 2007)
Ibuprofen (94.1; 10.9)
Carbamazepine (0.30; 0.50)
Caffeine (2.17; 1.24)
164,500 Not stated Activated sludge
Switzerland (Tauxe-
Wuersch et al. 2005)
Diclofenac (1.90; 1.90)
Ibuprofen (2.80; 0.60) 9,300 23,000
Modified
activated sludgeg
Switzerland (Maurer
et al. 2007) Metoprolol (0.15; 0.10) n.a. 36,000
Activated sludge
+ sand filter
Finland (Lindqvist et al. 2005)
Diclofenac (1.00; 0.35)
Ibuprofen (13.3; 1.10)
Bezafibrate (0.50; 0.33)
353,330 1,174,000 Modified activated sludgeg
UK (Kasprzyk-Hordern et al. 2009)
Acetaminophen (211; 11.7)
Diclofenac (0.07; 0.10)
Ibuprofen (1.68; 0.26)
Trimethoprim (2.19; 1.15)
Erythromycin (1.61; 1.39)
Sulfamethoxazole (0.03; 0.01)
Metoprolol (0.08; 0.07)
Bezafibrate (0.42; 0.23)
Carbamazepine (1.69; 2.50)
36,160 111,000 Trickling filter beds
UK (Jones et al.
2007)
Acetaminophen (2.00; 0.10)
Ibuprofen (4.00; 0.50) Not stated 150,000 Activated sludge
UK (Zhou et al.
2009)
Diclofenac (0.98; 0.08)
Sulfamethoxazole (0.18; 0.03)
Carbamazepine (1.83; 0.84)
34,992 32,000 Activated sludge
+ sand filter
UK (Roberts &
Thomas 2006)
Acetaminophen (27.3; 0.002)
Diclofenac (0.98; 0.34)
Ibuprofen (23.2; 12.8)
Trimethoprim (0.26; 0.40)
Erythromycin (0.11; 0.20)
230,000 Not stated Activated sludge
Sweden (Zorita et al.
2009)
Diclofenac (0.23; 0.49)
Ibuprofen (6.90; 0.09)
Oestrone (0.02; 0.07)
17β-oestradiol (0.003; 0.0025)
20,000 55,000 Activated sludge
Sweden (Lindberg et
al. 2005)
Trimethoprim (0.25; 0.22)
Sulfamethoxazole (0.41; 0.19) 1,400,000 644,000
Modified
activated sludgeg
Italy (Baronti et al.
2000)
Oestrone (0.04; 0.03)
17β-oestradiol (0.01; 0.002) 734,000 1,200,000 Activated sludge
Australia (Watkinson
et al. 2007)
Trimethoprim (0.34; 0.05)
Sulfamethoxazole (0.36; 0.27) 140,000 700,000 Activated sludge
a 3 lagoons, 8 conventional activated sludge plants and 2 activated sludge + media filtration plants.
b 2 conventional activated sludge plants, 1 modified activated sludge plant and 2 other plants.
c 5 activated sludge and 6 other plants.
d 5 conventional activated sludge plants.
e 10 conventional activated sludge plants.
f 2 activated sludge plants and 5 other plants.
g chemically enhanced.
h 1 activated sludge with ozone and ultrafiltration plant, 1 activated sludge and sand filtration plant, 1 oxidation ditch
plant and 1 activated sludge with microfiltration and reverse osmosis plant.
80
3.1.2. Estimation of influx of PPCP compounds into WWTP and removal rates
The estimations in this and the subsequent steps are predicated on the following
assumptions:
the amount of PPCP compounds in the WWTP influent is directly proportional to
the per-capita consumption of PPCP, meaning that a plant serving a larger number
of inhabitants will receive a proportionally higher number of compounds in its
influent;
the consumption of the target compounds is assumed to be constant throughout
the year due to a lack of data; although it is acknowledged that some compounds,
such as analgesics, are expected to have higher consumption values and, therefore,
influx in winter, the seasonal variations will even out over a year; and
daily variations in the influent volume and any reactions of the compounds in
urban effluents before reaching the WWTPs are not considered, again due to a
lack of data.
The annual per-capita influx of PPCP compounds into a WWTP is estimated as
follows, using the relevant data in Table 1 for each WWTP:
IMinf,i = 365 x 10−3 x [Cinf,i x Q
p] (mg/inhab.year) (5)
where:
IMinf,i annual per-capita influx of PPCP compound i into WWTP (mg/inhab.year)
Cinf,i concentration of PPCP compound i in the WWTP influent (µg/L)
Q daily wastewater influent into WWTP (L/day)
p population served by WWTP (number of inhabitants)
The removal rate is calculated based on the WWTP influent and effluent
concentrations of PPCP compounds (see Table 1):
Rrate,i =Cinf,i−Ceff,i
Cinf,ix100 (%) (6)
81
where:
Rrate,i removal rate of PPCP compound i in a WWTP (%)
Ceff,i concentration of PPCP compound i in the effluent of a WWTP (µg/L)
In addition to the influent concentrations, the removal rate is influenced by the
design and operation of WWTPs, which in turn affect the concentration of the compounds
in the effluent (Ratola et al. 2012; Clara, Kreuzinger, et al. 2005; Roig 2010; Verlicchi,
Al Aukidy & Zambello 2012). To account for the variation in different parameters, the
expected concentration ranges for each PPCP compound in the WWTP influent and
effluent are considered in step 3, as detailed in the next section.
3.1.3. Estimation of concentration ranges of PPCP compounds in WWTPs
The box plot method was used to determine the expected influent and effluent
ranges for the PPCP compounds. For these purposes, the values for IMinf,i and Rrate,i,
estimated in the previous step for each compound and WWTP in Table 1, were grouped
into two datasets (A and B), respectively. Each dataset was then divided into four equal
quartiles, each containing a quarter of the data. Then, the first quartile was defined as the
middle (median) value between the lowest and the median value of the data set, the second
quartile as the median of the data set and the third quartile as the middle value between
the highest and the median value of the data set. The interquartile range, defined as the
difference between the third (upper) and first (lower) quartile, assumes that the values
will be bundled around a central (or median) value, as per the box-plot method. As a
result, the interquartile range was considered to be representative of the whole dataset for
each PPCP compound if it contains more than 50% of values (Potter 2006). The
interquartile range can also be used to define outliers, i.e. the values too far from the
central value or the expected range. Here, the high and low outliners were defined,
respectively, as those 1.5 times above the upper quartile value and 1.5 times below the
lower quartile value, following the box-plot method. Accordingly, the daily influx range
for each PPCP compound can be estimated as:
αrange,i =λrange,i
365x103 x p λrange,i A (g/day) (7)
82
where:
αrange,i estimated daily influx range for compound i in WWTP (g/day)
λrange,i IMinf,i value for compound i within the interquartile range (mg/inhab.year)
A dataset of IMinf,i values (mg/inhab.year)
The influent concentration range βrange,i in a WWTP is calculated according to:
βrange,i =103
365x
λrange,i
q λrange,i A (µg/L) (8)
where:
βrange,i estimated influent concentration range for PPCP compound i (µg/L)
q average daily per-capita wastewater influent into WWTPs (L/inhab.day)
The expected range of removal rates Rrange,i for each PPCP compound i was
determined using dataset B, where Rrange,i represents the interquartile range of Rrate,i,s
values (%). Therefore, the effluent concentration range for PPCP compound i can be
estimated according to:
γrange,i = βrange.i x (Rrange,i x 10−2) Rrange,i B (µg/L) (9)
where
γrange,i estimated effluent concentration range for PPCP compound i (µg/L)
Rrate,i removal rate range of PPCP compound i in WWTPs (%)
The concentration range of PPCP compounds retained by the sludge can be
estimated using the solid-water distribution coefficient and the sludge solids content
(Jones et al. 2002):
Srange.i =αrange,i x 10−3
((p x q) Kd,i⁄ )+(p x SDM) (kg/kg) (10)
83
where:
Srange,i concentration range of compound i in the sludge (kg/kg)
Kd,i solid–water distribution coefficient of compound i (L/kg)
sDM per-capita amount of dry matter in the sludge (kg/inhab.day)
3.1.4. Estimation of freshwater concentrations of PPCP compounds
Finally, using the values estimated in the previous steps, the predicted freshwater
concentration of the target PPCP compounds after the release of the WWTP effluent can
be estimated according to the following equation:
PECrange,i = γrange,i x [p x q
F+(p x q)] (µg/L) (11)
where:
PECrange,i predicted environmental concentration range of compound i in freshwater
after the release of WWTP effluent (µg/L)
F daily flow of a freshwater body (L/day)
The estimate of PEC is based on the following assumptions: there is no previous
PPCP contamination of a freshwater body; there is no prompt degradation of PPCP
compounds after the effluent discharge; and spatial and time variations in the
concentration of the target compounds are homogeneous.
3.2. OPERATING PARAMETERS, RESOURCE RECOVERY AND REMOVAL
OF PPCP COMPOUNDS
This step involved selection of advanced wastewater and sludge treatment
techniques and defining their operating parameters. As mentioned in Chapters 1 and 2,
the techniques considered for advanced wastewater treatment are: (i) granular activated
carbon, (ii) nanofiltration, (iii) solar photo-Fenton, and (iv) ozonation. The following
sludge treatment methods were selected for evaluation: (i) agricultural application of
anaerobic digested sludge, (ii) agricultural application of composted sludge, (iii)
incineration, (iv) pyrolysis, and (v) wet air oxidation.
84
The treatment plants are assumed to be based in the UK. The operating parameters
and their ranges were defined using literature, existing operating plants and own
calculations. This is explained in more detail below.
3.2.1. Main operating parameters and products recovery
The design and operating parameters of full-scale WWTPs are generally defined
according to the required final effluent quality and receiving water body. As a rule, the
removal of suspended solids (SS), total organic carbon (TOC), natural organic matter
(NOM) turbidity, nitrogen, phosphorus, heavy metals and pathogens are the major
concerns in wastewater treatment (Peters et al. 2003). The sludge composition depends
on the wastewater composition, effluent treatment operating parameters and
thickening/dewatering process during sludge conditioning (Andreoli & Von 1997).
Therefore, since each WWTP has different requirements, significant differences in the
composition of secondary effluent and thickened sludge are expected, consequently
influencing the operating requirements and recovery of products from the advanced
treatment techniques (Wang et al. 2009; L. Wang et al. 2008; Romero-Hernandez 2004;
Romero-Hernandez 2005).
The ranges of operation of the advanced wastewater treatment techniques were
then estimated based on data from existing treatment facilities, theoretical calculations
and literature, as detailed in the rest of the thesis. For the sludge treatment techniques, the
variation in the recovery of products was selected as a key parameter for accounting for
different sludge composition and its utilization potential since sludge treatment
techniques are highly customizable for optimal recovery of products. The key parameters
relevant to the selected treatment techniques are detailed below.
3.2.1.1. Granular activated carbon
The key operating parameters for GAC – the amount of fresh GAC and the
number of regeneration cycles before the bed needs to be replaced – were estimated based
on two criteria commonly considered in the design of GAC: empty-bed contact time
(EBCT) and the bed service time (see section 2.5.3.1.1). The initial amount of fresh
granular activated carbon was estimated for different EBCTs, as follows (Wang et al.
2005; Yu et al. 2008; Reed et al. 1996):
85
VGAC = EBCT x Qinf (12)
where:
VGAC volume of granular activated carbon in the bed (m3)
EBCT empty-bed contact time (min)
Qinf influent flow to be treated (m3/min)
The maximum number of bed regenerations (nmax) ensure the bed’s initial
characteristics are maintained (San Miguel et al. 2001; Clements 2002), consequently
defining the total number of bed replacements (NBR) over the lifespan of the unit as
follows:
NBR = Ttreatment (nmaxx tGAC)⁄ (13)
where:
NBR total number of bed replacements over its lifespan
Ttreatment treatment time (d)
nmax maximum number of bed regenerations before replacement
tGAC bed service time (d)
The amount of fresh and regenerated GAC was then calculated according to:
FGAC = mGAC[1 + NBR + mloss(nr + NBRnmax)] (14)
RGAC = mGAC(nr + NBRnmax) (15)
where:
FGAC amount of fresh GAC needed for the treatment (kg)
RGAC amount of regenerated GAC (kg)
mGAC amount of granular activated carbon in the bed (kg)
mloss percentage of GAC lost during regeneration (%)
nr number of bed regenerations after the previous bed replacement
86
3.2.1.2. Nanofiltration
Nanofiltration treatments are straightforward in their operation. It is commercially
prebuilt in modules, with the total number required varying according to the desired final
effluent quality and influent flow to be treated (see section 2.5.3.1.2). The main operating
parameters for this treatment are: (i) electricity consumption for pre-filtration, high-
pressure pumping, water heating during winter months and for lighting; and (ii)
membrane cleaning procedures using variable amounts of cleaning agents according to
the filtration pressure, influent composition and membrane properties (Bolong et al. 2009;
Yoon et al. 2006).
3.2.1.3. Solar photo-Fenton
Solar photo-Fenton treatments for wastewater treatment are still under
development and at present only pilot and industrial-scaled plants are operating (see
section 2.5.3.1.3). Two main parameters are recognized as critical for this technique: (i)
hydrogen peroxide dosage; and (ii) catalyst optimal dosage, both varying greatly
according to the desired final effluent quality (Robert & Malato 2002; Lofrano 2012;
Klamerth 2011).
3.2.1.4. Ozonation
The main operating parameters that need to be considered for ozonation units are
the amount of ozone required for efficient treatment and electricity consumption (see
section 2.5.3.1.4). The former can be calculated as (Wang et al. 2005):
Dozone =100
TEx T (16)
where:
Dozone applied ozone dosage (mg/L)
T transferred ozone dosage (mg/L)
87
TE ozone transfer efficiency (%).
The electricity consumers in the ozonation treatment include production of ozone,
pumping, recirculation and destruction of residual ozone. These were estimated as being
directly proportional to the amount of wastewater treated as follows:
Eozonation = Dozone x Vinf x Eozone (17)
where:
EOzonation electricity consumption for ozonation (kWh)
DOzone applied ozone dosage (kgozone/m3)
Vinf influent volume to be treated (the functional unit in this work) (m3)
Eozone electricity consumption for ozone generation (kWh/kgozone)
3.2.1.5. Sludge treatment techniques
The recovery of products from thickened sludge depends on the advanced sludge
treatment technique applied and often recovery of more than one product is possible.
Table 2 lists the products recovered by the sludge treatment techniques considered in this
work and the treatments (description can be found in sections 2.5.3.2.1-2.5.3.2.5). They
were chosen because they are the most commonly considered options in Europe and
because of data availability. The range for the products recovery considered here ranged
from total recovery (maximum efficiency according to references) to no recovery,
therefore including variations in the quality (i.e. composition) of the thickened sludge.
Table 2 – Products that can be recovered from advanced sludge treatment techniques and products that they
potential displace
Sludge treatment technique Products recovered (avoided products)
Agricultural application of anaerobic digested sludge Biosolids (synthetic fertilizers) and electricity (electricity grid)
Agricultural application of composted sludge Biosolids (synthetic fertilizers)
Incineration Heating (district heating network) and electricity (electricity grid)
Pyrolysis Biochar (charcoal) and bio-oil (heavy fuel oil)
Wet air oxidation Liquid stream with short chain of organic acids (methanol)
88
3.2.2. Estimation of the removal of PPCP compounds
There is a great variation in the removal of PPCP compounds by different
treatment methods reported in the literature. To determine typical removal rates of PPCP
compounds, a comprehensive literature review was carried out (see treatment’s
description), following the steps below:
• first, references were identified for each advanced wastewater treatment
techniques containing data on the removal of PPCP compounds;
• second, from the studies identified in the first step, the ones with composition of
effluents most similar to typical secondary effluents were selected;
• next, from the references selected in the previous step, the ones considering the
parameters and PPCPs similar to those commonly found in WWTPS and PPCP
were chosen for further consideration; and
• finally, from the above studies, the ones with the highest number of the target
PPCP compounds for this research were selected.
The above data were then used to estimate the removal percentages of the target
PPCP compounds for each advanced wastewater treatment technique, using the
physicochemical properties of the target PPCP compounds.
3.3. SUSTAINABILITY ASSESSMENT
The following methods were used for sustainability assessment of advanced
wastewater and sludge treatment methods:
life cycle assessment (LCA) for environmental sustainability assessment;
life cycle costing (LCC) for economic sustainability assessment;
social sustainability indicators for social sustainability assessment and EWF
nexus assessment; and
multi-criteria decision analysis (MCDA) for an integrated sustainability
assessment.
The methodologies for each are described below.
89
3.3.1. Life cycle assessment
The LCA methodology applied in this work followed the ISO 14040/10044
standard (ISO 14044 2006) as summarized in Figure 20. An LCA study starts by defining
the goal and scope of the study. This is followed by a compilation of an inventory of
relevant material and energy inputs and emissions to air, water and land. These are then
used to estimate relevant environmental impacts. The final stage in LCA is interpretation
of the results. The LCA steps are discussed in more detail below in the context of this
work. The LCA results for advanced wastewater and sludge treatment techniques are
presented in Chapters 5 and 6, respectively.
Figure 20 - Life cycle assessment methodology according to ISO 14044 (2006)
3.3.1.1. Goal and scope definition
The goal of this study was to assess environmental impacts of different advanced
wastewater and sludge treatment technologies. The system boundaries were defined from
“cradle-to-grave”. As indicated in Figure 21, this includes inputs into the treatment plant,
its construction, operation and decommissioning and recovery of various products. In the
case of ozonation and the sludge treatment alternatives, construction and
decommissioning of the plants were not considered due a lack of data. The functional unit
for the effluent treatment was defined as “treatment of 1,000 m3 of effluent from
secondary treatment” and for sludge as “treatment of 1,000 kg of sludge on a dry matter
basis”.
3.3
.1.4
.
Inte
rpre
tati
on
3.3.1.1.
Goal and scope definition
3.3.1.3.
Impact assessment
3.3.1.2.
Inventory analysis
90
Figure 21 – System boundaries for life cycle assessment of the advanced wastewater and sludge treatment
techniques
3.3.1.2. Inventory analysis
There is a significant lack of data and uncertainties in inventories for WWTPs
(Corominas et al. 2013; Foley et al. 2010; Lane et al. 2012). To fill in data gaps and create
more reliable inventories, the methodology applied in this work for data collection is
outlined Figure 22. As indicated, data were collected from previous studies of PPCP
removal rates by different treatment methods, technical studies of their operating
requirements and previous life cycle inventory (LCI) studies. These were used either
directly or to estimate/extrapolate the missing data. Care was also taken that data between
different studies were compatible and comparable, to ensure consistency. The
background data were sourced from the Ecoinvent v2.2 database (Frischknecht et al.
2004). Further detail on the LCI data can be found in Chapters 5 and 6.
Secondary effluent /
Thickened sludge
Marketable products
(from sludge)
Treatment operation
Treatment decommissioning*
Emissions and wastes
*Not included in the potential environmental impacts for ozonation advanced sludge treatment techniques.
Avoided
Landfill
Irrigation
91
Figure 22 – The methodology for creation of life cycle inventories considered in this work
During data acquisition, it was necessary to scale-up some of the unit operations
and infrastructure to estimate the amount of materials used their manufacture or
construction. The “economies of scale” method was used for this purpose, based on the
approach in Coulson et al. (1993) as modified by Greening & Azapagic (2012):
C2 = C1x (c2
c1)
0.6
(18)
where:
C1 and C2 material requirements for smaller and larger scale, respectively
c1 and c2 respective treatment capacities
0.6 “economy of scale” factor
Influent and effluent PPCPs concentrations
Number of target PPCP compounds
astewater / Sludge characteristics
perating parameters
PPCP compounds removal studies
Previous LCIs
Co
mp
ati
bil
ity
92
3.3.1.3. Impact assessment
Two life cycle impact assessment (LCIA) methods were used in this work:
ReCiPe 2008 v1.08 (Goedkoop et al. 2009) and USEtox 1.0 (Rosenbaum et al. 2008).
The former was selected as the state-of-the-art LCIA method and the latter for its
relevance to freshwater ecotoxicity related to the PPCP compounds. The ReCiPe
midpoint impact categories listed in Table 3 were considered. The systems were modelled
and impacts calculated in LCA software Gabi 6.0 v4.3 (thinkstep 2015).
Table 3 – Recipe 2008 midpoint impact categories considered in this work
Impact category and description Abbreviation Unit/functional unit
Climate change
Environmental mechanisms linked to global warm potential CC kg CO2 Equiv.
Fossil fuel resource depletion:
Reduction of available hydrocarbons resources and its future availability FD kg oil Equiv.
Metal depletion
Reduction of available metals resources and its future availability MD kg Fe Equiv.
Water depletion
Amount of fresh water consumed WD m3
Ozone depletion
Environmental mechanism responsible for depletion of the stratospheric
ozone layer and consequent increase of solar UV- radiations on Earth’s
surface
OD kg CFC-11 Equiv.
Freshwater eutrophication
Nutrient enrichment of freshwater environments FE kg P Equiv.
Marine eutrophication
Nutrient enrichment of marine environments ME kg N Equiv.
Terrestrial acidification
Impact on soil acidity TA kg SO2 Equiv.
Ionizing radiation
Damage to ecosystems from radioactive materials in the environment IR kg U235 Equiv.
Freshwater ecotoxicity
Toxicological damage to freshwater species FET kg 1,4-DCBa Equiv.
Terrestrial ecotoxicity
Toxicological damage to terrestrial species TET kg 1,4-DCBa Equiv.
Marine ecotoxicity
Toxicological damage to marine species MET kg 1,4-DCBa Equiv.
Human toxicity
Damage to human health HT kg 1,4-DCBa Equiv.
Natural land transformation
Transformation of forests, seas and other natural environments NLT m2
Urban land occupation
Occupation of urban and area ULO m2.year
Agricultural land occupation
Occupation of agricultural area ALO m2.year
Photochemical oxidants formation
Formation of harmful chemical compounds from solar irradiation and air
pollutants
POF kg NMVOC Equiv.
Particle matter formation
Formation of atmospheric particulates PMF kg PM10 Equiv.
a Dichlorobenzene
93
3.3.1.4. Interpretation
As mentioned earlier, uncertainties in this kind of analysis can be significant due
to a lack of data. To ensure a more robust interpretation of the LCA results, sensitivity
and parametric analyses was carried out for the main operating parameters for the
advanced wastewater treatment techniques and for products recovery potential for the
sludge treatment methods. This is outlined in Figure 23. For the effluent from the
secondary treatment, its quality was assumed to range from superior to inferior for three
values of main operating parameters: minimum, maximum and mean. Similarly, the
sludge was also assumed to range from superior to inferior quality, and the potential
recovery of products from maximum to minimum, respectively.
Figure 23 – Sensitivity analysis assuming variations in the quality of the secondary effluent and thickened
sludge
3.3.2. Life cycle costing
Since there is no standardised LCC methodology, the code of practice proposed
by Swarr et al. (2011) was adopted as it is compatible with the LCA methodology. The
economic costs were then estimated based on the LCI. This is described in more detail
below. The results of LCC are discussed in Chapter 7.
Superior
Mean
Inferior
Secondary effluent
Quality
Treatment operation
Minimum main operating parameters
Mean main operating parameters
Maximum main operating parameters
Superior
Mean
Inferior
Thickened sludge
Quality
Products recovering*
Maximum recovery potential
Mean recovery potential
Minimum recovery potential
Operating range
Products recovery range
*Subject to commercial exploitation potential
94
3.3.2.1. Data sources for costs
The costs for construction and infrastructure were sourced from engineering
handbooks, previous estimates and data published in the literature. Prices of chemical
products, energy and other materials were estimated based on bulk prices from suppliers
in the UK, US and China. The costs were updated to the present values using the
corresponding exchange rates and inflation. The cost data are detailed in Chapter 7. The
next section gives an overview of how the costs were calculated.
3.3.2.2. Life cycle costs calculations
The LCC costs were estimated as follows:
LCC = CC + IRC + FC + VC + WC + TC – S (£/functional unit) (19)
where:
LCC total life cycle costs
CC construction costs
IRC infrastructure replacement costs (advanced wastewater treatment methods only)
FC fixed operating costs
VC variable operating costs (advanced wastewater treatment methods only)
WC waste management costs
TC transport costs
S revenue from the sales of recovered products (sludge treatment methods only)
95
3.3.3. Social sustainability indicators
The social assessment aim to include the impacts on society, from local to global
scale, of the stakeholders along the life cycle of the product or service under question.
However, several issues relative to measurement and quantification of social indicators
are yet to be solve since there is not yet a defined methodology for social assessments and
due to the inherent complexity of the topic (Finkbeiner et al. 2010). Nevertheless, a
number of generic social indicators are available concerning labour and commercial
trades, and methodologies and guidelines for social assessment is under development for
broader application in sustainability assessments (Benoît-Norris et al. 2011).
The social sustainability assessment was here developed from fundamental life
cycle social assessments principles (UNEP-SETAC 2009; Muthu 2015; Jørgensen et al.
2010; Jørgensen et al. 2008), guidelines and sustainability indicators concerning urban
water-wastewater infrastructure (Balkema et al. 1998; Balkema et al. 2002; McConville
& Mihelcic 2007; Hellström et al. 2000; Urkiaga et al. 2006; Fytili & Zabaniotou 2008;
Kennedy & Tsuchihashi 2005; Carvalho et al. 2009; Wüstenhagen et al. 2007). In total,
six social issues were selected and quantified using 16 indicators listed in Table 4. These
issues and indicators were selected for consideration due to their relevance to the
treatment methods considered here but also due to the data availability. The issues are
split into the national, supplier and consumer levels, as described below. The results of
the social sustainability assessment can be found in Chapter 8.
Table 4 – Social issues and indicators for social sustainability assessment of advanced wastewater and
sludge treatment techniques
Level Social issue Indicator Unit per functional unit
National
Water securitya Water stress -
Net water use m3
Energy securitya
Net energy use kWh
Imported fossil fuel avoidedb koec
Diversity of outputsb -
Food securitya Agricultural land use m2.year
Synthetic fertilizer avoidedb kg of Pd
Suppliers Adoption and the market Potential for product utilization -
Public opposition to the treatmentb -
Consumers
Human health Damage to human health DALYe
Emerging contaminants and heavy metalsf -
Product acceptance
Wastewater reuse acceptancef -
Presence of harmful substancesb -
Similarity to traditional productsb -
The rebound effect -
a Energy-water-food nexus issues and indicators. b Only for sludge treatment. c koe – kilograms of oil equivalent. d P – phosphorus. e DALY – Disability Adjusted Life Years. f Only for advanced wastewater treatment.
96
3.3.3.1. Social issues at the national level
At the national level, three issues pertinent to wastewater and sludge treatment
were considered, related to water, energy and food security. These were first considered
individually and then integrated to determine the potential effect of the treatment methods
on the EWF nexus.
3.3.3.1.1. Water security
This issue relates to water availability at the national level and the contribution of
the treatment methods to the national water reserves. This was quantified through two
indicators:
• national water stress, expressed as the national Water Stress Index (WSI) to determine
if water scarcity is an issue for the country; and
• net water use by the treatment plants, calculated as the difference between the amount
of the treated water that can be discharged into the environment and the water
consumed in the life cycle of the treatment process.
3.3.3.1.2. Energy security
As for the water security, this issue considers the effect of the treatment techniques
on the national security of energy supply. Three indicators were quantified within this
issue:
• net energy use, calculated as the difference in the energy (electricity and/or heat)
consumed and generated (if any) by the advanced treatment;
• imported fossil fuel avoided (sludge treatment only), which evaluates the potential
avoidance of fossil fuels by recovering products from sludge treatment; and
• diversity of outputs (sludge treatment only), indicating the degree to which the
treatment can contribute to diversification of energy supply.
97
3.3.3.1.3. Food security
This issue aims to evaluate the potential of the treatment methods to contribute
towards food production and improve the national security of food supply. This was
assessed through two indicators:
• land use, estimating agricultural land occupation in the life cycle of the treatment
methods; and
• avoidance of the use of synthetic fertilizers (sludge treatment only), representing
the amount of synthetic fertilizers potentially displaced by recovering nutrients
from the treated sludge. This indicator is considered a social issue for several
reasons: synthetic fertilizers are associated with human health issues; they deplete
natural phosphorous which is becoming a scarce resource and may not be
available to future generations, thus raising the intergenerational equity issues;
and using widely-available supplies of nutrients improves food security.
3.3.3.1.4. Energy-water-food nexus
Water, energy and food are intimately intertwined, creating the EWF nexus.
Consequently, improvements in one of these sectors can potentially have unwanted and
unpredictable negative influence in the other two, thereby impairing their sustainable
development. To enable evaluation of the impacts on the EWF nexus, a new methodology
was developed as part of this research using the above indicators for water, energy and
food security. The method is based on the integration of the nexus indicators and
examination of the delineated area (quantitative aspect) and shape (qualitative aspect)
generated by their plotting in a specific manner. Due to its generic nature, the method is
suitable for use in any case that this type of assessments is required and provide a visual
communication of their interconnection.
Nexus indicators
The methodology starts by calculating the nexus issues scores in energy, water
and food (see social issues and indicators at national level in Table 4) for each treatment
technology according eqns. (20)-(21). It is worth noting that scores are calculate for lower
values as preferable in all steps.
98
vk,a =∑ vk,i,a
′Ji=1
J (20)
where:
vk,a nexus score for nexus issue k (water, energy or food) for technology a
v’k,i,a normalized nexus indicator i for nexus issue k and technology a
J total number of nexus indicators i for nexus issue k and technology a
The normalized nexus indicator score is estimated as:
vi,a′ =
vi,a−yi,a
Yi,a−yi,a (21)
where:
vi,a estimated value of criterion i (nexus indicator) for technology a
yi,a minimum value of criterion i for technology a
Yi,a maximum value of criterion i for technology a
Nexus indicators integration and impact assessment
After the nexus issues scores are calculated, the result is plot inside the triangle
shown in Figure 24 (nexus triangle). It depicts three axis of unit length with origins at the
center of an equilateral triangle having angles of 120o among them, delimitating the
maximum nexus area. The nexus impact comprises: (i) two categories regarding the
triangle area and shape, assessing quantitative and qualitative aspects of the nexus
respectively; and (ii) one category depicting the relationship among the latter two
categories, giving the final score balancing the overall impact on the nexus. These are
commented next.
99
Figure 24 – Axis configuration for the integration of nexus indicators (nexus triangle)
Nexus influence
The total area of the triangle in Figure 24 is of approximately 1.30, calculated
using the Heron’s formula (eqns. 22-23) for nexus issues scores (vk) equal 1. On these
lines, the nexus influence (Anexus – eqn. 24) is defined as the normalized area of the
triangle formed by connecting the nexus issues scores (with sides xa, ya and za). In Figure
25 there are some key Anexus scores for the combination values for two nexus issues scores
(thus total score should be calculated by the summing the combination of the scores for
water-food, energy-food and water-energy).
Snexus,a = √(Sax(Sa − xa)x(Sa − ya)x(Sa − za)) (22)
Sa =(xa+ya+za)
2 (23)
Anexus,a =Snexus,a−wnexus
Wnexus−wnexus (24)
where:
Food security
Water-Food nexus Energy-Food nexus Water-Energy nexus
1
1
y
z
vw
ate
r
1
100
Snexus,a area in the nexus triangle for technology a
xa, ya and za sides of the triangle in the nexus triangle for technology a
Anexus,a nexus influence score for technology a
wnexus minimum area in the nexus triangle (0)
Wnexus maximum area in the nexus triangle (~1.30)
Figure 25 - Nexus influence (Anexus) according different vk values
Nexus homogeneity
For assessing the distribution homogeneity of the area in the nexus triangle, the
nexus homogeneity (SDnexus) category defines how nexus issues scores are close to form
an equilateral triangle; the standard deviation given by eqns. (25) and (26) describe the
previously mentioned:
SDnexus,a = √1
N−1∑ (vk,a − vaverage,a)
2Nk=1 (25)
vaverage,a =∑ vk,a
Nk=1
N (26)
0.00
0.05
0.10
0.15
0.20
0.25
0.30
0.35
0.00 0.10 0.20 0.30 0.40 0.50 0.60 0.70 0.80 0.90 1.00
Vk2 = 1.0 Vk2 = 0.90 Vk2 = 0.80 Vk2 = 0.70
Vk2 = 0.60 Vk2 = 0.50 Vk2 = 0.40 Vk2 = 0.30
Vk2 = 0.20 Vk2 = 0.10 Vk2 = 0.05 Vk2 = 0.01
vk1
Nex
us
infl
uen
ce (
An
exu
s )
101
where:
SDnexus,a nexus homogeneity score for technology a
vaverage,a average of the nexus issue scores for technology a
N total number of nexus issues (3)
Nexus score
Finally, the overall preference is given by the nexus score (Nscore), defined by the
relationship in eqn. 27:
Nscore,a =Anexus,a
(1−SDnexus,a) (27)
where:
Nscore,a nexus score for technology a
Anexus,a nexus influence score for technology a
SDnexus,a nexus homogeneity score for technology a
Interpretation
Nexus influence: this category estimate the interaction in the nexus using the three
indicators (energy-water-food); lower values (i.e. smaller triangle area) translates
to lower influence in the nexus; differentials in nexus indicators further away from
the origin have greater variations in the delineate triangle area (see Figure 25),
suitable to this assessment;
Nexus homogeneity: the interpretation in this step is that equilateral triangles
reflects homogeneous impacts in the nexus (i.e. optimum nexus influence
distribution), hence promoting more equilibrated distribution of the resources and
preferable if compared to other triangular shapes. In it, a value equal to zero
describes perfect equilateral triangles and higher scores less equilateral triangles;
and
Nexus score: this category reflects both quantitative and qualitative aspects on the
nexus simultaneously. Following the outlined above, low values are preferable
and high values are less preferable.
102
3.3.3.2. Social impacts at the supplier level
3.3.3.2.1. Adoption and marketing
This issue deals with concerns related to the adoption, commercialization and
distribution of products produced by the treatment techniques. It was evaluated through
the following two indicators:
• potential for the utilization of products, indicating their readiness for adoption and
implementation; and
• public opposition to the treatment (sludge treatment only), due to the aspects such
as pathogens content, odour, attraction of insects, air pollution, health and other
issues.
3.3.3.3. Social impacts at the consumer level
3.3.3.3.1. Human health
The human health issue relates to adverse effects on human health in the life cycle
of the advanced treatments through two indicators:
• damage to human health, estimated as the sum of the potential hazards caused by
climate change, human toxicity, ionising irradiation, ozone depletion, particulate
matter formation and photochemical oxidants and expressed as disability-adjusted
life year (DALY); and
• emerging contaminants and heavy metals (effluent treatment only), indicating
their presence in the treated water and the related effects on human health if the
advanced method is meant for reuse of the effluent as potable water.
103
3.3.3.3.2. Product acceptance
This issue recognises the importance of public acceptance of products produced
by the advanced treatment methods that could potentially substitute traditional products.
It comprises four indicators:
• wastewater reuse acceptance (effluent treatment only): currently wastewater reuse
is practiced at very few locations worldwide, mostly as a last resort so this
indicator considers how consumers may respond to a wider deployment of
wastewater reuse;
• presence of harmful substances (sludge treatment only) considers concerns related
to the pathogens, heavy metals and air pollution from sewage sludge treatment;
• resemblance to traditional products (sludge treatment only) compares the overall
quality of the products generated by sludge treatment to traditional goods they
aim to substitute; and
• the rebound effect takes into account that adoption of a “green” product can lead
to a rise in consumption of that or other products.
3.3.4. Multi-Criteria Decision Analysis
Trade-offs among environmental, economic and social criteria are inevitable in
sustainability evaluations, and ultimately, the choice of sustainable options will
ultimately depend on stakeholder preferences (Dodgson et al. 2009). In situations like
these, MCDA methods are frequently used to help integrate different aspects of
sustainability based on stakeholder preferences and help to identify the most sustainable
options among the alternatives considered (Azapagic & Perdan 2005a). From the many
MCDA methods, this work applied the multi-attribute value theory (MAVT), often used
in sustainability assessments. This is due to its suitability for application to cases in which
the stakeholder can comprehensibly express its preferences among the criteria and the
outcomes or consequences of the alternatives assessed are previously known (Azapagic
& Perdan 2005b; Seppälä et al. 2002; Rahimi & Weidner 2008), as is the case of this
study. In MAVT, the overall sustainability score (va) for each alternative is estimated first
for each sustainability aspect (vs,a) as follows:
vs,a = ∑ wi,sv′i,s,aIi=1 (28)
104
where:
vs,a sustainability score for technology a for sustainability aspect s (environmental,
economic or social)
wi,s weight of importance of decision criterion i (sustainability indicator) in
sustainability aspect s
v’i,s,a normalized sustainability score for technology a for criterion i (sustainability
indicator) and sustainability aspect s
I total number of decision criteria i (sustainability indicators) for sustainability
aspect s
The normalized sustainability score is estimated as:
vi,a′ =
vi,a−xi,a
Xi,a−xi,a (29)
where:
vi,a estimated value of criterion i (sustainability indicator) for technology a
xi,a minimum value of criterion i for technology a
Xi,a maximum value of criterion i for technology a
The above results are then used to estimate the overall sustainability score (va)
according to equation:
va = ∑ Wsvs,aAs=1 (30)
where:
vs,a sustainability score for technology a for sustainability aspect s (environmental,
economic or social)
Ws weight of importance of sustainability aspect s
A total number of sustainability aspects (3)
105
In this method, the low scores indicate greater sustainability. All sustainability
indicators and aspects were assumed to have the same importance. However, the
influence of the importance of the sustainability aspects on the results was tested through
sensitivity analysis. The assumed weights can be found in Table 5 and the results of the
MCDA are presented in Chapter 8.
Table 5 - Weights of importance for the environmental, economic and social indicators considered in the
MCDA
Aspect Total no. of indicators (I) Indicator weights
(wi,s)
Aspect weights
(Ws)
Environmental (LCA impacts) 18 = 1/18 = 0.056 = 1/7 = 0.143 (minimum)
Economic (LCC) 1 = 1/1 = 1 =1/3 = 0.333 (equal)
Social
(social sustainability indicators)
9 (water)
13 (sludge)
= 1/6 = 0.167 (water)
= 1/13 = 0.077 (sludge) = 5/7 = 0.714 (maximum)
106
4. METHODOLOGY FOR ESTIMATING CONCENTRATIONS OF
PPCP COMPOUNDS IN WWTPS
Despite an increasing number of studies on the compounds from PPCPs, data on
their concentrations in the environment are still scant. This is due to many factors,
including great variability in usage and physicochemical properties of these chemicals,
which contribute to their widespread presence and complex behaviour, particularly in the
aquatic environment. Aiming to contribute to a better understanding of the role that
WWTPs play in the presence of PPCP substances in the environment, this Chapter
proposes a new method for estimating the expected concentrations of these compounds
in conventional WWTPs, their expected discharge and related concentrations in
freshwaters. The proposed method can assist with future ecotoxicological and
environmental risk assessments as well as the development of policies and regulation
related to PPCP compounds.
4.1. Estimation of Influent flow in WWTPs
As can be seen in Figure 26, the data for the influent flow Q and the served
population p range widely. For example, the smallest treatment facility has an average
flow of 7,200 m3/day and the largest 2,785,000 m3/day; the population served varies from
23,000 to 6.1 million. However, as indicated in Figure 26, the influent flow and the
population served are well correlated linearly (R2 = 0.9225). Based on these data, the
average per-capita influent flow q is equivalent to 428 L/inhab.day. It is acknowledge that
this estimate is greater than averages for European countries (of around 150-200
L/inhab./day) but in the range of US, Canada and Australia (400-500 L/inhab./day). Since
the data was gathered from these regions and Asian countries, combined to the fact that
many of the considered WWTPs presumably collects storm water (aquaterra 2008;
Sperling 2007), this value is assumed representative for this study. Further assessment on
this value or use of other average per-capita influent flow can be calculated accordingly
in future research. Anyhow, this value is used for the estimations of different parameters
in the next steps due to the fact it reduces variation in the concentration of these
substances since dilution factor is controlled.
107
Figure 26 – Correlation between daily water influent Q and population p served by WWTPs based on the
data in Table 1
4.2. Estimation of influx of PPCP compounds into WWTPs and removal rates
The annual per-capita influx IMinf,i into WWTPs of the target PPCP compounds
and their removal rates Rrate,i,, estimated using data in eqns. (5) and (6), are shown in
Figure 27 and Figure 28, respectively. As can be seen in Figure 27, the great majority of
the IMinf,i values fall between 1 and 100 mg/inhab.year, with only three being above 1,000
and three below 1 mg/inhab.year. Similarly, most of the removal rates Rrate,i,s in Figure 28
vary between 20% and 100%, with only a few falling below 20%. It can also be observed
that removal rates for some compounds have negative values – this is due either to their
accumulation (Gao et al. 2012; Li & Zhang 2011; Katsoyiannis & Samara 2005; Quintana
et al. 2005) or chemical reactions during the treatment process (Carballa et al. 2004;
Schlüsener & Bester 2008; Xu et al. 2012; Esperanza et al. 2007) which can lead to higher
concentrations in the effluent than in the influent (see topic 2.3.3.).
The estimated IMinf,i and Rrate,i values are then grouped respectively into the
datasets A (Table 36, SI) and B (Table 37, SI) for each target compound to determine the
interquartile values and the outliers. The latter are excluded from further consideration. It
can be noticed in Table 6 that there are only four outliers for IMinf,i, out of 85 data points
in total. All of these are for the WWTPs based in the UK, with one being shared with
Switzerland. Given that consumption of PPCPs in the UK is amongst the highest in the
R² = 0.9225
0.0E+00
5.0E+08
1.0E+09
1.5E+09
2.0E+09
2.5E+09
3.0E+09
0.0E+00 2.0E+06 4.0E+06 6.0E+06 8.0E+06
Wast
ewate
r in
flu
ent
Q (
L/d
)
Population served, p (no.)
108
world (WHO 2004; Roig 2010), this would suggest that higher consumption leads to their
higher influx into WWTPs (Lindqvist et al. 2005; Oosterhuis et al. 2013; Zhang & Geißen
2010).
To test this assumption further, the data for IMinf,i were analysed by world region
excluding the outliers to determine if there is a relationship between the influx of PPCP
compounds into the WWTPs and the consumption of PPCPs. First, the data were grouped
into the low, mid-range and high values. As shown in Table 7, out of 81 data points,
excluding the outliers, there is an equal number of low and high data values (14, with the
rest being in the mid-range. Of this, 65% of the low values are located in Asia and 29%
in Europe, the two regions for which the data are most abundant. For the high values, half
are in Europe and 36% in Asia. Therefore, a trend can be noticed with the low values
located in Asia and the high in Europe, corresponding to the respective PPCP
consumption in these regions (WHO 2004; Roig 2010).
Figure 27 – Annual per-capita discharge IMinf,i of target PPCP compounds estimated using eqn. (5) and
data from Table 1. Each point on the graph represents IMinf for one target compound i
1.E-01
1.E+00
1.E+01
1.E+02
1.E+03
1.E+04
1.E+05
IMin
f,i(m
g/i
nh
ab
.yea
r)
109
Figure 28 – Removal rates Rrate,i for target PPCP compounds estimated using eqn. (6) and data from Table
1. Each point in the graph represents Rrate,i for one target compound i
However, to account for the fact that the number of data is not evenly distributed
among the regions, they were first normalised with respect to the range of concentrations
and then weighted to consider the differing number of data available for different regions.
The results are shown in Figure 29; the method applied for the normalisation and
weighting can be found in section in SI, together with the range of IMinf,i values for each
of the 14 PPCP compounds considered (Figure 74, outliers in Table 6). The results in
Figure 29 demonstrate that indeed the lowest values are found for Asian WWTPs and the
highest in North America and Europe (although the highest values in Figure 29 are for
Australia, they are least reliable due to only two data points available). These trends are
congruent with the consumption of PPCPs in these regions, further corroborating the
assumption that the influx of PPCP compounds into WWTPs is correlated with their
consumption. For the removal rates (dataset B, outliers in Table 6), out of 142 data points,
10 are outliers, with the majority being for WWTPs in the UK and Spain. However, no
correlation is apparent between the number of outliers and the type of WWTP or
operational parameters despite a wide range covered by the data in the literature.
-260
-220
-180
-140
-100
-60
-20
20
60
100
Rra
te,i
(%)
110
Table 6 – Outliers for the influx of PPCP compounds (A dataset) and removal rates (dataset B) in
WWTPs (data points in SI Table 36 and Table 37)
Compound i WWTP location Dataset A, IMinf,i
(mg/inhab.year)
Total number
of data points
Dataset B, Rrate,i
(%)
Total number
of data points
Acetaminophen UK a 4 94.45; 95.00 10 Diclofenac UK 391.14 10 14
Ibuprofen Spain; UK 10 44.83; 64.05 18
Trimethoprim UK 260.40 8 13 Erythromycin South Korea; UK 5 -81.82; 80.00 8
Sulfamethoxazole UK 71.84 8 15
Metoprolol US 3 90.48 6 Gemfibrozil Spain 4 -133.33 7
Bezafibrate - 4 5
Carbamazepine UK 730.40 7 13 Oestrone - 7 10
17β-oestradiol Sweden 7 16.67 8
Triclosan 4 6 Caffeine Spain 4 42.86 9
a One of the values was much higher than the rest (see SI Table 36), but due to a small data sample, that value has not been considered
as an outlier.
Table 7 – Distribution of data for IMinf,i (dataset A) in different world regions
Values No. of points
North America Asia Europe Australia Total
Low 1 9 4 0 14
Mid-range 3 30 18 2 53
High 2 5 7 0 14
Total 6 44 29 2 81
Figure 29 – Normalized and weighted results for the number of data points for IMinf,i (dataset A) by world
region
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
Australia North America Europe Asia
No
rma
lize
d a
nd
weig
hte
d v
alu
es
111
4.3. Estimation of concentration ranges of PPCP compounds in WWTPs
4.3.1. Daily influx ranges
The expected range of the daily influx of the target compounds into WWTPs,
αrange,i,, estimated by eqn. (7), is given in Figure 30 (see SI Table 38 for the slopes values).
The figure shows minimum (Figure 30a) and maximum (Figure 30b) values, taking into
account the size of the population p served by WWTP. As can be seen, the expected daily
influx of PPCP compounds is correlated linearly with the size of the population. This is
in congruence with the assumption discussed in the previous section that a greater per-
capita consumption of PPCPs leads to a higher influx of their compounds into WWTPs.
For example, it can be inferred from Figure 30 that a WWTP serving 200,000 inhabitants
has an expected daily influx of acetaminophen in the range of 600 g/day to 10 kg/day
while that serving twice as many people can expect double the influx. The PPCP
compound with the highest estimated influx is acetaminophen, followed by ibuprofen and
caffeine. This is not surprising since all three products are available over the counter and
used widely. The lowest influx is found for the hormones 17β-oestradiol and oestrone.
112
a) Minimum daily influx
b) Maximum daily influx
Figure 30 – Minimum (a) and maximum (b) daily influx of target PPCPs estimated according to eqn. (7)
for different size of the population served by WWTPs
1
10
100
1000
10000
100000
10,000 100,000 1,000,000 10,000,000
Acetaminophen
Caffeine
Ibuprofen
Erythromycin
Triclosan
Carbamazepine
Trimethoprim
Diclofenac
Sulfamethoxazole
Bezafibrate
Oestrone
Gemfibrozil
Metoprolol
17β-oestradiol
Population served by the plant, p (no. of inhabitants)Est
imate
d m
inim
um
dail
y i
nfl
ux o
f com
pou
nd
s,α
min
,i
(g/d
)
1
10
100
1000
10000
100000
10,000 100,000 1,000,000 10,000,000
Acetaminophen
Ibuprofen
Caffeine
Carbamazepine
Erythromycin
Diclofenac
Triclosan
Bezafibrate
Trimethoprim
Sulfamethoxazole
Gemfibrozil
Metoprolol
Oestrone
17β-oestradiol
Est
imate
d m
axim
um
dail
y i
nfl
ux o
f co
mp
ou
nd
s,α
ma
x,i
(g/d
)
Population served by the plant, p (no. of inhabitants)
113
4.3.2. Influent concentration ranges
The influent concentration range of PPCP compounds in a WWTP, βrange,i,
calculated according to eqn. (8) for the average per-capita influent of 428 L/inhab.day, is
given in Table 8. As can be seen, the expected mean concentration for most compounds
ranges between 0.02 µg/L for 17β-oestradiol to 66.9 µg/L for acetaminophen. In the worst
case, the latter can reach 127 µg/L; the next worst are ibuprofen with 6.1 µg/L and
caffeine at 5.7 µg/L. The lowest minimum concentrations can be expected for metoprolol
and 17β-oestradiol (~ 0 µg/L or below detection levels).
Table 8 - Estimated influent concentration ranges for the target PPCPa
Compound i βrange,i (µg/L) V (βmax,i -βmin,i)
(µg/L) βmin,i βmean,i βmax,i
Acetaminophen 6.87 66.9 127 120
Diclofenac 0.06 0.52 0.97 0.91 Ibuprofen 1.04 3.57 6.10 5.06
Trimethoprim 0.08 0.16 0.24 0.16
Erythromycin 0.57 0.84 1.11 0.54 Sulfamethoxazole 0.03 0.11 0.20 0.17
Metoprolol 0.00 0.04 0.08 0.08
Gemfibrozil 0.02 0.09 0.17 0.15 Bezafibrate 0.03 0.19 0.35 0.31
Carbamazepine 0.10 0.69 1.29 1.19
Oestrone 0.03 0.05 0.06 0.03 17β-oestradiol 0.00 0.02 0.04 0.03
Triclosan 0.42 0.59 0.76 0.33
Caffeine 3.06 4.40 5.74 2.68 a q = 428 L/inhab.day.
4.3.3. Expected removal ranges
To estimate the expected concentration range γrange,i of PPCP compounds in the
WWTP effluent (eqn. (9)), it was first necessary to determine the expected range of
removal rates Rrange,i for each compound using dataset B. As shown in Figure 31 (and SI
Table 39), the expected removal rates vary greatly with the lowest removal (≤ 25%)
expected for erythromycin, metoprolol, carbamazepine and oestrone and the highest (>
90%) for acetaminophen, ibuprofen and caffeine. The lowest variation in the removal
rates (≤ 1%) was found for caffeine and acetaminophen, suggesting that their removal is
not dependent on the type of treatment or operating conditions of the plant.
114
For some compounds (gemfibrozil, carbamazepine and oestrone), negative
removal rates can be expected resulting in a higher concentration in the effluent than in
the influent into WWTP. As mentioned earlier during the literature review (see section
2.3.3.2), this is due to possible transformation, desorption, recombination, conjugation
and/or accumulation of the compounds during the secondary treatment (Gao et al. 2012;
Kagle et al. 2009; Koh et al. 2008; Verlicchi, Al Aukidy & Zambello 2012) However, the
variation in the removal rates could also be attributed to a wide variation in their
physicochemical properties which can impair their removal by conventional wastewater
treatment methods.
To better illustrate the above, carbamazepine, for instance, is already known to be
amongst the PPCP compounds with lowest sorption and biodegradability potential during
wastewater treatments, suitable as anthropogenic marker in aquatic environments (Clara
et al. 2004; Onesios et al. 2009; Ying et al. 2009). The eventual increased effluent
concentration of this substance was attributed by Zhang et al. (2008) to sampling and
measurements issues. Oestrone is also known as one of the most troublesome substances
during biological wastewater treatment, with several studies demonstrating its
unpredictable behaviour, specially relative to its irregular sorption potential, dependence
of the treatment’s oxidation conditions and microbial activity in the biological reactor
(Koh et al. 2008; Atkinson et al. 2012; Esperanza et al. 2007; Evgenidou et al. 2014).
Figure 31 – Estimated range of WWTP removal rates (Rrange,i) for the target compounds (q = 428
L/inhab.day)
-120
-100
-80
-60
-40
-20
0
20
40
60
80
100
120
Rem
ov
al
rate
s,R
ran
ge,
i(%
)
115
To examine the possible effects of this fact, some physicochemical properties of
the target compounds were considered in relation to their removal rates estimated here.
Among these, their acidity, measured by the acid dissociation constant (pKa), showed an
interesting trend (Table 9). The acidic compounds (low pKa) were found to have moderate
removal rates (30-62.5%) and higher removal variation ( > 35%), with ibuprofen being
the only exception. The basic compounds (high pKa) exhibited more extreme removal
rates (lower than 22.5% or higher than 80%) and lower variation ( < 30%). These
observations are in agreement with the findings of other authors related to the behaviour
of acidic PPCP compounds during biological treatment (Quintana et al. 2005; Metcalfe
et al. 2003; Thomas & Foster 2005). Furthermore, the compounds with extremely low
and high pKa (carbamazepine and oestrone, respectively), were found to be the exception
to the rule, with negative mean removal rates and the greatest removal variation (> 90%)
of all target compounds (see previous paragraph for more details).
Table 9 – Effect of acid dissociation constant (pKa) on estimated removal of PPCP compounds by
conventional WWTPs
Compound i Acid dissociation constant
(pKa)a
Mean removal rate, Rmean,i
(%)
Removal rate variation
(%)
Carbamazepine 2.30 -5.00 90.0 Bezafibrate 3.60 55.00 40.0
Diclofenac 4.20 40.00 80.0 Ibuprofen 4.90 93.50 11.0
Gemfibrozil 4.90 30.00 70.0
Sulfamethoxazole 5.70 62.50 35.0 Trimethoprim 7.10 42.50 65.0
Triclosan 8.10 89.50 15.0
Erythromycin 8.90 12.50 25.0 Acetaminophen 9.40 99.49 1.00
Metoprolol 9.60 22.50 25.0
17β-oestradiol 10.23 80.00 30.0 Caffeine 10.40 99.45 0.90
Oestrone 10.40 -25.00 150
a Source: Muñoz (2008).
4.3.4. Effluent concentration ranges
The expected concentration ranges γrange,i of PPCP compounds in the WWTP
effluent, estimated by using the influent concentration and the removal rate ranges (eqn.
(9)), are summarised in Table 10. The results suggest that, similar to the influent
concentrations, the minimum mean effluent concentration is expected for 17β-oestradiol
(0.01 µg/L). However, unlike the influent concentrations, the highest mean effluent value
is found for carbamazepine (0.99 µg/L); this is due to its negative removal rate in WWTP.
116
Table 10 – Estimated effluent concentration ranges for the target PPCP compoundsa
Compound i γrange,i (µg/L) V (γmax,i -γmin,i)
(µg/L) γmin,i γmean,i γmax,i
Acetaminophen 0.00 0.64 1.28 1.28
Diclofenac 0.01 0.49 0.97 0.96
Ibuprofen 0.01 0.37 0.73 0.72
Trimethoprim 0.02 0.12 0.21 0.19
Erythromycin 0.43 0.77 1.11 0.68
Sulfamethoxazole 0.01 0.06 0.11 0.10
Metoprolol 0.00 0.04 0.07 0.07
Gemfibrozil 0.01 0.09 0.17 0.17
Bezafibrate 0.01 0.12 0.22 0.22
Carbamazepine 0.06 0.99 1.93 1.87
Oestrone 0.02 0.07 0.13 0.11
17β-oestradiol 0.00 0.01 0.01 0.01
Triclosan 0.01 0.07 0.14 0.12
Caffeine 0.00 0.03 0.06 0.05
a q = 428 L/inhab.day.
4.3.5. Sludge concentration ranges
The concentration ranges Srange,i of the target PPCP compounds in the sludge from
WWTPs, calculated according to eqn. (10), are given in Table 11. Following the trend
for the influent concentrations, the highest mean concentration in the sludge is expected
for triclosan (3,528 µg/kg) and the lowest for sulfamethoxazole and metoprolol (both
around 1.0µg/kg). These results are in broad agreement with measurements available in
the literature, as shown in the review of Verlicchi & Zambello (2015).
For example, in the study of McClellan & Halden (2010), triclosan, erythromycin,
caffeine and ibuprofen were found at highest concentration in 94 wastewater treatment
plants in the district of Columbia, US. This is in accordance with this work results, where
these three compounds have the highest mean concentrations. Furthermore, although in
triclosan showed a much higher mean concentration in US WWTPs (average values
around 10,000 µg/kg), caffeine and ibuprofen were in the same range as here. The results
for the other PPCP compounds are also in agreement with the ranges obtained in the
literature (Sim et al. 2011; Yu & Wu 2012; J. Radjenović et al. 2009; Jones-Lepp &
Stevens 2007). The only exceptions are for the compounds trimethoprim, gemfibrozil and
carbamazepine which appear to be underestimated in this work, possibly due to their
recalcitrant behavior (see Figure 31).
117
Table 11 – Estimated sludge concentration ranges for the target PPCP compoundsa
Compound i LogKd,i
(L/kg)
Srange,i (µg/kg) Variation (Smax,i - Smin,i)
(µg/kg) Smin,i Smean,i Smax,i
Acetaminophen 0.3 13.70 133.40 253.11 239.41
Diclofenac 2.7 30.31 245.49 460.68 430.37 Ibuprofen 2.1 129.45 443.17 756.87 627.42
Trimethoprim 1.8 4.81 9.82 14.83 10.02
Erythromycin 2.2 88.65 130.49 172.32 83.67 Sulfamethoxazole 1.0 0.31 1.15 1.98 1.67
Metoprolol 1.3 0.10 0.88 1.66 1.56
Gemfibrozil 1.3 0.38 1.85 3.31 2.93 Bezafibrate - - - - -
Carbamazepine 1.2 1.52 10.94 20.35 18.84
Oestrone 2.5 9.76 14.64 19.52 9.76 17β-oestradiol 2.6 1.46 8.04 14.61 13.15
Triclosan 4.3 2,530.69 3,527.60 4,524.65 1,993.96
Caffeine 2.3 596.61 857.46 1,118.32 521.71 a For q = 428 L/inhab.day, and sDM=50 g/inhab.day (Rulkens 2007)
4.4. Estimation of freshwater concentrations of PPCP compounds
The expected mean concentration in freshwater bodies (PECmean,i), estimated
according to eqn. (11) and based on the mean concentrations γmean,i of PPCP compounds
in the WWTP effluent, are given in Figure 32. The figure shows the PECmean,i values for
different freshwater flows, ranging from 50 ML/day to 5 bn L/day and for effluent flows
from WWTP varying from 31.5-442 ML/day. The latter corresponds to the typical size
range of WWTP, serving 11.7 million inhabitants. For example, if a WWTP discharges
442 ML/day to a freshwater body with a 50 ML/day average flow, the mean expected
concentration of acetaminophen is around 580 ng/L. If a WWTP discharges 150 ML/day
to a body with a flow of 500 ML/day, the average PEC of acetaminophen is expected to
be around 150 ng/L. The best-fit curves given in Figure 32 can be used to estimate the
values of PECmean,i for each target PPCP compound over the flow ranges considered here.
It can be noted from the figure that the best-fit relationship between the PEC and the
volume of WWTP effluent changes with the freshwater flow: it is logarithmic for the
lower flow range (50-100 ML/day), polynomial for the mid-range (500 ML/day) and
linear for the highest flow (5 bn L/day).
118
Figure 32 – Predicted environmental concentration (PEC) of target PPCP compounds in freshwaters for the
mean expected effluent concentration (γmean,i) and for different freshwater flows: F1 = 5bn L/day; F2 = 500
M L/day; F3 = 100 M L/day; F4 = 50 ML/day
y = 1E-07x
R² = 0.999
y = -1E-15x2 + 1E-06x
R² = 0.9984
y = 142.68ln(x) - 2309.6
R² = 0.9981
y = 125.44ln(x) - 1902.8
R² = 0.9852
0
100
200
300
400
500
600
700
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
p x q (L/day)
PE
Cm
ean
(ng/L
)
Ace
tam
ino
ph
en
y = 9E-08x
R² = 0.999
y = -8E-16x2 + 9E-07x
R² = 0.9984
y = 109.7ln(x) - 1775.6
R² = 0.9981
y = 96.44ln(x) - 1462.9
R² = 0.9852
0
100
200
300
400
500
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Dic
lofe
na
c
PE
Cm
ean
(ng
/L)
p x q (L/day)
y = 7E-08x
R² = 0.999
y = -6E-16x2 + 7E-07x
R² = 0.9984
y = 82.621ln(x) - 1337.4
R² = 0.9981
y = 72.637ln(x) - 1101.8
R² = 0.9852
0
50
100
150
200
250
300
350
400
450
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Ibu
pru
fen
PE
Cm
ean
(ng/L
)
p x q (L/day)
y = 2E-08x
R² = 0.999
y = -2E-16x2 + 2E-07x
R² = 0.9984
y = 25.857ln(x) - 418.54
R² = 0.9981
y = 22.732ln(x) - 344.83
R² = 0.9852
0
20
40
60
80
100
120
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Tri
met
ho
pri
m
PE
Cm
ean
(ng/L
)
p x q (L/day)
y = 1E-07x
R² = 0.999
y = -1E-15x2 + 1E-06x
R² = 0.9984
y = 170.78ln(x) - 2764.3
R² = 0.9981
y = 150.14ln(x) - 2277.5
R² = 0.9852
0
100
200
300
400
500
600
700
800
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Ery
thro
myci
n
p x q (L/day)
PE
Cm
ean
(ng
/L)
y = 1E-08x
R² = 0.999
y = -9E-17x2 + 1E-07x
R² = 0.9984
y = 12.843ln(x) - 207.89
R² = 0.9981
y = 11.291ln(x) - 171.28
R² = 0.9852
0
10
20
30
40
50
60
70
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Su
lfa
met
ho
xa
zole
PE
Cm
ean
(ng/L
)
p x q (L/day)
y = 7E-09x
R² = 0.999
y = -6E-17x2 + 7E-08x
R² = 0.9984
y = 8.6813ln(x) - 140.52
R² = 0.9981
y = 7.6322ln(x) - 115.77
R² = 0.9852
0
5
10
15
20
25
30
35
40
45
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Met
op
rolo
l
p x q (L/day)
PE
Cm
ean
(ng/L
)
y = 2E-08x
R² = 0.999
y = -1E-16x2 + 2E-07x
R² = 0.9984
y = 20.194ln(x) - 326.87
R² = 0.9981
y = 17.754ln(x) - 269.31
R² = 0.9852
0
10
20
30
40
50
60
70
80
90
100
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Gem
fib
rozi
l
PE
Cm
ean
(ng
/L)
p x q (L/day)
y = 2E-08x
R² = 0.999
y = -2E-16x2 + 2E-07x
R² = 0.9984
y = 25.928ln(x) - 419.69
R² = 0.9981
y = 22.795ln(x) - 345.78
R² = 0.9852
0
20
40
60
80
100
120
140
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Bez
afi
bra
te
p x q (L/day)
PE
Cm
ean
(ng/L
)
y = 2E-07x
R² = 0.999
y = -2E-15x2 + 2E-06x
R² = 0.9984
y = 221.17ln(x) - 3580.1
R² = 0.9981
y = 194.45ln(x) - 2949.6
R² = 0.9852
0
200
400
600
800
1000
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Ca
rba
ma
zep
ine
p x q (L/day)
PE
Cm
ean
(ng/L
)
y = 1E-08x
R² = 0.999
y = -1E-16x2 + 1E-07x
R² = 0.9984
y = 16.027ln(x) - 259.42
R² = 0.9981
y = 14.09ln(x) - 213.74
R² = 0.9852
0
10
20
30
40
50
60
70
80
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Est
ron
e
p x q (L/day)
PE
Cm
ean
(ng/L
)
y = 1E-09x
R² = 0.999
y = -1E-17x2 + 1E-08x
R² = 0.9984
y = 1.5172ln(x) - 24.559
R² = 0.9981
y = 1.3339ln(x) - 20.234
R² = 0.9852
0
1
2
3
4
5
6
7
8
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
17β
-est
rad
iol
PE
Cm
ean
(ng/L
)
p x q (L/day)
y = 1E-08x
R² = 0.999
y = -1E-16x2 + 1E-07x
R² = 0.9984
y = 16.54ln(x) - 267.73
R² = 0.9981
y = 14.541ln(x) - 220.58
R² = 0.9852
0
10
20
30
40
50
60
70
80
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Tri
clo
san
p x q (L/day)
PE
Cm
ean
(ng/L
)
y = 6E-09x
R² = 0.999
y = -5E-17x2 + 5E-08x
R² = 0.9984
y = 6.7228ln(x) - 108.82
R² = 0.9981
y = 5.9104ln(x) - 89.656
R² = 0.9852
0
5
10
15
20
25
30
35
0.0E+00 1.0E+08 2.0E+08 3.0E+08 4.0E+08 5.0E+08
F1 F2 F3 F4
Ca
ffei
ne
PE
Cm
ean
(ng/L
)
p x q (L/day)
119
4.5. Chapter conclusions
This paper has proposed a new methodology for estimating expected
concentrations of PPCP compounds ubiquitously found in influents, effluents and sludge
of conventional WWTPs, as well as their expected concentrations in freshwaters.
Application of the methodology has been illustrated for 14 PPCP compounds for which
the data were available; however, the methodology is generic and can be applied to further
PPCP compounds or other emerging pollutants if and when the data become available
(and for new compounds not assessed in this study). The methodology and the results
from this work can be used for several purposes. First, the expected concentrations in
freshwaters could be used as a basis for ecotoxicological tests to help determine their
impact on aquatic species. Knowing how high the concentrations of compounds in
freshwater can also help identify likely synergistic effects between them and ranking
according the concentration in influents and effluents.
Secondly, the results could assist in environmental risk assessment (ERA) by
linking consumption of PPCPs with environmental concentrations taking into account the
actual measured data, rather than relying solely on production or consumption data for
PPCPs. Furthermore, the outputs could be used for development of policy and regulations
as currently the presence of PPCP compounds in the environment is not regulated. For
example, regulation could impose limits on the concentrations of these compounds in
WWTPs effluents, also determine the necessity for monitoring the effluents for the
presence of PPCP compounds. This is important not only because of the environmental
pollution but also due to the increasing pressure on traditional water resources associated
with pollution, urbanization and climate change, which is necessitating reuse of
wastewater in many regions worldwide. As a result, legislation to limit the presence of
PPCP compounds in wastewaters intended for reuse as potable water has been considered
in some regions. For example, California has recently introduced regulations for
monitoring of some PPCP compounds in wastewaters intended for reuse (NRC 2012;
EPA 2012).
120
Depending on the intended wastewater reuse, the adoption of advanced treatment
techniques in wastewater treatment plants may be necessary in future to aid the removal
of PPCP compounds. Thus, the methodology proposed in this work could also be applied
to estimate the concentrations that such plants should expect in their influents from
conventional treatment and the removal rates that they should achieve to render the reused
water safe for human health. This in turn could aid the selection and design of most
effective advanced treatment plants to enable wastewater reuse.
Furthermore, environmental legislation for the traditional sewage sludge handling
routes, such as agricultural spreading, is becoming increasingly more stringent, with some
PPCP compounds already monitored in some European countries (Ellis 2006; Moran &
Dann 2008; Roig 2010). Therefore, the results of this research could also be helpful for
these purposes, helping to determine the expected concentrations in the sludge and set the
appropriate legislative limits.
121
5. LIFE CYCLE ASSESSMENT OF WATERWATER TREATMENT
TECHNIQUES
Compounds from Pharmaceutical and Personal Care Products (PPCPs) are of
increasing interest because of their ecotoxicological properties and environmental
impacts. Wastewater Treatment Plants (WWTPs) are the main pathway for their release
into the environment due to the inefficiency of conventional WWTPs in removing these
contaminants from effluents. Therefore, different advanced wastewater treatment
techniques have been proposed for their treatment. However, it is not known at present
how effective these treatment methods are and whether on a life cycle basis they cause
other environmental impacts which may outweigh the benefits of the treatment. In an
effort to provide an insight into this question, this paper considers life cycle
environmental impacts of the following advanced treatment techniques aimed at reducing
freshwater ecotoxicity potential of PPCP compounds: granular activated carbon (GAC),
nanofiltration (NF), solar photo-Fenton (SPF) and ozonation. This Chapter depicts
information related to methodology for assessing variations in the operating parameters
of the treatment techniques (section 5.2). It also presents the removal estimation of PPCP
compounds by these same treatments and consequently their confrontations in effluents
after the advanced treatments. After the life cycle impact results (section 5.3-5.5), there
is a discussion about their potential for wastewater reuse (section 5.6) and the chapter
conclusions (section 5.7).
5.1. Goal and scope
The goal of the study was to estimate and compare life cycle environmental
impacts of the four advanced techniques for treatment of PPCP compounds. A further
goal was to estimate the ecotoxicity of PPCP compounds in the effluent after advanced
treatment and compare it to the equivalent impact of the effluent from conventional
wastewater treatment plants (WWTP) without the advanced treatment. The system and
the system boundaries are outlined in Figure 33. As indicated in the figure, the scope of
the study is from ‘cradle to grave’, encompassing construction, operation and
decommissioning of the treatment plant. The advanced wastewater treatment techniques
are assumed to be coupled to a conventional WWTP with membrane bioreactors (MBR)
(see topic 2.5.3.).
122
The assumed capacity of MBR plant is 64,000 m3/d, corresponding to the average
capacity of WWTPs in the UK (DEFRA 2012). Furthermore, it is assumed that the
advanced wastewater treatment methods are capable of treating secondary effluents to the
drinking water standards since they are controlled for pH, pathogens, hardness and heavy
metals. The functional unit was defined as “treatment of 1,000 m3 of secondary effluent”.
The facility is assumed to be located in the UK. The lifetime of the plants is assumed at
60 years.
Figure 33 – System boundaries and life cycle stages of the advanced wastewater treatment techniques
considered in the study (*Excluded for ozonation due to a lack of data)
5.2. Inventory analysis
5.2.1. Overview of advanced wastewater treatments operating parameters
The inventory data were sourced from the literature and own estimates as detailed
in the next sections. Industry data were not available as the removal of PPCP compounds
is still not targeted by the water industry due to a lack of legislation. The life cycle data
were taken from Ecoinvent 2.2 (Frischknecht et al. 2004). The following sections give a
brief description of the advanced treatment methods considered, followed by an overview
of the estimation of the boundaries defined for the operating parameters of the advanced
wastewater treatment techniques.
Decommissioning*Waste treatment
and disposal
Energy
Chemicals
Other
materials
Influent
(from secondary treatment
Transport
Effluent
Transport
Plant operation
(PPCP treatment)
Construction* Part replacements*
123
5.2.1.1. Granular activated carbon (GAC)
GAC treatments removes PPCP compounds by physical adsorption onto the GAC
bed and to a lesser extent through biodegradation, thus avoiding generation of harmful
reaction by-products. Moreover, a high removal of metals is expected (Goel et al. 2005;
Qu et al. 2013). After a certain time in use, the bed needs to be regenerated or replaced
(Figure 37). Hence, the key influencing parameters are the amount of fresh GAC required
for the treatment and the number of regeneration cycles before the bed needs to be
replaced, as depicted in section 3.2.1.1.
The assumptions for the variables in eqns. (12)-(15) are summarised in Table 12,
corresponding to a maximum number of bed regenerations (nmax) of 10. The necessity of
defining a maximum number of regenerations for the granular activated carbon in the
contactors is to maintain its adsorption capacity (e.g. porous structure and reactivity), and
by guaranteeing that over half of the activated carbon in the beds are being regenerated
less than five times after ten bed replacements (since at every regeneration there is the
addition of 10% of the fresh carbon due losses during the process), as shown in Figure 35
(San Miguel et al. 2001; Bayer et al. 2005). Moreover, it provides an estimate for
optimum environmental-economical replacement periods during the treatment life cycle.
As can be seen, three different EBCTs and bed service times were considered,
based on the range of values reported for large GAC treatment facilities (Wang et al.
2005; Reungoat et al. 2011; Clements 2002). The trends for the fresh and regenerated
GAC requirements according to different bed service times are given in Figure 34; the
actual values can be found in Table 12. As expected, the required amount of GAC
decreases with the increasing bed service time. These results were used in LCA to
determine the influence on the environmental impacts of the variation in the key
parameters.
124
Figure 34 – Estimated amounts of fresh and regenerated granular activated carbon for 1,000 m3 of
wastewater treated for different bed service times and empty-bed contact times (EBCT) (nmax:10, mloss:10%,
GAC density: 564 kg/m3)
Figure 35 – Amount of fresh and regenerated granular activated carbon in contactors according the
number of bed regenerations (mloss:10%)
0
20
40
60
80
100
120
140
160
180
0
5
10
15
20
25
30
35
40
100 140 180 220 260 300 340
Minimum of fresh GAC
This work (fresh GAC)
Maximum of fresh GAC
Minimum of regenerated GAC
This work (regenerated GAC)
Maximum of regenerated GAC
Bed service time (days)
Fresh
GA
C (
kg/1
,00
0m
3)
Regen
era
ted
GA
C (
kg/1
,00
0 m
3)
0%
10%
20%
30%
40%
50%
60%
70%
80%
90%
100%
0 1 2 3 4 5 6 7 8 9 10
No regenerations 1 regeneration 2 regenerations 3 regenerations
4 regenerations 5 regenerations 6 regenerations 7 regenerations
8 regenerations 9 regenerations 10 regenerations
Total of regenerations
Reg
en
era
ted
GA
C in
th
e c
on
tacto
rs
125
5.2.1.2. Nanofiltration (NF)
NF treatments uses membranes to remove PPCP compounds from the effluent
(Figure 37). First, the wastewater is passed through pre-filters under a high pressure to
remove larger particles and thereafter to the membranes with the pore sizes of 0.10 to 1.0
nm, separating the influx into the permeate (treated effluent) and concentrate
(contaminants, here assumed redirected to the WWTPs influent). as depicted during
section 3.2.1.2. The data for NF were taken from real facilities in Canada and France
(Bonton et al. 2012; Cyna et al. 2002).
5.2.1.3. Solar photo-Fenton (SPF)
SPF treatments are advanced oxidation processes, known for their high efficiency
in degrading most organic contaminants and simple operation. The process consists of
adding a catalyst and hydrogen peroxide to the influent which is then passed through
reactors irradiated by solar light to generate OH radicals and oxidize PPCP contaminants
(Figure 37), as depicted during section 3.2.1.3. The data for the operation of SPF
treatments are still subject of ongoing research, but some values are based on pilot-plants
(Robert & Malato 2002; Lofrano 2012; Klamerth 2011) which were considered here.
5.2.1.4. Ozonation
The ozonation treatment (disinfection), works through direct and indirect
reactions of PPCP compounds with OH radicals generated by ozone decomposition in the
contactors (Figure 37). The overall treatment efficiency is directly dependent on the
influent pH and organic matter content. After treatment, the effluent needs to be balanced
by the addition of NaOH. The main parameters that need to be considered in the design
of ozonation units are the amount of ozone required for efficient treatment and electricity
consumption, as depicted during section 3.2.1.4.
126
The range of transferred ozone dosages and the transfer ozone efficiencies
considered here can be found in Table 12. These values were chosen to match the ranges
found in the literature (Xu et al. 2002; Burns et al. 2007; Wang et al. 2005; Petala et al.
2006). The electricity consumption in the ozonation treatment include production of
ozone, pumping, recirculation and destruction of residual ozone. These were estimated as
being directly proportional to the amount of wastewater treated as in eqns. (16)-(17).
Based on the data in Table 12, the transferred ozone dosage (T) was estimated in
the range from 4.0 to 42 mg/L (Wang et al. 2005; Xu et al. 2002; Tripathi & Tripathi
2011) The average electricity consumption for ozone generation (Eozone) from
atmospheric air was estimated at 16.5 kWh/kgozone (Wang et al. 2005). Thus, the estimated
electricity consumption for ozonation (EOzonation) ranges from around 90 to 2,780
kWh/1,000 m3 for the transfer efficiency TE between 25% and 75% (Figure 36). This
variation is due to the strong dependence of the TE on the reactor type and size, influent
composition and the required level of treatment (disinfection).
Figure 36 – Estimated electricity consumption per 1,000 m3 of wastewater for different ozone transfer
efficiencies and applied ozone dosage
0
300
600
900
1,200
1,500
1,800
2,100
2,400
2,700
3,000
4.0 23.0 42.0
Transfer efficiency 75%
Transfer efficiency 50% (this work)
Transfer efficiency 25%
Applied ozone dosage (mg/L)
Ele
ctri
city
con
sum
pti
on
(k
Wh
)
127
Table 12 – Operating parameters for GAC (eqns. (12)-(15)) and ozonation (eqns. (16)-(17)) per 1,000 m3
of wastewater
GAC
EBCT
(min)
Qinf
(m3/h)
VGAC
(m3)
mGACa
(kg)
tGACb
(d) NBR
c nr Fresh GACd
(kg)
Regenerated
GAC (kg)
20
2,667
889 501,396 330 6 6 6,818,986 33,092,136
30 1,334 752,094 220 9 9 14,966,671 74,457,306
40 1,778 1,002,792 110 19 9 40,011,401 199,555,608
Ozonation
T
(kg/m3)
TE
(%)
DOZONE
(kg/m3)
Vinf
(m3)
EOZONE
(kWh/kgozone)
EOzonation
(kWh)
0.004 75 0.005
1,000
88
759
2772
0.023 50 0.046 16.5
0.042 25 0.168
a GAC density: 564 kg/m3
b Values from Figure 34
c nmax = 10; Ttreatment: 21,900 days (over the 60-year lifespan)
d mloss = 10% (Clements 2002)
Figure 37 – Graphical illustration of the advanced treatment methods considered in the study
Granular activated carbon Nanofiltration
Solar photo-Fenton
GAC production
Contactors
GAC regeneration
Hard coal
Influent Effluent
Chemicals
Landfill
Membrane
materials and
assembly
Nanofiltration
module
Cleaning
agent
Influent Effluent
Chemicals
Incineration
Solar panelInfluent Effluent
Catalyst and
hydrogen peroxide Precipitates Landfill
Ozonation
ContactorsInfluent Effluent
Ozone
generator
Ozone
destruction
Sodium
hydroxide
Chemicals
128
5.2.2. Estimation of removal rates of target PPCP compounds
The concentrations of PPCP compounds in the effluents of WWTPs vary greatly
in the literature (see Table 10). The similar is expected to be true for their removal rates
by advanced wastewater treatments, with only few data available from experimental
results. For that reason, it was necessary to estimate the potential removal rates of the
target compounds for different ranges of the operating parameters defined earlier,
defining representative values for the removal of PPCP substances by the advanced
wastewater treatment techniques. This was calculated according the presented in section
3.2.2, and the experimental conditions for each technique used for estimation of the final
concentration of PPCP compounds after advanced treatment is shown in Table 13.
Table 13 – Original data of the advanced wastewater treatment techniques operation
Treatment Experimental conditions Concentration of PPCP
compounds in the influent of
advanced treatment plants
Other
parameters
Source
Granular activated
carbon
Pilot scale
Acticarb BAC GA1000N
Apparent density: 554 kg/m3
Bed service time: 120 days
Filtration rate: 1.6 m/h Empty bed contact time: 60
min
Bed depth: 3 m Temperature: 26 °C
Diclofenac ~ 0.21 µg/L
Ibuprofen ~ 0.15 µg/L
Trimethoprim ~ 50 µg/L Carbamazepine ~ 0.50 µg/L
Dissolved organic
carbon: 8.7 mg/L
Dissolved oxygen: 5 mg/L
Reungoat et al.
(2011)
Nanofiltration Bench scale
Dow FilmTec NF270-400 Thin polyamide membrane
Applied pressure: 680 kPa
Molecular weight cut off: 400 Daltons
Zeta potential: -87 mV
Contact angle: 29.8° Temperature: 21 °C
Sulfamethoxazole ~ 0.72 µg/L
Carbamazepine ~ 0.68 µg/L Oestrone ~ 0.55 µg/L
17β-oestradiol ~ 0.52 µg/L
Dissolved organic
carbon: 3.7 mg/L Total organic
carbon: 3.7 mg/L
Comerton et al.
(2005)
Solar
photo-Fenton
Pilot scale
Applied pressure: 300 kPa
pH: 2.9 Influent flow: 1.5 m3 / h
Hydrogen peroxide dosage:
43 mg/L Iron salt dosage: 5.0 mg/L
Hydraulic retention time:
90 min Irradiation time: 30 min
Temperature: 30 °C
Diclofenac ~ 5.0 µg/L
Ibuprofen ~ 5.0 µg/L
Sulfamethoxazole ~ 5.0 µg/L Carbamazepine ~ 5.0 µg/L
Triclosan ~ 5.0 µg/L
Dissolved organic
carbon: 20 mg/L
Total organic carbon: 10 mg/L
Klamerth
(2011)
Ozonation Bench scale Contact time: 5 min
Applied ozone dosage: 14
mg/L Transfer efficiency: 14.5%
Transferred ozone dosage:
2 mg/L Ozone consumption: 1.6
mg/L
Temperature: 20 °C
Compounds not discriminated Minimum ~ 0.002 µg/L
Maximum ~ 0.774 µg/L
Dissolved organic carbon: 3.1 mg/L
Kim & Tanaka (2011)
129
The physicochemical properties selected for the estimation of the removal rates
are those that are compatible with the main removal mechanisms of each treatment:
hydrophobic interactions for GAC (Wang et al. 2005), sieving for NF (Nghiem & Hawkes
2007) and oxidation for SPF and ozonation (Huber et al. 2003). These parameters are
shown in Figure 38, along with the corresponding best-fit curves for the removal rates.
The results suggest that for GAC the physicochemical property best-fitting (i.e.
determining) the removal rates are the octanol-water partition coefficient (KOW); for NF,
it is the molecular weight (MW) of the target compounds and for SPF and ozonation,
hydroxyl radical reaction in air (HRA).
The removal efficiency estimated for the target compounds using the best-fit
curves in Figure 38 can be found in Table 14. As shown, GAC has the highest removal
efficiency across the target compounds, ranging from 89%-99%. Ozonation is the next
most efficient treatment method with the range of 80%-99%, except ibuprofen and
triclosan, for which the removal efficiency is 42% and 52%, respectively. NF and SPF
have similar average efficiencies of 62% and 67%, respectively. The highest removal
efficiency is found for erythromycin: 86% for NF and 99% for the other three methods.
However, the removal rates for the other target compounds are quite variable. Based on
the above, the concentration ranges for each target PPCP compound after the advanced
treatments were estimated as given in Table 15.
Figure 38 – Best-fit curves for the estimation of the removal rates of the target PPCP compounds by the
advanced treatment techniques based on experimental data in the literature. Data points include some non-
target compounds to improve the reliability of the estimates
y = -2.5516x2 + 16x + 76.361
R² = 0.70140
20
40
60
80
100
120
-1.0 0.0 1.0 2.0 3.0 4.0 5.0 6.0
Log Kow
Rem
ov
al
(%)
Granular activated carbon
y = 4E+21x2 - 9E+11x + 86.962
R² = 0.5020
20
40
60
80
100
120
1.0E-11 6.0E-11 1.1E-10 1.6E-10 2.1E-10
HRA (cm3/molec.sec)
Rem
ov
al
(%) Solar photo-Fenton
y = -2E+41x4 + 1E+32x3 - 3E+22x2 + 3E+12x + 11.103
R² = 0.8125
0
20
40
60
80
100
120
1.0E-11 6.0E-11 1.1E-10 1.6E-10 2.1E-10
HRA (cm3/molec.sec)
Rem
ov
al
(%)
Ozonation
y = 0.0009x2 - 0.8389x + 217.36
R² = 0.6203
0
20
40
60
80
100
120
100 200 300 400 500 600 700 800
MW (g/mol)
Rem
ov
al
(%) Nanofiltration
130
Table 14 – Estimated efficiencies for the removal of the target PPCP compounds in the advanced treatment
plants (%)
Granular
activated
carbon (%)
Nanofiltration
(%)
Solar
photo-Fenton (%)
Ozonation
(%)
Diclofenac 97 48 47 90
Ibuprofen 99 83 77 42 Trimethoprim 89 50 69 95
Erythromycin 99 86 99 99
Sulfamethoxazole 89 63 67 90 Carbamazepine 99 69 40 99
Oestrone 99 56 75a 80b
17β-oestradiol 99 56 75a 80b
Triclosan 94 50 74 52 a Feng et al. (2005) - pH 3.0; aqueous solution; hydrogen peroxide: 34 mg/L; iron salt: 0.59 mg/L; irradiation time 150 min;
oestrone removal assumed similar. b Ternes et al. (2003) - pH 7.2; secondary effluent; dissolved organic carbon: 23.0 mg/L; applied ozone dosage: 5.0 mg/L; 17β-oestradiol removal assumed as similar
Table 15 – Estimated concentrations of target PPCP compounds in effluents after the advanced wastewater
treatment (µg/L)
Compound
Granular
activated carbon Nanofiltration
Solar
photo-Fenton Ozonation
Min Mean Max Min Mean Max Min Mean Max Min Mean Max
Diclofenac 0.0000 0.0162 0.0321 0.0000 0.2555 0.5058 0.0000 0.2579 0.5106 0.0000 0.0490 0.0970
Ibuprofen 0.0001 0.0037 0.0073 0.0017 0.0644 0.1270 0.0023 0.0855 0.1686 0.0058 0.2128 0.4199
Trimethoprim 0.0023 0.0136 0.0237 0.0101 0.0604 0.1057 0.0062 0.0371 0.0649 0.0010 0.0060 0.0105
Erythromycin 0.0043 0.0077 0.0111 0.0582 0.1042 0.1503 0.0043 0.0077 0.0111 0.0043 0.0077 0.0111
Sulfamethoxazole 0.0011 0.0068 0.0124 0.0037 0.0224 0.0411 0.0033 0.0198 0.0363 0.0010 0.0060 0.0110
Carbamazepine 0.0006 0.0099 0.0193 0.0184 0.3030 0.5908 0.0358 0.5903 1.1509 0.0006 0.0099 0.0193
Oestrone 0.0002 0.0007 0.0013 0.0087 0.0306 0.0568 0.0050 0.0175 0.0325 0.0040 0.0140 0.0260
17β-oestradiol 0.0000 0.0001 0.0001 0.0000 0.0044 0.0044 0.0000 0.0025 0.0025 0.0000 0.0020 0.0020
Triclosan 0.0006 0.0039 0.0079 0.0050 0.0351 0.0701 0.0026 0.0185 0.0371 0.0048 0.0336 0.0672
5.2.3. Other inventory data
The inventory data for the construction, operation and decommissioning of the
advanced treatment plants over their lifespan are detailed in Table 16. The data for GAC
and NF are based on full-scale facilities treating 2,000 m3/d (Bonton et al. 2012). For
SPF, data from a pilot-scale study treating 7 m3/d were used (Ortiz 2006). For the latter,
it was necessary to scale up the plant to estimate the amount of materials used in the
construction of an industrial-size plant. For ozonation, no construction or
decommissioning data were available and these are thus excluded from consideration (see
Figure 33). Regarding the operating data, for each treatment method, minimum, mean
and maximum values were considered for each key parameter discussed (for details, see
Table 16 and topic 5.2.1.). The data for the GAC production and regeneration processes
were based on Bayer et al. (2005) and Jeswani et al. (2015).
131
The electricity consumption, cleaning agents (ethylenediaminetetraacetic acid –
EDTA and sodium hydroxide), and nanofiltration membrane lifespan (estimate to be
around 10 years) used in NF were based on the information reported in Bonton et al.
(2012) and Cyna et al. (2002). For SPF, the dosage of catalyst (iron salts) and hydrogen
peroxide were based on the studies by Ortiz (2006), Trovó et al. (2013) and Klamerth
(2011). Sodium hydroxide used for alkalinity balancing in ozonation was estimated from
Muñoz et al. (2007). For further information on these treatments please refer to SI Table
40. The worn-out parts were assumed to be replaced after 15 years. After the end of their
useful lifetime, supposed 60 years, the plants were assumed to be decommissioned and
waste treated using current waste management practices in the UK for recycling and
landfilling of construction materials (BRE/DEFRA 2010).
Table 16 – Inventory data for the advanced wastewater treatment techniques (per 1,000 m3 of secondary
effluent)
Ecoinvent data Granular
activated carbon
Nanofiltration Solar
photo-Fenton
Ozonation Unit
Construction and part replacements
Steel, low-alloyed 0.0523 0.0063 kg
Reinforcing steel 0.4150 0.0901 0.0496 kg
glass fibre 0.0175 0.0188 kg Concretea 0.0008 0.0002 m3
Polyvinyl chloride 0.1200 kg
Chromium steel 18/8 0.0125 kg Aluminium, production mix 0.0154 kg
Section bar extrusion,
aluminium
0.0154 kg
Anodising, aluminium sheet 0.0087 m2
Glass tube, borosilicate 0.0106 kg
Operation
Activated carbon production
(min/mean/max)
Membrane filtration
(min/mean/max)
Catalyst (min/mean/max)
Ozone generation
(min/mean/max)
Hard coal supply mix 15/33/66 kg Hard coal, burned at industrial
furnace 1-10 MW
304/669/1338 MJ
Natural gas, burned in industrial furnace >100 kW
66/145/290 MJ
Water, deionized, at plant 60/132/264 kg
Electricity, medium voltage, at grid (Germany)
8.0/17.6/35.2 kWh
Electricity, medium voltage, at
grid (UK)
270/412/554 150/750/1300 kWh
Iron sulphate, at plant 13.62/34.06/54.50 kg
Regeneration
(min/mean/max)
Cleaning
agents
(min/mean/max)
Hydrogen
peroxide
(min/mean/max)
Hard coal, burned at industrial
furnace 1-10 MW
75/165/330 MJ
Natural gas, burned in industrial furnace >100 kW
260/572/1144 MJ
Steam, for chemical process 15/33/66 kg
Electricity, medium voltage, at grid (UK)
0.75/1.65/3.30 kWh
EDTA 0.164/0.250/0.336 kg
Sodium hydroxide, 50% H2O 0.164/0.250/0.336 kg Hydrogen peroxide, 50% H2O 20/110/200 kg
Other operational data
Electricity, medium voltage, at
grid (UK)
19.56 0.42 kWh
Sodium hydroxide, 50% H2O 60.0 80.0 80.0 kg
Aluminum sulphate, powder 80.0 kg
132
Acrylonitrile-butadiene-styrene copolymer
0.30 kg
Carbon dioxide, liquid 14.0 31.0 kg
Calcium hydroxide 7.00 31.0 kg Phosphoric acid, industrial
grade
1.10
Sulphuric acid, liquid 36.0 130 kg Chlorine, liquid 0.60 0.60 kg
Spiral wound modulesb 0.3584 kg
Tap water, at user 2,200 kg Cement, hydrated, 0% water,
to residual material landfill
46.06 kg
Decommissioningc
Disposal steel, 0% water, to
inert material landfill
0.0234 0.0048 0.0025 kg
Disposal, inert waste, 5% water, to material landfill
0.0175 0.0188 kg
Disposal, concrete, 5% water,
to inert material landfill
0.4500 0.0900 kg
Disposal, bitumen, 1.4% water,
to sanitary landfill
5/11/22 kg
Disposal, polyethylene, 0.4% water, to sanitary landfill
0.0204 kg
Disposal, aluminium, 0%
water, to sanitary landfill
0.0015 kg
Disposal, glass 0% water, to
inert material landfill
0.0106 kg
Disposal, plastics, 15.3% water, municipal incineration
0.3584 kg
Steel – recycled 0.4439 0.0914 0.0471 kg
Concretea – recycling 1.4250 0.2850 kg Polyethylene – recycling 0.0996 kg
Aluminium – recycling 0.0293 kg
Chromium steel 18/8 – recycling
0.0119 kg
Transportd
Transport lorry, 16-32 t, Euro
5
44/57/81 19.8e 58/80/102 16 t.km
a Concrete density: 2,300 kg/m3 b See Table 41 in SI for details
c Concrete: 24% recycled and 76% landfilled; glass: 100% landfilled; glass fiber: 100% landfilled; metals: 95% recycled and 5% landfilled; plastics: 83% recycled and 17% landfilled d All distances were set to 200 km except for fresh GAC transport to the wastewater treatment site, assumed at 1,000 km (imported
from central Germany) e Negligible variation
5.3. Life cycle impacts results and discussion
The ReCiPe 2008 method (Goedkoop et al. 2009) was used to estimate the
environmental impacts of the advanced PPCP treatment options. All eighteen impacts
included in ReCiPe are considered here as discussed in the next section. In addition,
freshwater ecotoxicity potential of nine PPCP compounds was estimated using the
USEtox methodology (Rosenbaum et al. 2008; Henderson et al. 2011) to find out if the
treatment reduces the ecotoxicity associated with PPCP compounds on a life cycle basis.
Gabi 6.0 (thinkstep 2015) was used for LCA modelling and estimating the impacts. The
LCA results are first presented for the mean operating parameters (see Table 12) for each
impact in turn. The overview of these results is given in Figure 39 and Figure 40, where
the error bars represent the results for the minimum and maximum values of the
parameters, discussed in section 5.4. As can be seen from Figure 40, the majority of the
impacts across all the treatment techniques are from the operation of the plants.
133
Climate change potential
The results in Figure 39 suggest that GAC and SPF have a similar impact on
climate change, with the mean values estimated at 252 and 248 kg CO2 Equiv./1,000 m3,
respectively. The next best option is NF with 316 kg CO2 Equiv. At 543 CO2 Equiv./1,000
m3, ozonation is the worst alternative, with around two times higher impact than GAC.
For the latter, 41% of the impact is due to the production of fresh activated carbon and
26% due to the energy used for its regeneration, with the rest being due mostly to
aluminium sulphate production. Since fresh GAC is imported from Germany, 4% of the
impact is due to road transport.
In NF, 77% of the impact is from electricity generation and the remainder from
the productions of liquid carbon dioxide (used for fouling control) and calcium hydroxide
(for effluent balancing). For SPF, almost half (47%) of the impact is due to hydrogen
peroxide production and another 47% from other operational activities, with the rest being
from transport. The majority of the climate change potential of ozonation is from
electricity (83%) with the rest being due to sodium hydroxide production.
Resource depletion potential – fossil fuels and metals
As can be seen in Figure 39, GAC, NF and SPF have similar fossil resource
depletion potentials (89, 84 and 81 kg oil Equiv./1,000 m3, respectively) while ozonation
has nearly two times higher impact (155 kg oil Equiv.). For GAC, 50% of the impact is
related to the production of fresh activated carbon, 24% to energy used for regeneration
and 22% to the treatment process (Figure 40). For NF and ozonation, electricity
consumption is the main contributor to the depletion of fossil resources (86% and 83%,
respectively). The lowest metal depletion potential was found for GAC and NF (4.9 and
5.6 kg Fe Equiv./1,000 m3, respectively) and the highest for SPF (22.9 kg Fe Equiv.). For
GAC, the credit for materials recycling after decommissioning reduces the impact by
10%. For SPF, the majority of the impact is due to hydrogen peroxide (39%), sodium
hydroxide (29%) and sulphuric acid (21%).
134
Water depletion potential
The highest water consumption was obtained for ozonation, estimated at 1,296 m3
per 1,000 m3 of water treated (Figure 39). Therefore, more water is consumed along the
life cycle than treated. A half of this is due to water consumption during electricity
generation and another half from sodium hydroxide production. SPF is the second worst
option with 1,147 m3/1,000 m3. By contrast, GAC and NF consume roughly three times
less water. Therefore, these results demonstrate that to have a positive net generation of
water during advanced wastewater treatment the latter two alternatives are the only
recommended for contributor to the increase of water availability.
Ozone depletion potential
NF is the best option for this impact, followed by GAC (and 10.2 and 16.2 mg
CFC-11 Equiv./1,000 m3, respectively). The highest impact is from the SPF treatment (23
mg CFC-11 Equiv./1,000 m3), the majority of which (59%) is due to hydrogen peroxide
production; the contribution of transport is also relevant for this treatment option (11%).
For ozonation, 2/3rds of the impact, estimated at 17 mg CFC-11 Equiv./1,000 m3, originate
from electricity generation and 1/3rd from sodium hydroxide production.
Eutrophication potential – freshwater and marine
Ozonation has the highest freshwater eutrophication potential, equal to 0.22 kg P
Equiv./1,000 m3, with the main contributor (~60%) being phosphate emissions to
freshwater related to electricity generation. The equivalent impact for the other
alternatives ranges from 0.10 and 0.16 kg P Equiv./1,000 m3, almost exclusively due to
the operation of the treatment facilities. Regarding marine eutrophication, all the options
are fairly similar, with GAC and ozonation having a slightly higher impact than NF and
SPF. Unlike other impacts, the contribution to this category in GAC is dominated by
disposal of activated carbon (58%), due to the organically-bound nitrogen.
135
Acidification potential – terrestrial
For this impact, SPF is the worst option (2.5 kg SO2 Equiv./1,000 m3), largely due
to SO2 emissions from sulphuric acid production (66%). The next worst option is
ozonation, with 1.8 kg SO2 Equiv./1,000 m3, of which 53% is from SO2 emissions from
electricity generation. NF is the best option at 1.3 kg SO2 Equiv./1,000 m3, followed
closely by GAC with 1.5 kg SO2 Equiv./1,000 m3 (Figure 39).
Ionizing radiation potential
As can be seen in Figure 39, GAC has the lowest ionizing radiation potential (43
kg U235 Equiv./1,000 m3). This is almost seven times lower than for ozonation (270 kg
U235 Equiv./1,000 m3) which represents the worst option for this impact. SPF is the
second-best option, followed by NF. For all the options, the source of ionizing radiation
is nuclear power present in the electricity mix used for the operation of the plant.
Ecotoxicity potential – freshwater, marine and terrestrial
All three types of ecotoxicity exhibit a similar trend, with NF being the best option
and SPF the worst (Figure 39). For example, freshwater and marine ecotoxicity potentials
of SPF are more than three times higher than those of NF, largely because of emissions
to water of copper, nickel and zinc associated with the production of hydrogen peroxide
and sodium hydroxide. For freshwater and marine toxicity, the main contributor to the
impacts is the operation of the plants (Figure 40). However, terrestrial ecotoxicity is
largely caused by transport, particularly for GAC, NF and SPF to which it contributes
from 55-80%. For ozonation, the operation of the plant is the main contributor but
transport still adds 35% of the impact.
Human toxicity potential
Ozonation has the highest estimated human toxicity potential, equal to 188 kg 1,4-
DB Equiv./1,000 m3, 60% of which is from emissions of manganese to freshwaters. The
next highest impact is from SPF with a total of 165 kg 1,4-DB-Equiv./1,000 m3, with
52% attributed to the production of sodium hydroxide, again mostly due to manganese
136
emissions. NF is the best treatment alternative for this impact, with a value of 74 kg 1,4-
DB-Equiv./1,000 m3, followed by GAC at 92 kg 1,4-DB-Equiv.
Land transformation potential – natural, urban and agricultural
Ozonation has the greatest transformation potential for all three types of land,
requiring 0.1 m2 of natural, 2.5 m2 of urban ad 10 m2 of agricultural land per 1,000 m3 of
wastewater treated. By contrast, GAC uses 0.04, 1.9 and 5.5 m2, respectively. GAC is the
best option for natural and NF for urban land; for agricultural land they share joint first
place (0.055 m2), followed closely by SPF (0.057 m2). Much of the land transformation
for all options is associated with the operational requirements. For example, for ozonation
this is due to forest transformation, industrial areas and landfill sites associated with
electricity generation.
Particular matter formation potential
For this category, NF and GAC are the best alternatives, with a similar impact of
~0.4 kg PM10 Equiv./1,000 m3. SPF and ozonation have around a 1/3rd higher impact
(~0.6 kg PM10 Equiv./1,000 m3). For SPF, the impact is mainly from the productions of
sulphuric acid (57%) and sodium hydroxide (19%) due to the emission of SO2 which
contributes to the formation of particulates. For ozonation, emissions to air of NOx and
SO2 from electricity generation account for 62% of the total potential.
Photochemical oxidants formation potential
At 0.6 kg NMVOC/1,000 m3, GAC and NF are the best options for this category.
The worst option – ozonation – has almost twice as high impact (1.1 kg NMVOC/1,000
m3. The impact from SPF is estimated at 0.8 kg NMVOC/1,000 m3. The main contributor
across all four alternatives is the operation of the plant, related to NOx emissions from
electricity generation. Transport also contributes around 10% to the impact from GAC
and SPF.
137
Figure 39 – Life cycle impact of the advanced wastewater treatment techniques for PPCP compounds (error
bar represents minimum and maximum values for the parameters as specified in Table 16). All impacts are
expressed per 1,000 m3 of wastewater
0
200
400
600
800
1,000
1,200
1,400
1,600
1,800
Climate change
[kg CO2-Equiv.]
Fossil depletion
[kg oil Equiv.]
Metal depletion
[kg Fe Equiv. x0.1]
Water depletion
[m3]
Ozone depletion
[mg CFC-11Equiv. x 0.1]
Freshwater
eutrophication[g P Equiv.]
Marine
eutrophication[g N Equiv.]
Terrestrial
acidification[kg SO2 Equiv. x
0.01]
Ionizing radiation
[kg U235 Equiv.]
Granular activated carbon
Nanofiltration
Solar photo-Fenton
Ozonation
0
20
40
60
80
100
120
140
160
180
200
220
240
260
280
Freshwater
ecotoxicity[kg 1,4-DB Equiv.
x 0.1]
Marine ecotoxicity
[kg 1,4-DB Equiv.x 0.1]
Terrestrial
ecotoxicity[kg 1,4-DB Equiv.
x 0.01]
Human toxicity
[kg 1,4-DBEquiv.]
Natural land
transformation[m2 x 0.001]
Urban land
occupation[m2a x 0.1]
Agricultural land
occupation[m2a x 0.1]
Particulate matter
formation[kg PM10 Equiv.
x 0.01]
Photochemical
oxidants formation[kg NMVOC x
0.01]
Granular activated carbon
Nanofiltration
Solar photo-Fenton
Ozonation
138
Figure 40 – Contribution of different life cycle stages to the impacts of advanced treatment options
-10%
0%
10%
20%
30%
40%
50%
60%
70%
80%
90%
100%
110%
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
Climate
change
Fossil
depletion
Metal
depletion
Water
depletion
Ozone
depletion
Freshwater
eutrophication
Marine
eutrophication
Terrestrial
acidification
Ionizing
irradiation
Building Operation Decommissioning Transport
-10%
0%
10%
20%
30%
40%
50%
60%
70%
80%
90%
100%
110%
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
GA
C
NF
SP
F
OZ
O
Freshwater
ecotoxicity
Marine
ecotoxicity
Terrestrial
ecotoxicity
Human
toxicity
Natural land
occupation
Urban land
occupation
Agricultural
land
occupation
Particular
matter
formation
Photochemical
oxidant
formation
Building Operation Decommissioning Transport
139
5.4. Parametric analysis
As mentioned earlier, the results discussed in the previous sections refer to the
mean values of the key operating parameters shown in Table 16 for the respective
treatment options. To examine the influence of these parameters, a parametric analysis
was carried out assuming in turn the minimum and maximum values of the parameters in
Table 16. The resulting variations in the impacts are shown as error bars in Figure 38. As
can be seen, most impacts from SPF and ozonation are susceptible to the variations in the
key operating parameters and vary widely. Due to this, for some of the categories they
become comparable to the other two alternatives. These include climate change, ozone
depletion, eutrophication, acidification and photochemical oxidants, where the minimum
values for ozonation are lower than the respective mean values for GAC. On the other
hand, NF showed little sensitivity to the variation in the operating parameters.
5.5. Freshwater ecotoxicity potential of PPCP compounds
This section considers freshwater ecotoxicity potential of the target PPCP
compounds when released with the effluent directly to freshwaters or to agricultural soils,
the latter if the effluent is used for irrigation. Both treated and untreated effluents are
considered to find out if and by how much the advance treatment could contribute to
reducing the overall ecotoxicity potential of the target PPCP compounds compared to the
effluent from the conventional WWTP (here termed as “untreated”, i.e. not subjected to
the advanced treatment). The USEtox methodology was used for these purposes
(Henderson et al. 2011; Rosenbaum et al. 2008). Note that the freshwater potential of the
advanced treatment options was estimated using the ReCiPe methodology so that the
estimates presented in this section are not comparable. Note also that it was not possible
to use the ReCiPe methodology to estimate the ecotoxicity potential of the target PPCP
compounds due to a lack of the characterisation factors.
140
The USEtox characterization factors for freshwater ecotoxicity potential of the
target PPCP compounds are given in Table 17, distinguishing between their potential
impact when released to freshwaters and agricultural soils. It can be noted that the latter
is much lower than the former for each PPCP. These values were used together with the
concentrations of the target compounds in the effluent before and after the advanced
treatment (see Table 10 and Table 15, respectively) to estimate the overall ecotoxicity
potential per 1,000 m3 of effluent. The results are given in Figure 41, showing the range
of values for the minimum, mean and maximum operating parameters (see SI Table 42
and Table 43 for the totals and results obtained for each compounds).
Table 17 – USEtox characterization factors for freshwater ecotoxicity of target PPCP compounds
Compound
Characterisation factor (CTUe/kg)a
Emission to
freshwaterb
Emission to
agricultural soilb (irrigation)
Diclofenac 2,670 105
Ibuprofen 209 4
Trimethoprim 474 19
Erythromycin 24,900 3,120
Sulfamethoxazole 2,990 195
Carbamazepine 854 13
Oestrone 21,400 19
17β-oestradiol 184,000,000 255,000
Triclosan 106,000 200
a CTUe: comparative toxic units. It represents an estimate of the potentially affected fraction
of species (PAF) over time and volume per mass of a compound emitted to the environment. CTUe/kg = (PAF.m³.day)/kg (Henderson et al. 2011; Rosenbaum et al. 2008).
b Values from Alfonsín et al. (2014).
a) Release to freshwaters b) Release to agricultural soil
Figure 41 – Freshwater ecotoxicity potential of effluents released from advanced wastewater treatments
compared to the impact from effluent with no advanced treatment (estimated using USEtox methodology).
The impact for “No treatment” in figure b) has been multiplied by a factor of 10 to show on the scale
0
500
1,000
1,500
2,000
2,500
3,000
3,500
4,000
4,500
5,000
Granular activated
carbon
Nanofiltration Solar photo-Fenton Ozonation No treatment
Effluent with PPCP compounds
Treatment
Fre
shw
ate
r eco
toxic
ity p
ote
nti
al
(CT
Ue/1
,00
0 m
3)
0
500
1,000
1,500
2,000
2,500
3,000
3,500
4,000
4,500
5,000
Granular activated
carbon
Nanofiltration Solar photo-Fenton Ozonation No treatment (x0.1)
Treatment
Effluent with PPCP compounds
Fre
shw
ate
r eco
toxic
ity p
ote
nti
al
(CT
Ue/1
,00
0 m
3)
141
As indicated in Figure 41a, releasing the PPCP compounds in the effluent to
freshwater without the advanced treatment has the mean ecotoxicity potential of 950
CTUe/1,000 m3. By comparison, the effluent from GAC has the equivalent potential of
19.1 CTUe/1,000 m3, reducing the impact from the untreated effluent by almost 99% (see
SI). Treating the effluent by ozonation reduces the ecotoxicity by 80% and with SPF by
75%. NF is the least efficient but still lowers the effluent ecotoxicity by more than a half
(56%). However, releasing the untreated effluent to agricultural soils achieves a much
higher reduction of freshwater ecotoxicity potential than treating it by any of the advanced
treatments and releasing to freshwaters. In that case, the mean freshwater ecotoxicity
potential is equivalent to 5.05 CTUe/1,000 m3 (see Figure 41b). This is 374 times lower
than the impact on freshwaters and almost four times lower than using the best treatment
method (GAC) and releasing the effluent to freshwaters. If the effluent is treated and then
released to agricultural soils, the benefit is even greater, reducing the impact by about
70% for NF to 99% for GAC.
This gap between treated and untreated effluent widens when the life cycle impact
from the treatment is taken into account. As can be seen in Figure 41a, treating the effluent
by GAC and releasing it to freshwaters reduces the mean ecotoxicity potential by almost
40% (1,144 CTUe/1,000 m3), relative to the untreated effluent (1,871 CTUe/1,000 m3).
The next best option – NF – provides on average only a 10% reduction. However, the
other two treatment options have a higher impact than if the effluent is left untreated: SPF
by 45% and ozonation by 18%. In the worst case, assuming the worst operating
conditions, the impact increases by up to 2.5 times for SPF and by 65% for ozonation.
Thus, these findings suggest that these two options should not be used on the grounds of
reducing the freshwater ecotoxicity impact, which is the primary motivation for advanced
treatment of PPCP compounds, but instead in more eminent uses (as wastewater reuse).
The difference between treating and not treating the effluent is even starker when
considering the release to soil, where the life cycle impacts of all the treatment options
outweigh the direct impacts of untreated effluent by several orders of magnitude (Figure
41b). For example, NF, the best option in this case, has around 170 times higher mean
ecotoxicity potential than the untreated effluent while for the worst option – SPF – this
difference ranges from 110-810 times in favour of the untreated effluent. Therefore, on
the basis of these findings, and also taking into account that the treatment generates many
additional impacts, it could be argued that PPCP compounds should be left untreated after
the conventional wastewater treatment and utilised on agricultural soils for irrigation.
This is discussed further in the next section with a focus on wastewater reuse.
142
5.6. Wastewater reuse
Agricultural irrigation is at present the most common option for reuse of
wastewater in Europe. It is particularly practiced in water-stressed regions. Its use is
favoured also because of lower effluent quality requirements compared to potable water,
which can be achieved by conventional secondary or simple tertiary treatments (Bixio et
al. 2006; Bdour et al. 2009). Furthermore, given that agriculture is one of the largest water
consumers, reusing wastewater provides a reliable and cheaper source of freshwater,
reduces water stress and the need for other water sources (Barceló & Petrović 2011).
As demonstrated in this work, the PPCP compounds in the effluent discharged
after conventional treatment and those subjected to the advanced treatment have low
freshwater ecotoxicity potential when applied to agricultural soils, although some
compounds, such as diclofenac and carbamazepine, shows evidence of soil accumulation
(Ternes et al. 2007; Yu et al. 2013; Xu et al. 2009; Liu & Wong 2013). Moreover, the
presence of heavy metals in the effluents (not assessed here) can pose risks to the
environment, especially in effluents are used for the irrigation of areas that receives
biosolids for crop cultivation (Karvelas et al. 2003; Tripathi & Tripathi 2011). Still, their
removal from wastewaters would potentially create a greater ecotoxicity potential as well
as a number of other impacts, such as climate change, acidification, eutrophication,
human toxicity, etc.
The results of this work also suggest that the removal of PPCP compounds to
achieve the water quality similar to potable water for release to freshwaters is not
environmentally sustainable since it creates a similar or greater freshwater ecotoxicity
impact than the untreated effluent. However, if the treatment is aimed at reuse of treated
water for drinking, then advanced wastewater treatment is environmentally more
sustainable than some drinking-water treatment methods, particularly desalination (Sala
& Serra 2004; Pasqualino et al. 2011; Muñoz & Fernández-Alba 2008; Raluy et al. 2004).
However, many obstacles need to be overcome to enable direct potable reuse of
wastewater from advanced treatment methods, such as pumping and buffering (blending
ratios with drinking water, reservoir maintenance) requirements, generation of harmful
by-products, regulations, social acceptance and economic viability (Salgot et al. 2006;
Urkiaga et al. 2006; Lim et al. 2008; NRC 2012).
143
While advanced treatment of PPCPs may not be warranted, site-control and
monitoring of compounds originating from WWTPs effluents and in freshwaters are
required to avoid contamination of water sources and consequently drinking water
supplies. This is particularly important as water in many cities are found to contain these
substances (Webb et al. 2003; Benotti et al. 2009; Huerta-Fontela et al. 2011; Kleywegt
et al. 2011). However, the risks posed to humans by these chemicals in potable water
supplies remain unknown (Jones et al. 2005; Gaffney et al. 2015).
5.7. Chapter conclusions
This Chapter considered life cycle environmental impacts of four advanced
treatment techniques for nine target PPCP compounds. The results suggest that, on
average, NF has the lowest impacts for 10 out of 18 categories. GAC is the best alternative
for six impacts, including climate change (together with SPF); but, it has the highest
marine eutrophication of all the options. SPF is the best technique for the latter and for
fossil depletion, in addition to climate change. However, it is the least sustainable for
seven other impacts. Nevertheless, ozonation can be considered the worst option overall,
with 10 impacts higher than for any other alternative.
However, most impacts from SPF and ozonation vary widely with the operating
parameters and, when considering their ranges rather than the mean values, for some
impacts they become comparable to the other two alternatives. These include climate
change, ozone depletion, eutrophication, acidification and photochemical oxidants, where
the minimum values for ozonation are lower than the respective mean values for GAC.
On the other hand, GAC and NF are favoured since they have greater removal
efficiencies for heavy metals and avoid production of harmful by-products during the
treatment, thus being more suitable for potable reuse of wastewater (see section 2.5.3.1).
Moreover, they are the only two alternatives with the life cycle freshwater ecotoxicity
lower than the effluents released from conventional WWTPs to freshwaters without
advanced treatment of PPCP compounds. However, releasing the untreated effluent to
agricultural soils achieves a much higher reduction of freshwater ecotoxicity potential
than treating it by any of the advanced treatments and releasing to freshwaters. Therefore,
the use of advanced wastewater treatment for agricultural purposes is not recommended.
144
Remarks concerning updates of the UK electricity grid supply
The consumption of electricity contributed more than significantly to the potential
environmental impacts of nanofiltration and ozonation due to their high consumption
(mean values of 412 and 750 kWh/1,000m3 respectively, Table 16). Since this work
utilized the Ecoinvent 2.0 database and it is based in data previously to the year 2006
(Dones et al. 2007), updates in the UK electricity grid during the last decade can have
important influence in the many impact categories. The shift in the electricity grid fuel
supply in the UK between 2000-2015 is shown in Figure 42, indicating that the
contribution of fossil fuels has decreased from 72% to around 60%, while other fuels
sources (e.g. renewables such as solar and wind) increased from 5 to over 15% in the
same period. This demonstrates that the UK has currently less traditional sources of
energy for electricity generation, targeting compliance to climate change and energy
security goals (Stamford & Azapagic 2012; Stamford & Azapagic 2014).
Figure 42 – Fuel sources used in the electricity grid supply between 2000 and 2015 in the UK
The above suggest that, due to the decrease of many potential environmental
impacts derived from the reduction of fossil fuel use (Stamford & Azapagic 2012),
nanofiltration may have currently a more accentuate advantage over granular activated
carbon as the best for advanced wastewater treatment, possibly reaching similar impacts
in climate change potential and freshwater ecotoxicity removal (see Figure 39); the
decrease in the potential impacts in ozonation is not expected to reach similar profile to
solar-Photo Fenton treatment because its impacts are expressively higher in most
categories (see Figure 39). Therefore, no substantial changes in impact categories ranking
is expected among these techniques due to electricity grid updates.
0
20
40
60
80
2000 2010 2015
Fossil Nuclear Others
Pe
rce
nta
ge o
f th
e e
lectr
icit
y g
rid
su
pp
ly (
%)
145
6. LIFE CYCLE ASSESSMENT OF SLUDGE TREATMENT
TECHNIQUES
This chapter presents the results of environmental life cycle assessment of sludge
treatment techniques. It first defines the goal and scope of the study in the next section,
followed by the inventory analysis (section 6.2) and impact assessment (section 6.3). The
effect on the impacts of different recovery rates of the products from sludge treatment is
explored through a sensitivity analysis (section 6.4). In the end of the chapter the
contribution of PPCP compounds and heavy metals on freshwater toxicity is evaluated
(section 6.5) and the conclusions are drawn (section 6.6).
6.1. Goal and scope
The goal of the study was to estimate and compare life cycle environmental
impacts of the five chosen sludge handling routes with diverse resource and energy
recovery potentials. A further goal was to estimate freshwater ecotoxicity of PPCP
compounds and heavy metals in the sludge and determine the extent and significance of
the impact from these contaminants. The functional unit was defined as “treatment of
1,000 kg of thickened sludge on a dry matter basis” (sludge dry solids mass). The scope
of the study was from ‘cradle to grave’. Construction and decommissioning of the
treatment plants were excluded due to a lack of data. This is not considered a significant
limitation of the study as previous studies indicated that their contribution to the impacts
is mostly insignificant (Johansson et al. 2008; Yoshida et al. 2013).
6.2. Inventory analysis
The data for the operation of the treatment plants were sourced from existing
facilities in Europe. The life cycle data were taken from Ecoinvent 2.2 (Frischknecht et
al. 2004). The sludge treatment techniques are illustrated in Figure 43 and described
below. The inventory data for each technique are summarised in Table 18, including the
resources recovered for which they were credited. A range of potential recovery rates of
resources were considered for each treatment method, from maximum to no recovery.
Biogenic CO2 emitted during the treatment process was not considered; however,
biogenic methane is included.
146
6.2.1. Overview of sludge treatment methods
6.2.1.1. Agricultural application of anaerobically digested sludge
As can be seen in Figure 43, in this method the thickened sludge is digested
anaerobically to generate biogas which is used to maintain the digester at 35ºC and the
excess is used for electricity generation, as commonly practiced in the UK
(DECC/DEFRA 2011). The process data for anaerobic digestion is from digesters of a
plant treatment sludge from 90,000 inhabitants (Hospido et al. 2004; Hospido et al. 2005).
The data for electricity generation from anaerobic digestion was based in the studied of
(Houdková et al. 2008). The digested sludge is then mixed with a dewatering agent
(polymer) and directed to filter beds to reduce the water content. The product, containing
24% of dry matter (DM) is then distributed to where it is used as a substitute to synthetic
fertilizers. The digested sludge is considered of high-quality (compatible with the US
EPA’s Class A standards (Lu et al. 2012; Jones-Lepp & Stevens 2007) and applied on
land following regulations (Iranpour et al. 2004) to minimise pathogens and freshwater
eutrophication.
The system was credited for the displacement of the equivalent amount (in mass)
of synthetic NPK 15-15-15 fertilizer (fraction of the total mass in ammonia nitrogen,
phosphorus pentoxide and potassium oxide) and electricity generation from biogas (in
kWh). The amount of the displaced synthetic fertilizer was estimated according to the
phosphorus and nitrogen content in the treated sludge (~16 kg/1,000 kg DM (Hospido et
al. 2005). To account for the variability in the nutrient content, two cases are considered:
maximum recovery, displacing 100 kg of synthetic fertilizer per 1,000 kg DM and the
mean value of 50 kg/1,000 kg DM (Table 18). The electricity generation was based the
maximum and mean generation data in the work of Houdková et al. (2008). A third option
assuming no recovery of nutrients or electricity was also considered to compare the effect
on the environmental impacts.
6.2.1.2. Agricultural application of composted sludge
In this technique, thickened sludge is first mixed with the bulking agent, e.g. wood
chips or saw dust. As these typically represent waste, they were not considered here. The
mixture is then composted under controlled conditions to achieve a desired composition
147
of compost. The data is from facilities receiving sludge of over 400,000 inhabitants, and
the compost is used as a fertilizer substitute and the system was credited for the equivalent
amount of synthetic fertilizer, based on phosphorus content data (Sablayrolles et al.
2010). Like the anaerobic digestate, the compost was assumed to be of high quality, with
minimum contribution to freshwater eutrophication and pathogens contamination. Due to
high uncertainty on emissions from composting (Sánchez et al. 2015), only methane
emissions was included, assumed to be half that of anaerobic digestion (Zigmontiene &
Zuokaite 2010; de Guardia et al. 2010) (see Table 18). The same recovery rates range
(100 kg of NPK to no recovery) from composting were assumed as for anaerobic
digestion but a median value of 25 kg of NPK was set due lower content of nutrients in
the compost.
6.2.1.3. Incineration
Before being incinerated, thickened sludge is mixed with a polymer to aid its
dewatering in centrifuges and to increase its calorific value (Figure 43). After reaching
the DM content of 35% in the centrifuges, the sludge is incinerated in a fluidised bed
combustor (FBC - two 4 tonne/h and one with 5.2 tonne/h capacity) at 850ºC, with the
addition of fuel and lime to improve the combustion efficiency and to control acid gases,
respectively (Hospido et al. 2005; Hall 2014; Gottschalk et al. 1996). The addition of
meat and bones was not included since these typically represent waste and newer
developments in FBC have made the process more efficient over the years (Hall 2014).
The heat from incineration is used to generate electricity and heat and the system was
credited for both. The range of heat-to-electricity ratios considered can be found in Table
1 and was based in similar fluidised bed combustor in Houdková et al. (2008). However,
since the UK have little infrastructure for heat distribution, the heat recovery potential
was set the current reach of district heating networks in the UK (<1% of the population)
(Which? 2015). The bottom ash is landfilled in a sanitary landfill and fly ash disposed of
as hazardous waste.
148
6.2.1.4. Pyrolysis
This technique involves first filter pressing to reduce the water content in the
thickened to 30% of DM. The sludge is then dried and pyrolysis produce tar and char,
depending on the pyrolysis temperature (syngas was not accounted due lack on
information concerning its composition). The data were sourced from several facilities
operating in the temperature range from 300–900ºC (Hospido et al. 2005). The production
of tar and char was credited for the equivalent amounts of heavy fuel oil and charcoal,
respectively (see Table 18).
6.2.1.5. Wet air oxidation
After the addition of a dewatering agent (polymer), the thickened sludge is
pumped to the reactor to be mixed with generated oxygen at high temperatures and
pressure (at 235ºC and 40 bar). The process data is from an estipulate treatment of the
sludge generated by 300,000 inhabitants. The output of this process is a carbon-rich
effluent which can be used as a substitute for methanol in the denitrification process in
wastewater treatment plants (Houillon & Jolliet 2005). The addition of catalyst (copper
sulphate) has been excluded due lack of data in the Ecoinvent database. The mass of
methanol avoided by using the effluent was estimated based on the energy content of
methanol (35 MJ/kg) and the recovered by process (7.5 GJ) (Houillon & Jolliet 2005).
6.2.1.6. Transport
All transport distances were assumed at 200 except for agricultural application
site (45 km from the facility), while the treatments themselves take place near the
wastewater treatment plant and hence there was no need for transport. For transport of
fertilizers from anaerobic digestion and composting, the same mass was assumed when
transporting to agricultural site.
149
Figure 43 - Overview of the sludge treatment methods considered in the study showing the recovery of
resources (fertilizer, heat, electricity, fuels and methanol)
Table 18 – Inventory data for the sludge treatment techniques (per 1,000 kg of dry matter)
Ecoinvent dataa
Agricultural
application of
anaerobically digested
sludgeb
Agricultural
application of
composted sludgec
Incinerationb Pyrolysisb Wet air
oxidationd Unit
Anaerobic digestion Compost mixing Centrifuge Filter press Polymer addition
Electricity,
medium voltage,
at grid (UK)
88.6 33.2 52.50 40.0 kWh
Diesel 8.91 kg
Polymer 3.72 5.00 0.10 kg
Carbon monoxide
emission 0.84 kg
Nitrogen dioxide
emission 0.85 kg
Nitrous oxide
emission 0.02 kg
Particles to air
(PM10) 0.08 kg
Filter bed Fermentation /
Maturation Incineration Thermal drying
High-pressure
oxidation
Electricity,
medium voltage,
at grid (UK)
49.1 501 9.50 118 796.8 kWh
Heat, natural gas,
at industrial
furnace > 100kW
1,638 kWh
Heavy fuel,
burned in furnace 31 3.40 kg
Polymer 5.50 kg
Sodium
hydroxide 12.2 kg
Lime, hydrated,
loose 4.96 kg
Ammonia, liquid 3.72 kg
Tap water, at user 15.20 m3
Carbon monoxide
emission 0.15 mg
Nitrogen dioxide
emission 1.00 mg
VOC emission to
air 44.30 g
Agricultural application of anaerobically digested sludge with nutrients recovery
Agricultural application of composted sludge with nutrients recovery
Incineration with heat and electricity recovery
Pyrolysis with heat and fuels recovery
Wet air oxidation with methanol recovery
Polymer
Thickened sludge Centrifuge IncinerationHeat and electricity
Ash to landfill
Polymer
Thickened sludge Filter press PyrolysisFuels
Inert waste to landfill
Thermal
drying
Polymer
Thickened sludgeHigh-pressure
pumping
Wet air
oxidationMethanol
Thickened sludgeAnaerobic
digestionFilter bed Storage Fertilizer
Polymer
Thickened sludge Mixing Composting
Inert waste to landfill
Fertilizer
Electricity
150
Particles to air
(PM10) 2.00 µg
Furan emission 3.0E-5 ng
Agricultural
application
Agricultural
application
Waste
management Pyrolysis
Waste
management
Electricity,
medium voltage,
at grid (UK)
58.5 244 kWh
Diesel 0.73 0.73 kg
Methane
emission 3.2 1.6 kg
Carbon monoxide
emission 480 g
Nitrogen dioxide
emission 217 g
Nitrous oxide
emission 3.66 g
Particles to air
(PM10) 43.5 g
Disposal,
hazardous waste,
0% water, to
underground
deposit
19 kg
Disposal, inert
waste, 0% water,
to sanitary
landfill
273 35.5 kg
Resource
recovery Maximum/mean/nil Maximum/mean/nil
Maximum/
mean/nil Maximum/mean/nil Maximum/mean/nil
NPK 15-15-15 100/50/0 100/50/0 kg
Electricity,
medium voltage,
at grid (UK)e
794/397/0 454/227/0 kWh
Heat, at local
distributione 24/12/0 kWh
Charcoal 230/115/0 kg
Heavy fuel oil 40/20/0 kg
Methanol 214/107/0 kg
Disposal, inert
waste, 0% water,
to sanitary
landfill
0/135/270 kg
Transport
Digested and
composted sludge 107 107 t.km
Wastes 58 0/54 7.1 t.km
Chemicals 13 14 11 1 0.7 t.km a The specific life cycle inventory datasets used from Ecoinvent to estimate the environmental impacts of each treatment method. b Hospido et al. (2005)
c Sablayrolles et al. (2010)
d Houillon & Jolliet (2005) e Houdková et al. (2008)
6.3. Life cycle impacts results and discussion
The results are summarised in Figure 44 and Figure 45 and discussed below. The
error bars in Figure 44 represent the results obtained when using minimum and maximum
values of the parameters in Table 18, while the chart bars relate to the mean values of the
parameters. The discussion below refers to the latter. As can be inferred, no treatment
method is environmentally superior across all the impact categories. However, the
agricultural application of anaerobic digested sludge could be considered
environmentally the most sustainable option, with 13 out of 18 impact categories lower
than for any other option while wet air oxidation is the worst alternative with 7 highest
impacts. Most the impacts are related to grid electricity.
151
Climate change potential
Application of sludge from anaerobic digestion to agricultural land has the lowest
climate change potential, equal to -355 kg CO2 Equiv./1,000 kg DM after the credits for
electricity and fertilizer. The main contributors are the emissions of methane from the
digester and CO2 from grid electricity generation (80 and 113 kg CO2 Equiv./1,000 kg
DM respectively). Thus, using sludge as fertilizer avoids 127 kg CO2 Equiv./1,000 kg
DM and electricity recovery 480 kg CO2 Equiv./1,000 kg DM. The second lowest impact
was found for incineration, estimated at -79 kg CO2 Equiv./1,000 kg DM. This is largely
due to the emissions from heavy fuel oil from the incinerator of 108 kg CO2 and electricity
consumption of 38 kg CO2 Equiv./1,000 kg DM. The credits for energy recovery reduces
the emissions by around 285 kg CO2 Equiv./1,000 kg DM, of which 4% is due to heating
and 96% due to electricity.
Composting and wet air oxidation have impacts of 359 and 328 kg CO2
Equiv./1,000 kg DM. For both, the impact is mostly (> 80%) due to grid electricity. The
credits for the resource recovery decrease the total in 15% and 25% respectively. At 439
kg CO2 Equiv./1,000 kg DM, pyrolysis is the worst option for this impact, with natural
gas used for drying accounting for 422 kg CO2 Equiv. and grid electricity for 242 kg CO2
Equiv./1,000 kg DM. The avoidance of charcoal and fuel oil reduce the climate change
potential of pyrolysis by 249 kg CO2 Equiv./1,000 kg DM (~25%).
Resource depletion potential – fossil fuels and metals
Pyrolysis is also the worst option for depletion of fossil resources (183 kg oil
Equiv./1,000 kg DM), mostly due to the use of natural gas for drying of sludge. This is
despite the 30% reduction in the impact from the resource recovery (8 and 47 kg oil
Equiv./1,000 kg DM for charcoal and fuel oil, respectively). The impact from incineration
is -14 kg oil Equiv./1,000 kg DM, indicating that this amount of fossil resources is saved.
Anaerobic digestion is the best option in this category (-113 kg oil Equiv./1,000 kg DM)
with -138 kg from electricity and -29 kg oil Equiv./1,000 kg DM from fertilizers. Wet air
oxidation is the next best alternative, with -29 kg oil Equiv./1,000 kg DM.
Pyrolysis and wet air oxidation have the lowest metal depletion (0.69 kg Fe
Equiv./1,000 kg DM). Incineration has metals depletion over four times this value, 2.90
kg Fe Equiv./1,000 kg DM. The impact is predominantly from the disposal of hazardous
waste (2.5 kg Fe Equiv./1,000 kg DM) but is compensated by the energy recovery
152
potential (2.2 kg Fe Equiv.). Composting causes the highest depletion of metals,
estimated at 4.5 kg Fe Equiv./1,000 kg DM, 65% of which is related to the life cycle of
grid electricity. Finally, anaerobic digestion is the preferred option in the category,
avoiding depletion of 1.97 kg Fe Equiv./1,000 kg DM (3.8 kg of which from electricity
recovery).
Water depletion potential
Wet air oxidation and composting have the highest water depletion potential (581
and 476 m3/1,000 kg DM, respectively), mostly due to grid electricity (> 80%). Anaerobic
digestion is again the best option, saving -550 m3/1,000 kg DM by displacing grid
electricity.
Ozone depletion potential
The lowest ozone layer depletion potential is found for wet air oxidation, followed
by anaerobic digestion, with -222 and -53 mg CFC-11 Equiv./1,000 kg DM, respectively.
Pyrolysis is the worst option (466 mg CFC-11 Equiv./1,000 kg DM), mainly due to the
use of natural gas. The impact from incineration and composting, estimated at 115 and
162 CFC-11 Equiv./1,000 kg DM, respectively, is largely caused by diesel, heavy fuel oil
and for transport.
Eutrophication potential – freshwater and marine
Wet air oxidation and composting have the highest freshwater eutrophication
potential (120 and 96 and g P Equiv./1,000 kg DM, respectively); this is attributed to the
emissions of PO43- during generation of grid electricity. Pyrolysis is the third worst option
with 61.5 kg P Equiv., also related to grid electricity. Anaerobic digestion is the best
alternative, saving -101 g P Equiv./1,000 kg DM, followed by incineration, saving -47 g
P Equiv./1,000 kg DM.
Anaerobic digestion and incineration are also the preferred alternatives for marine
eutrophication, saving -39.0 and 10.8 g N Equiv./1,000 kg DM, largely due to the credit
for electricity, preventing emissions of 32 g from fertilizer and 65 g N Equiv./1,000 kg
DM from grid electricity in anaerobic digestion, and of 37 g N Equiv./1,000 kg DM from
grid electricity in incineration. As for freshwater eutrophication, composting and wet air
153
oxidation have the highest marine eutrophication of 50 and 55 g N Equiv./1,000 kg DM).
This is mainly due to NOx emissions to air and NO3- to freshwater during generation of
grid electricity. Due to the recovery of electricity, anaerobic digestion and incineration
save -39 and -11 g N Equiv./1,000 kg DM.
Acidification potential – terrestrial
Anaerobic digestion is also the best option for this impact, saving -0.62 kg SO2
Equiv./1,000 kg DM. By contrast, the worst options – composting and wet air oxidation
– emits 1.30 kg SO2 Equiv. The impact from the other two techniques is also very similar,
of 0.71 and 0.77 kg SO2 Equiv./1,000 kg DM (incineration and pyrolysis, respectively).
The main contributors (>50%) across all the options are NOx and SO2 emissions from
grid electricity generation. For anaerobic digestion and pyrolysis ammonia emissions
from digestate and NOx emissions from combustion of natural gas during drying of the
sludge are also significant contributors to this impact (0.47 kg and 0.30 kg SO2
Equiv./1,000 kg DM respectively).
Ionizing radiation potential
The highest impact for this category is from wet air oxidation (140 kg U235
Equiv./1,000 kg DM) and the lowest for anaerobic digestion (-108 kg U235 Equiv.). The
former is due to nuclear power in the grid electricity mix and the latter due to its avoidance
through the recovery of electricity from sludge, saving 145 kg U235 Equiv./1,000 kg DM.
Ecotoxicity potential – freshwater, marine and terrestrial
Freshwater and marine ecotoxicity follow a similar trend, with anaerobic
digestion being the best option, saving -1.4 kg and -0.14 kg 1,4-dichlorobenzene (DB)
Equiv./1,000 kg DM, respectively. This is followed by pyrolysis with the respective
impacts of 0.34 and 0.06 kg 1,4-DB Equiv./1,000 kg DM. For both categories,
composting and wet air oxidation are the worst alternatives, with the impacts five to six
times higher than from pyrolysis. For all the treatment techniques, emissions of heavy
metals from grid electricity generation are the main contributors to these two impacts.
154
For terrestrial ecotoxicity, pyrolysis is the best treatment method, saving -0.9 kg
1,4-DB Equiv./1,000 kg DM, followed by wet air oxidation with 0.22 kg 1,4-DB Equiv.
For the latter, the credit for avoiding the use of methanol cancels out the impact from grid
electricity. Composting and aerobic digestion are the least preferred for this category,
with 0.35 and 0.27 kg 1,4-DB Equiv./1,000 kg DM, respectively. The impacts from these
two options, as well as from incineration, are dominated (>50%) by transport because of
emissions of chlorine to industrial soil.
Human toxicity potential
The two treatment methods consuming more electricity than the other options, i.e.
composting and wet air oxidation, have the highest human toxicity potential, estimated at
83.5 and 79.7 kg 1,4-DB Equiv./1,000 kg DM, respectively. This is due to emissions of
manganese and arsenic in the life cycle of electricity. Incineration and pyrolysis have a
similar impact of 37.1 and 39.5 kg 1,4-DB Equiv./1,000 kg DM, respectively. Anaerobic
digestion is the best option with -69.0 kg 1,4-DB Equiv.).
Land transformation potential – natural, urban and agricultural
Anaerobic digestion requires by far the least natural land, saving -0.06
m2.year/1,000 kg DM. By comparison, composting as the worst option occupies 0.09
m2.year /1,000 kg DM. This is largely attributed to grid electricity, which is also the
reason why anaerobic digestion is the best option as it avoids its use. The trend is quite
different for urban and agricultural land, with pyrolysis now by far the best alternative,
avoiding the use of -8.7 and -981 m2.year/1,000 kg DM, respectively. The former is
largely related to the avoidance of charcoal transport and the associated infrastructure and
the latter due to the avoidance of conventional charcoal production and related forest land.
The worst option for urban and agricultural land is wet air oxidation, due to the land area
required in the life cycle of electricity.
155
Figure 44 - Life cycle impacts of sludge treatment techniques expressed per 1,000 kg DM (The error bars
represent the minimum values for the recovery of resources specified in Table 1. DB: dichlorobenzene;
PM10: particulate matter, 10µm; NMVOC: non-methane volatile
-600
-500
-400
-300
-200
-100
0
100
200
300
400
500
600
700
Climate change[kg CO2-Equiv.]
Fossil depletion[kg oil Equiv.]
Metal depletion[kg Fe Equiv. x
0.01]
Water depletion[m3]
Ozone depletion[mg CFC-11
Equiv.]
Freshwatereutrophication[g P Equiv.]
Marineeutrophication[g N Equiv.]
Terrestrialacidification
[kg SO2 Equiv. x0.01]
Ionizing radiation[kg U235 Equiv.]
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
-140
-120
-100
-80
-60
-40
-20
0
20
40
60
80
100
120
140
Freshwaterecotoxicity[kg 1,4-DB
Equiv. x 0.1]
Marineecotoxicity[kg 1,4-DB
Equiv. x 0.1]
Terrestrialecotoxicity[kg 1,4-DB
Equiv. x 0.01]
Human toxicity[kg 1,4-DB
Equiv.]
Natural landtransformation[m2 yr x 0.001]
Urban landoccupation
[m2 yr x 0.1]
Agricultural landoccupation
[m2 yr x 0.1]
Particulate matterformation
[kg PM10 Equiv.x 0.01]
Photochemicaloxidants
formation[kg NMVOCEquiv. x 0.01]
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
-9,810 -277
156
Figure 45 - Contribution of different life cycle stages to the impacts of advanced treatment options (The
values refer to the maximum recovery of resources. ADG: anaerobic digestion; COM: composting: INC:
incineration; PYR: pyrolysis; WAO: wet air oxidation)
-100%
-75%
-50%
-25%
0%
25%
50%
75%
100%
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
Climate change Fossil depletion Metal depletion Water depletion Ozone depletion Freshwatereutrophication
Marineeutrophication
Terrestrialacidification
Ionizingirradiation
Treatment Waste management Transport System credits
-100%
-75%
-50%
-25%
0%
25%
50%
75%
100%
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
AD
G
CO
M
INC
PY
R
WA
O
Freshwaterecotoxicity
Marineecotoxicity
Terrestrialecotoxicity
Human toxicity Natural landoccupation
Urban landoccupation
Agriculturalland occupation
Particularmatter
formation
Photochemicaloxidants
formation
Treatment Waste management Transport System credits
157
Particulate matter formation potential
Anaerobic digestion has the lowest impact, reducing the particulate matter
formation by -0.23 kg PM10 Equiv./1,000 kg DM. The next best option is pyrolysis which
generates 0.16 kg PM10 Equiv./1,000 kg DM. The worst treatment methods for this
impact are composting and wet air oxidation with 0.44 kg and 0.41 kg PM 10
Equiv./1,000 kg DM respectively. In all cases, NOx and SO2 emissions from grid
electricity generation are the main contributors.
Photochemical oxidants formation potential
The highest impact in this category is from composting (1.24 kg NMVOC
Equiv./1,000 kg DM) due to emissions of NOx from electricity and diesel burning during
composting. Pyrolysis has the lowest impact, avoiding -2.77 kg NMVOC Equiv./1,000
kg DM because of the credits for avoiding carbon monoxide emissions. Incineration have
relative little impact (-0.6 kg NMVOC Equiv./1,000 kg DM) or displacing electricity.
Although also recovering electricity, due emissions from digestion (0.85 kg NMVOC
Equiv./1,000 kg DM) the agricultural application of anaerobic digested sludge has
potential for photochemical oxidants formation of around 0.23 kg and NMVOC Equiv.
6.4. Sensitivity analysis
This section considers how the impacts and the ranking of the treatment options
may change if the resource recovery potential is varied within the ranges specified in
Table 18 for each technique. The results, given in Figure 46, indicate that anaerobic
digestion and incineration are the most sensitive to the assumptions on resource recovery.
For instance, in anaerobic digestion, if energy and fertilizer were not recovered, climate
change would increase from -355 kg to 255 kg CO2 Equiv./1,000 kg DM and human
toxicity from -65 to 41 kg 1,4-DB Equiv./1,000 kg DM. Still, these values are still lower
than any other of the techniques. The other impacts would also be affected to a varying
degree (Figure 46).
158
A similar trend can be noticed for incineration with most impacts affected by
resource recovery potential, except for terrestrial ecotoxicity. For example, in climate
change low recovery of energy would increase impacts from -79 kg to 205 kg CO2
Equiv./1,000 kg DM. Other ecotoxicities (freshwater and marine), human toxicity and
most of other impacts would increase 2 to 3 times. The effect on pyrolysis of resource
recovery potential is also high, being most noticeable for metal depletion (five times
higher if no recovery of resources compared to the maximum recovery), ecotoxicities (5-
10 times) and human toxicity (2.5 times higher); climate change potential would increase
by around 50%.
Composting is the least affected, with most impacts unchanged with the variation
in the assumptions for resource recovery. The exceptions are climate change and marine
eutrophication which have increases of 15%. For wet air oxidation, the most significant
effect is found for fossil, metals and ozone depletions and natural land transformation (a
factor of six); most other impacts increase by 20-50% if methanol is not recovered
compared to the maximum recovery.
Although resource recovery rates have an effect in the impacts the relative ranking
of the alternatives have little change. For instance, agricultural application of anaerobic
digested sludge is consistently the prefer technique compared to any of the others at same
recovery potential. At any recovery potential, this option is the best in 13 out of 18
impacts. The switch in ranking happens only among incineration and pyrolysis. At
maximum recovery, pyrolysis is the best in 4 impacts while incineration does not score
as the prefer option in any impact. Towards mean and minimum products recovery, this
alternative in prefer in nil impacts and incineration in 4. The ranking for impacts
considering different resource recovery potential among the techniques should be
interpreted accordingly.
159
Figure 46 – The effect of different resource recovery rates on the environmental impacts of different sludge treatment
techniques (100%, 50% and 0% refer on the x-axis represent the maximum, mean and minimum values, respectively,
for the recovery of resources from different treatment options. ADG: agricultural application of anaerobically digested
sludge; COM: composting; INC: incineration;), PYR: pyrolysis; WAO: wet air oxidation)
-600
-400
-200
0
200
400
600
800
100% 50% 0%
ADG COM
INC PYR
WAO
Cli
mate
ch
an
ge
pote
nti
al
[kg
CO
2-E
qui
v.]
Recovery potential
-200
-100
0
100
200
300
100% 50% 0%
ADG COM
INC PYR
WAO
Fo
ssil
reso
urc
e d
ep
leti
on
po
ten
tia
l
[kg
oil
Eq
uiv.
]
Recovery potential
-400
-200
0
200
400
600
100% 50% 0%
ADG COM
INC PYR
WAO
Me
tal d
ep
leti
on
po
ten
tia
l
[kg
Fe
Eq
uiv.
x0
.01]
Recovery potential
-800
-600
-400
-200
0
200
400
600
800
100% 50% 0%
ADG COM
INC PYR
WAO
Wa
ter
de
ple
tio
n p
ote
nti
al
[m3]
Recovery potential
-250
-100
50
200
350
500
650
100% 50% 0%
ADG COM
INC PYR
WAO
Ozo
ne
de
ple
tio
n p
ote
nti
al
[mg
CF
C-1
1 E
qui
v.]
Recovery potential
-100
-50
0
50
100
150
100% 50% 0%
ADG COM
INC PYR
WAO
Fre
shw
ate
r e
utr
op
hic
ati
on
p
ote
nti
al
[g P
Eq
uiv.
]
Recovery potential
-60
-40
-20
0
20
40
60
80
100% 50% 0%
ADG COM
INC PYR
WAO
Mari
ne e
utr
op
hic
ati
on
p
ote
nti
al
[g N
Equi
v.]
Recovery potential
-100
-60
-20
20
60
100
140
100% 50% 0%
ADG COM
INC PYR
WAO
Terr
est
ria
l a
cidif
ica
tio
n po
tenti
al
[kg
SO
4E
qui
v.
x0.0
1]
Recovery potential
-140
-100
-60
-20
20
60
100
140
180
100% 50% 0%
ADG COM
INC PYR
WAO
Ion
izin
g
rad
iati
on
p
ote
nti
al
[kg
U2
35
Eq
uiv]
Recovery potential
-30
-20
-10
0
10
20
30
40
100% 50% 0%
ADG COM
INC PYR
WAO
Fre
shw
ate
r eco
toxic
ity p
ote
nti
al
[kg
1-4
DM
Equi
v x0
.1]
Recovery potential
-20
-10
0
10
20
30
100% 50% 0%
ADG COM
INC PYR
WAO
Ma
rin
e e
coto
xic
ity
po
ten
tia
l
[kg
1-4
DM
Eq
uiv
x 0
.1]
Recovery potential
-30
-20
-10
0
10
20
30
40
100% 50% 0%
ADG COM
INC PYR
WAO
Terr
est
rial
eco
toxic
ity p
ote
nti
al
[kg
1-4
DM
Equi
v x
0.0
1]
Recovery potential
160
Figure 46 – (cont.) The effect of different resource recovery rates on the environmental impacts of different
sludge treatment techniques (100%, 50% and 0% refer on the x-axis represent the maximum, mean and
minimum values, respectively, for the recovery of resources from different treatment options. ADG:
agricultural application of anaerobically digested sludge; COM: composting; INC: incineration; PYR:
pyrolysis; WAO: wet air oxidation)
6.5. Freshwater ecotoxicity of PPCP compound and heavy metals
In this section, the freshwater ecotoxicity of PPCP compounds and heavy metals
are assessed using the USEtox methodology. For PPCP substances, the results obtained
in section 4.3.5 of this work are considered, and information about concentration and
availability of heavy metals in sewage sludge is given next.
The concentration of heavy metals in the sludge is only a partial information about
the risks that they might pose once in the environment (Shrivastava & Banerjee 2004).
The study of heavy metals in sludge must analyse their speciation and bonds formed with
other substances aiming to determine their mobility and bioavailability potential (Silveira
et al. 2003). For this intent, sequential chemical extraction procedures such as the adopted
by European Community Bureau of Reference (BCR) method is the most broadly
methodology to determine heavy metals speciation in the sludge. It divides the heavy
metals contained in samples in: (i) exchangeable, (ii) associated with carbonates; (iii)
associated with hydrated iron and manganese oxides; (iv) associated with organic matter
and sulphides; and (v) residual.
-80
-60
-40
-20
0
20
40
60
80
100
120
100% 50% 0%
ADG COM
INC PYR
WAO
Hum
an to
xic
ity
po
tenti
al
[kg
1-4
DM
Eq
uiv
]
Recovery potential
-125
-75
-25
25
75
125
100% 50% 0%
ADG COM
INC PYR
WAO
Na
tura
l la
nd
tra
nsf
orm
ati
on
p
ote
nti
al
[m2
yr x
0.0
00
1]
Recovery potential
-100
-75
-50
-25
0
25
100% 50% 0%
ADG COM
INC PYR
WAO
Urb
an l
and o
ccupati
on p
ote
nti
al
[m2
yrx
0.1
]
Recovery potential
-100
-80
-60
-40
-20
0
20
40
60
80
100
100% 50% 0%
ADGCOMINCPYRWAOA
gri
cult
ura
l la
nd o
ccupa
tio
n p
ote
nti
al
[m2
yrx
0.1
]
Recovery potential
-40
-30
-20
-10
0
10
20
30
40
50
60
100% 50% 0%
ADG COM
INC PYR
WAOP
art
icula
r m
att
er
form
ati
on
po
tenti
al
[kg
PM
10
Eq
uiv
x 0
.01]
Recovery potential
-100
-75
-50
-25
0
25
50
75
100
125
150
100% 50% 0%
ADG
COM
INC
PYR
WAO
Pho
toch
em
ica
l;
ox
ida
nt
form
ati
on
po
tenti
al
[kg
NM
VO
C
x0.0
1]
Recovery potential
161
From the fractions commented above, the shares found in fractions i and ii are
suppose the most easily mobile and bioavailable (Dabrowska & Rosińska 2012; koro et
al. 2012; Gleyzes et al. 2002) and were the ones considered for this assessment. The data
was originated from thermophilic anaerobic digested sludge (55ºC for 30 days incubation
period) in Dabrowska & Rosińska (2012) and for composting estimated from aerobic
composted sludge (150h period) found in Liu et al. (2007). Concerning species released
during incineration, from the target metals only zinc was found to have significant release
of mobile species from volatile fractions at 900ºC in Liu et al. (2010). The USEtox
characterization factors for freshwater ecotoxicity of the target PPCP compounds and
heavy metals are shown in Table 20; their impact, estimated using the data in Table 11
and Table 19, can be found in Table 21.
Table 19 - Heavy metals in sludge applied on agricultural land and emitted by incineration
Sludge concentrationa
(mg/kg DM)
Emission by
incinerationd
(mg/kg DM)
Exchangeable and associated with carbonates
(%)
Minimum Mean Maximum Mean Anaerobic
digested b Composted c Incinerated e
Cadmium 0.4 2.1 3.80 0.57 0 25.0 0
Chromium 16 146 275 2.37 0.20 0 0
Copper 39 340 641 n. a. 0.40 0.04 0 Nickel 9 50 90 n. a. 36.0 28.3 0
Zinc 142 1,071 2,000 0.98 14.1 11.5 29.1 a European Comission (2001) b Dabrowska & Rosińska (2012) c Liu et al. (2007) d Hong et al. (2009) e Liu et al. (2010)
Table 20 – USEtox characterization factors for freshwater ecotoxicity potential of PPCP compounds and
heavy metals
Emissions to agricultural soil
(CTUe/kg)a
Emissions to air
(CTUe/kg)a
PPCP compoundsb
Diclofenac 105 -
Ibuprofen 3.67 - Trimethoprim 19.2 -
Erythromycin 3,120 -
Sulfamethoxazole 195 - Carbamazepine 12.5 -
Oestrone 19.3 -
17β-oestradiol 255,000 -
Triclosan 200 -
Heavy metals
Cadmium (II) 4,900 3,900 Chromium (III) 650 520
Chromium (VI) 53,000 42,000
Copper (II) 23,000 29,000 Nickel (II) 7,700 6,100
Zinc (II) 21,000 17,000
a CTUe: comparative toxic units. It represents an estimate of the potentially affected fraction of species (PAF) over time and volume
per mass of a compound emitted to the environment. CTUe/kg = (PAF.m³.day)/kg (Henderson et al. 2011; Rosenbaum et al. 2008).
b Values from Alfonsín et al. (2014).
162
Table 21 – Freshwater ecotoxicity potential of PPCP compounds and heavy metals contained in sludge
using the USEtox methodology
Freshwater ecotoxicity potential (CTUe/1,000 kg DM)
Anaerobic digested / composted sludge
Minimum Mean Maximum
Diclofenac 0.00 0.03 0.05
Ibuprofen 0.00 0.00 0.00
Trimethoprim 0.00 0.00 0.00
Erythromycin 0.28 0.41 0.54
Sulfamethoxazole 0.00 0.00 0.00
Carbamazepine 0.00 0.00 0.00
Estrone 0.00 0.00 0.00
17β-Estradiol 0.38 2.07 3.72
Triclosan 0.53 0.74 0.94
Total PPCPs 1.19 3.24 5.26
Anaerobic digested sludge Composted sludge Incineration
Minimum Mean Maximum Minimum Mean Maximum
Cadmium (II) 0.00 0.00 0.00 0.49 2.57 4.66 0
Chromium (III) 0.01 0.09 0.18 0.00 0.00 0.00 0
Chromium (VI) 0.85 7.74 14.6 0.41 3.60 6.79 0
Copper (II) 3.59 31.3 59.0 0.36 3.13 5.90 0
Nickel (II) 24.9 139 249 19.6 109 196 0
Zinc (II) 420 3,171 5,922 343 2,586 4,830 4.83
Total heavy metals 450 3,349 6,245 364 2,705 5,043 4.83
TOTAL 451.2 3,352 6,250 365.2 2,708 5,048 4.83
It is shown that when sludge is used in agriculture, the PPCP compounds at their
maximum concentration still have a freshwater ecotoxicity potential nearly 70 times
lower than the heavy metals at their minimum content in the sludge (5.26 and 364
CTUe/1,000 kg DM respectively). Assuming the mean concentrations of heavy metals,
their impact is 930 times higher than the mean impact of PPCP compounds; for their
maximum content in sludge, it is higher by a factor of nearly 1200. The main contributors
to the ecotoxicity of heavy metals is zinc (II), with 95% of the total for the mean
concentrations. This corroborate with previous findings indicating zinc as one of the most
problematic metals in soils due its mobility and decreasing of soil quality (Udom et al.
2004; Mantovi et al. 2005; Wong et al. 2001).
The results also suggest that the impact of heavy metals released by sludge
incineration (4.83 CTUe/1,000 kg DM) is relatively small, several orders of magnitude
lower than the impact of heavy metals applied to the agricultural land (364-6,245
CTUe/1,000 kg DM). However, relative to the PPCP compounds content in the sludge,
the impact from heavy metals emitted during incineration is similar to the sludge applied
on agricultural land (~1.19-5.26 CTUe/1,000 kg DM). The estimated freshwater
ecotoxicity considering the impact of the PPCP compounds and heavy metals in
combination to the freshwater ecotoxicity of the treatments life cycle is shown in Figure
47. It demonstrates that the methods that involve agricultural application of sludge
(anaerobic digestion and composting) have the highest freshwater ecotoxicity potential,
with composting being the worst option on average for this impact category.
163
As discussed above, this is largely due to the heavy metals content in the sludge.
In it, incineration have the lowest impact, nearly 4 times lower than composting.
However, assuming the lower bound of heavy metals content in the sludge used in
agriculture and maximum freshwater ecotoxicity during the life cycle of the thermal
methods, their impacts are somehow comparable. It can also be noted that the ranking of
the thermal options is congruent with that for freshwater ecotoxicity estimated using the
ReCiPe method. Given the large contribution of heavy metals in this impact category,
anaerobic digestion and composting indeed changed their results in comparison to
thermal treatments (see Figure 44).
Figure 47 – Total freshwater ecotoxicity potential (including PPCPs and heavy metals) of the sludge
treatment techniques according to the USEtox methodology (ADG: anaerobic digestion; COM:
composting: INC: incineration; PYR: pyrolysis; WAO: wet air oxidation)
To contextualize these results, Figure 48 shows freshwater ecotoxicity of heavy
metals estimated based on the legislative limits for application for sludge to land in some
European countries and in the US (Table 22). These are compared to the impact from
heavy metals discussed above and in Table 21.The results in Figure 48 suggest that the
Dutch legislation is the most stringent and potentially the most effective in limiting
freshwater ecotoxicity, requiring concentrations 2.5 times lower than the typical average
for sludge in Europe. All other countries have more lax standards in this respect, with the
Spanish and the US legislation allowing slightly higher freshwater ecotoxicity than the
maximum found in European sludge. In the UK, the limit values are defined according
land loading rates and local restrictions (European Comission 2001).
-1,000
0
1,000
2,000
3,000
4,000
5,000
6,000
7,000
ADG COM INC PYR WAO
Fre
shw
ate
r eco
toxic
icty
pote
nti
al
(CT
Ue/1
,000
kg
DM
)
164
Table 22 – Legislative limits for some heavy metals in the sludge applied on agricultural land in some
European countries and the US
Maximum concentration
(mg/kg DM)
Cadmium Chromium Copper Nickel Zinc
Spaina Soil pH < 7 20 1,000 1,000 300 2,500
Soil pH > 7 40 1,750 1,750 400 4,000
Denmarka Dry matter basis 0.80 100 1,000 30 4,000
Netherlandsa 1.25 75 75 30 300
UK Article 5, paragraph 2(b) of Directive 86/278/EEC
USAb 39 1,200 1,500 420 2,800 a European Comission (2001). b Iranpour et al. (2004).
Figure 48 – Freshwater ecotoxicity potential estimated according to the USEtox methodology and based
on the legislative limits for heavy metals in sludge applied to agricultural land in some European countries
and in the US in relation to the range of impact from heavy metals estimated in this work for different
sludge treatment methods (horizontal red lines). The impact takes into account only direct emissions from
the application of the sludge (i.e. it is not on a life cycle basis)
The terrestrial ecotoxicity of PPCP compounds and heavy metals were not
assessed in this work due to the lack of characterization factors, and these are expected to
be more critical than freshwater ecotoxicity since these substances are applied directly to
agricultural soils. Nevertheless, studies in literature concerning the latter already
suggested that long term application of sewage sludge can damage soils and compromise
crops if not closely controlled (Udom et al. 2004; Singh & Agrawal 2008), and thus the
agricultural application should be carefully monitored at a local scale to avoid undesired
environmental impacts.
0
2,000
4,000
6,000
8,000
10,000
12,000
14,000
Spain
(pH < 7)
Spain
(pH > 7)
Denmark Netherlands USA
Anaerobic digested
Composted
Fre
shw
ate
r eco
toxic
icty
po
ten
tia
l
(CT
Ue/
1,0
00 k
g D
M)
European range
165
6.6. Chapter conclusions
This study considered life cycle environmental impacts of five sludge treatment
techniques. The results suggest that agricultural application of anaerobic digested sludge
has the lowest environmental impacts for 13 out of 18 categories. Wet air oxidation and
composting are the worst alternatives at the mean and maximum recovery of products,
with the highest impacts for seven and eight categories respectively. Pyrolysis is the best
option for four impacts at the maximum recovery potential; however, at the lower
recovery, this technique is less competitive in relation to the others.
The impacts are sensitive to the assumptions on the recovery of resources in the
case of incineration and pyrolysis, affecting the ranking of the options. For the maximum
resource recovery, pyrolysis is the best option for three and four impacts out 18 impact
categories. However, at no resource recovery, incineration is the best for four and
pyrolysis in nil impacts. At all resource recovery potentials assessed in this study, the
agricultural application of anaerobic digested sludge is the best techniques for 13 out of
18 impacts. The smallest effect of resource recovery rates was found for composting.
The sludge from anaerobic digestion has, on average, the highest freshwater
ecotoxicity because the speciation of heavy metals contains more (bio)available species,
however followed closely by the composted sludge. The contribution of the PPCP
compounds in the sludge in the overall freshwater ecotoxicity is small if compared to
heavy metals (less than 2%), especially when combining with the freshwater ecotoxicity
potential from the life cycle impacts of the treatments. Therefore, stricter control of heavy
metals (more specifically zinc) should be enforced, aiming standards similar to the Dutch,
in order to agricultural application of biosolids have similar freshwater ecotoxicity
potential than thermal sludge treatments.
Remarks concerning updates of the UK electricity grid
As commented in the last Chapter, shifts in the UK electricity grid over the last
decade (Figure 42 in Chapter 5 conclusions) may significantly influence the potential
impacts of the sludge treatment techniques, and eventually their relative ranking. Among
the alternatives assessed, the ones relying more electricity generation/consumption are
the ones expected to be susceptible to step changes in their potential environmental
impacts. These are: anaerobic digestions (maximum net generation of 686.4 kWh/1,000
kg of DM); composting and wet air oxidation (consumption of 543.2 and 796.8
kWh/1,000 kg of DM respectively) (see Table 18).
166
Regarding the first, as shown in Figure 44, this treatment had its environmental
impacts results with a clear advantage over the other methods and, although increases in
its potential environmental impacts are expected for the current electricity grid supply, a
change in its relative ranking is not expected because of the meaningful advantage over
other methods. Concerning composting and wet air oxidation, the shift to cleaner
electricity grid supply might beneficiate the later since it had better results in 10 out of 18
categories and similar impacts in other 4 in relation to the former option (see Figure 44).
This is due to the greater reliance on electricity consumption of this treatment that, by its
turn, tends decrease the potential environmental impacts more sharply in cleaner
electricity grids. Nevertheless, change in the relative ranking is not expected since these
two alternatives showed results significantly higher than intermediate options,
incineration and pyrolysis.
167
7. LIFE CYCLE COSTING OF ADVANCED WASTEWATER AND
SLUDGE TREATMENT TECHNIQUES
This chapter presents the results of life cycle costing of the advanced wastewater
and sludge treatment techniques. It starts with the definition of the goal and scope,
defining the system boundaries and the functional unit of the study. The cost data used in
the study are detailed in section 7.2 and the results are discussed in section 7.3. The
sensitivity analysis for different prices of energy, chemicals and other materials is
presented in section 7.4, followed by a discussion on the economic feasibility of
wastewater reuse and resource recovery from sludge in section 7.5. The conclusion can
be found in section 7.6.
7.1. Goal and scope
The goal of the study was to estimate the life cycle costs of the selected advanced
wastewater treatment methods aimed at wastewater reuse and sludge handling techniques
aimed at resource recovery. The scope was from cradle to grave, comprising plant
construction and operation, equipment replacement, waste management and recovery of
resources. The sludge treatment systems were credited for the revenue from the sales of
the recovered resources but wastewater plants were not for recovery of potable water for
the reasons discussed in section 2.6.2.1. It was assumed that the advanced plants are
coupled with the conventional WWTPs serving 150,000 inhabitants and treating 64,000
m3/day of wastewater, which generates 7,000 kg/day of sludge (on a dry basis). To enable
a more efficient operation of the advanced treatment plants.
The functional unit for the advanced wastewater treatment was defined as the
“treatment of 1,000 m3 of effluent from conventional wastewater treatment”. For sludge,
the functional unit was “treatment of 1,000 kg of thickened sludge on a dry matter basis”
(sludge dry solids mass). All plants were assumed to be located in the UK, with the chosen
plant size representing the average capacity of WWTPs in the UK (DEFRA 2012). This
size is also suitable as some of the advanced treatment techniques are not yet available
for treating larger amounts of effluent or sludge. The lifetime of the plants was assumed
at 60 years. The following section gives a brief overview of the treatment techniques
considered, followed by a description of the methodology and data used for the estimation
of LCC. For an overview of the treatments please see Chapter 5 and Chapter 6.
168
7.2. Costs estimation and data sources
The construction costs CC are the costs of building the plant and they can be found
in Table 23. The infrastructure replacement costs IRC represent the expenditure for
replacing the pipes, pumps, control equipment, etc., assumed to occur every 15 years over
the 60-year lifespan of the plant. These costs are given in Table 24. The replacement costs
for the advanced sludge treatments were not considered due a lack of data, and its lifespan
were assumed similar to water treatment works of up to 30-years (Cashman et al. 2014).
The fixed operating costs FC relate to the cost of materials and energy which are used
regardless of the level of treatment of water or sludge (e.g. electricity for pumping or a
dewatering agent). On the other hand, the variable operating costs VC refer to the
materials and energy whose usage varies depending on the required quality of the treated
water (e.g. the amount of activated carbon) such as removal of suspended solids,
dissolved organic carbon, turbidity and pollutants; thus, they were only considered for
the wastewater treatment methods.
The data for the operating parameters for the wastewater plants can be found in
Table 25 and for sludge processing in Table 26; the corresponding FC and VC are detailed
in Table 27-Table 29. To account for the uncertainty in the sourcing (origin) of different
materials and chemicals, both UK production and imports from China were considered,
the latter being a significant exporter of goods worldwide. As can be seen in Table 27,
there is a large difference between the costs in the respective countries so that the average
values were used for the estimation in the base case; the effect of these differences on the
LCC was considered in a sensitivity analysis.
The costs of waste disposal WC include landfilling and incineration of waste and
they are shown in Table 30, together with the transportation costs T, which include
transport of materials, chemicals and wastes; transport of recovered resources to the point
of sale is excluded. The infrastructure for heat and electricity distribution is also excluded
as that is already in existence regardless of sludge treatment. The sludge treatment plants
were credited for the revenue S from the sales of recovered resources, based on their
amounts (Table 25-Table 26) and the market prices of the products that they potentially
replace (Table 31 ). The costs were converted to British pounds (£) using SI Figure 75.
169
The water treatment plants were not credited for a potential revenue for selling
tap water as wastewater is currently not used for this purpose in the UK so it is not known
at what price the reused wastewater would be sold. In addition, new infrastructure would
be required to enable distribution of tap water from wastewater treatment plants and the
data for this were not available (section 2.6.2.1). There are also consumer perception
issues which would need to be understood and resolved before reclaiming wastewater as
tap water. By contrast, the recovery of resources from sludge is well established and most
of the co-products are used commercially, including in agriculture and for energy supply.
As some operating parameters vary significantly for some of the treatment
methods, a range of values were considered as specified in Table 25 and Table 26. The
amount of the recovered resources sold was also varied, ranging from complete to no sale
of products (see Table 26). Labour costs for the operation of the plants were not included
as it was assumed that they are similar across the methods considered, given that they
would be integrated within a conventional WWTP. However, it is acknowledged that
some of the methods may incur higher labour costs due to the need for a specialized
workforce or more intensive maintenance (Andreoli & Von 1997; Healy et al. 2008;
European Commission 2001b; Tyagi & Lo 2013; Wang et al. 2005).
Table 23 – Construction costs for the advanced wastewater and sludge treatment techniques
Treatment
Construction
costs
(£M)
Included Excluded Sources
Granular
activated carbon 0.57 Contactors, pipes, pumps sand electrical equipment Pre-coagulation tanks Wang et al. (2005)
Nanofiltration 4.50
Housinga, high-pressure pumps, pressure tubes, tubes
support, diaphragms, joints, flux control equipment
and electric installations
Treatment of the
concentrate b
Bonton et al.
(2012) Elazhar et
al. (2009)
Solar photo-
Fenton 1.10 Solar panel materials and flux control equipment Precipitate separation Ortiz (2006)
Ozonation 2.60
Housing, air preparation & ozone generation units,
ozone contactors, ozone diffusers and monitoring
equipment
- Wang et al. (2005)
Anaerobic
digestion 2.20 Digesters and filter bed
Land application
machinery and storage
facility
Hung et al. (2013)
Composting 2.10 Aerated composting facility
Land application
machinery and storage
facility
Hung et al. (2013)
Incineration 3.20 Centrifuge and incinerator Hung et al. (2013)
Pyrolysis 6.10 Filter press, thermal drying and pyrolysis apparatus - Hung et al. (2013)
Wet air oxidation 4.90 Wet air oxidation plant - Hung et al. (2013) a Assumed similar to ozonation. b Concentrate assumed redirected to conventional treatment line.
170
Table 24 – Infrastructure replacement costs for the advanced wastewater treatment techniques over the
lifespan of the plant (60 years)
Treatment Infrastructure
replacement costs
(£M)
Included Sources
Granular activated carbon 1.30 Pipes, pumps and electrical equipment Wang et al. (2005)
Nanofiltration 3.90 High-pressure pumps, pressure tubes, tubes support,
diaphragms and joints
Bonton et al. (2012)
Elazhar et al. (2009)
Solar photo-Fenton 3.30 Solar panel materials and flux control equipment Ortiz (2006)
Ozonation 7.50 Air preparation and ozone generation units, ozone
contactors and ozone diffusers
Wang et al. (2005)
Table 25 – Operating, waste management and transport data for the advanced wastewater treatment
techniques (per 1,000 m3 of secondary effluent)
Granular activated carbon Nanofiltration Solar
photo-Fenton Ozonation
Unit
(per 1,000 m3)
Fixed operating parameters
Electricity 19.56 0.42 kWh
Aluminium sulphate (powder) 80 kg
Calcium hydroxide 7 31 kg
Carbon dioxide, liquid 14 31 kg
Chlorine, liquid 0.60 0.60 kg
Phosphoric acid 1.10 kg
Polymer (dewatering aid) 0.30 kg
Sodium hydroxide 60 80 80 kg
Sulphuric acid 36 130 kg
Variable operating parametersa
Electricity 270 / 412 / 554 150 / 750 / 1,300 kWh
Ethylenediaminetetraacetic acid (EDTA) 0.16 / 0.25 / 0.34 kg
Fresh granular activated carbon 5 / 11 / 22 kg
Regenerated granular activated carbon 25 / 55 / 110 kg
Hydrogen peroxide 20 / 110 / 200 kg
Iron sulphate 14 / 34 / 55 kg
Sodium hydroxide 0.16 / 0.25 / 0.34 kg
Spiral-wound membranes 0.08 module
Waste managementa
Incineration (waste) 0.3584 kg
Landfill (sanitary) 5 / 11 / 22 46 kg
Transport abc
16-32 tonne lorry 44 / 57 / 81 20 58 / 80 / 102 16 t.km a Minimum/average/maximum values where shown.
b Over the lifetime of the plant c All transport distances assumed 200 km, except for fresh granular activated carbon ( 1,000 km). Transport of spiral-wound membranes to the treatment plant is not
included.
Table 26 – Operating, waste management and transport data for the sludge treatment plants (per 1,000 kg
of dry matter)
Agricultural
application of
anaerobic digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Unit
(per 1,000
kg DM)
Fixed operating
parameters
Electricity 197 534 62 402 797 kWh
Diesel 0.73 9.6 kg
Heavy fuel oil 31 3.4 kg
Natural gas 1,638 kWh
Ammonia, liquid 3.7 kg
Calcium hydroxide 5.0 kg
Polymer 5.5 4.0 5.0 0.10 kg
Sodium hydroxide,
50%
12 kg
Resource recoverya
Bio-char 0 / 115 / 230 kg
Bio-oil 0 / 20 / 40 kg
Electricity 0 / 397 / 794 0 / 227 / 454 kWh
Fertilizer (NPK) 0 / 50 / 100 0 / 25b / 100 kg
Heat (district heating) 0 / 12/ 24 kWh
Methanol 0 / 107 / 214 kg
Waste management
Landfill (inert waste) 0 / 135 / 270a 36 kg
Landfill (sanitary) 273 kg
Landfill (hazardous
waste)
19 kg
Transporcb
16-32 tonne lorry 121 122 70 1.0 / 28 / 55 8.0 t.km
a Nil/average/maximum values. b Half the amount of digested sludge due a half content of phosphorus. c Chemicals transport distances assumed at 200 km. Sludge transport to agricultural distances assumed 45 km. For pyrolysis, transport of
minimum/average/maximum amount of inert waste considered. Transport of recovered resources to the point of sale is excluded.
171
Table 27 – Prices of chemicals in the UK and imported from Chinaa
Price (£/kg)
Sources UK
Imports from
China Average
Aluminium sulphate (granules) 1.95 0.30 1.13 Easychemtrade (2016) / Shandong Sanfeng group
(2016)
Ammonia (liquid) 1.70 0.50 1.00 ReAgent (2016) / Shijiazhuang Xinlongwei Chemical (2016)
Calcium hydroxide (powder) 1.32 0.25 0.79 Mistral Industrial Chemicals (2016) / Guangdong
Qiangda New Materials Technology (2016)
Carbon dioxide, liquid 1.75 0.30 1.00 Gas UK (2016) / Anqiu Hengan Gas Manufacture
Factory (2016)
Chlorine 1.50 0.50 1.00 Alliance UK (2016) / Qingdao Huatuo Chemical (2016)
Ethylenediaminetetraacetic acid, EDTA (powder)
4.90 2.00 3.45 Mistral Industrial Chemicals (2016) / Jinan Yuxing Chemical (2016)
Hydrogen peroxide 0.90 0.60 0.75 Easychemtrade (2016) / Zhengzhou Qiangjin
Science and Technology Trading (2016)
Iron sulphate (powder) 1.20 0.40 0.80 Mistral Industrial Chemicals (2016) / Zhuzhou
Rongda Chemical (2016)
Phosphoric acid 1.68 0.85 1.27 Easychemtrade (2016) / Guangxi Qinzhou Capital Chemical (2016)
Polymer (granules) 1.60 0.85 1.23 British Plastics Federation (2016) / Hebei
Xiongye Machine Trade (2016)
Sodium hydroxide 0.36 0.25 0.31 Easychemtrade (2016) / Qingdao Huatuo
Chemical (2016)
Sulphuric acid 0.45 0.35 0.40 Easychemtrade (2016) / Wuhan Guotai Hongfa Commodity (2016)
a Costs of Chinese supplies were estimated from their average costs and shipping to the UK, at a rate of £200/t from
worldfreightrates.com. (accessed in February 2016).
Table 28 – Prices of granular activated carbon and nanofiltration membranesa
Price (£/unit)
Unit Remarks Sources Minimum Maximum Average
Granular activated carbon
(fresh) 1.20 1.60 1.40 kg -
Chengde Hongya Activated Carbon
(2016) / Jeswani et al. (2015)
Granular activated carbon
(regeneration) 0.60 0.80 0.70 kg -
Jeswani et al. (2015) / Bayer et al.
(2005);
Spiral-wound membrane 450 610 530 module NF90-8040 / NF270-400
Elazhar et al. (2009) / ServApure
(2016)
a Costs in Europe and the US. For the exchange rates, see Figure 75 in the SI.
Table 29 – Energy prices in the UKa
Price (£/unit)
Unit Minimum Maximum Average
Diesel 1.20 1.40 1.30 kg
Electricity 0.08 0.12 0.10 kWh
Heavy fuel oilb 0.48 0.64 0.56 kg
Natural gas 0.02 0.04 0.03 kWh
a Source: Department of Energy and Climate Change (2015).
b Used in incineration and recover product in pyrolysis.
172
Table 30 – Costs of waste disposal and transport
Cost (£/unit) Unit Sources
Incineration (waste) 90.0 tonne WRAP (2013)
Landfill (inert) 25.0 tonne WRAP (2013)
Landfill (sanitary) 35.0 tonne WRAP (2013)
Landfill (hazardous) 84.4 tonne WRAP (2013)
Transport 0.29 t.km Spielmann et al. (2007)
Table 31 – Market prices of products replaced by the equivalent resources recovered by sludge treatment
Average price
(£ /unit) Unit Remarks Sources
Charcoal 1.78 kg Made from wood
Ganzhou Green Top Biological Technology (2016)
/ Treewood (2016)
District heating 0.10 kWh Heating from diverse sources
WHICH? (2015)
Methanol 0.75 kg - Shijiazhuang City Horizon Chemical (2016) / ReAgent
(2016).
Synthetic
fertilizer 0.85 kg NPK 15-15-15
Zouping Runzi Chemical Industry (2016) / Agroshop
(2016)
7.3. Results and discussion
The life cycle costs of the advanced wastewater and sludge treatment methods are
summarized in Figure 49 and Figure 51, respectively, showing the contribution of
different life cycle stages and the range of costs, depending on the assumptions for the
operating variables and the sale of resource. The results are discussed in the next sections,
first for the advanced wastewater and then for the sludge treatment techniques. If not
stated otherwise, the discussion refers to the mean values of the parameters in Table 25
and Table 26, respectively.
7.3.1. Advanced wastewater treatment techniques
As can be seen in Figure 49, the lowest average LCC were found for ozonation
(£112/1,000 m3) and the highest for SPF (£215/1,000 m3), followed closely by GAC
(£205/1,000 m3). The costs of NF are estimated at £144/1,000 m3. However, taking into
account the variation in their operating parameters, the GAC costs in the best case
(£172/1,000 m3) approach the costs of ozonation at its worst operating conditions
(£162/1,000 m3). In the best case, the costs of SPF (123/1,000 m3) are also comparable
with the minimum NF costs (£129). However, ozonation is by far the cheapest option
assuming its best performance, costing only £47/1,000 m3.
173
The main contributor to the total LCC are the operating costs (~90%) for all the
treatment methods. On average, GAC and SPF are the most costly to operate
(~£188/1,000 m3) and ozonation is the least expensive (£100/1,000 m3). For GAC, 45%
of the total cost is due to aluminium sulphate used for coagulation (Figure 50) and 20%
due to the other chemicals used in the process. The energy-intensive regeneration of the
spent carbon contributes 18% to the total, with the cost of the fresh adsorbent adding a
further 7%. In the case of SPF, hydrogen peroxide represents 38% of the total costs and
iron sulphate (catalyst) 13%. For NF, chemicals account for 50%, electricity ~205% and
the membrane module 12% of the total LCC. The majority of the costs for ozonation are
due to electricity (65%) and sodium hydroxide (20%).
Transport costs are significant only for GAC and SPF (~10% of the total), the
former due to the transport of fresh GAC which is imported from Germany and the latter
due to the relatively large quantity of chemicals that need to be transported to the plant.
The construction, infrastructure replacements and waste management costs have a minor
contribution to the total LCC. NF is the most expensive plant to build (£3.2/1,000 m3)
while ozonation has the highest infrastructure replacement costs (£5.35/1,000 m3).
Figure 49 - Life cycle costs of the advanced wastewater treatment techniques showing the contribution of
different stages (The data labels represent the costs for the average and the error bars for the minimum and
maximum values of the parameters in Table 25)
0.4
1
0.9
3
53.9
0
132.8
5
0.3
9
16.5
3
205.0
0
3.2
1
2.7
8
61.6
7
70.4
9
0.0
3
5.8
0
143.9
8
0.7
8
2.3
5
109.7
0
76.8
4
1.6
1
23.2
0
214.4
9
1.8
6
5.3
5
75.0
0
24.8
0
4.6
4
111.6
5
0
50
100
150
200
250
300
Construction
costs
Infrastructure
replacement costs
Operating costs
(Variable)
Operating costs
(Fixed)
Waste
management costs
Transport costs Total
Granular activated carbon
Nanofiltration
Solar photo-Fenton
Ozonation
Lif
e c
ycle
co
sts
(£ /
1,0
00
m3
of
seco
ndar
yef
flu
ent)
174
Figure 50 - Contribution of different life cycle stages to the costs advanced of advanced wastewater
treatment techniques for the average operating parameters (For the latter, see Table 25. NF: nanofiltration;
EDTA: ethylenediaminetetraacetic acid)
7.3.2. Sludge treatment techniques
For the average recovery of products, pyrolysis is the best sludge treatment option
with an overall negative LLC, or a net profit of £29/1,000 kg DM (Figure 51). Anaerobic
digestion is the next least costly alternative with £8.7 followed by wet air oxidation at
£69. Composted sludge is the most expensive method with the cost estimated at
£107/1,000 kg DM. However, the costs vary widely, particularly for anaerobic digestion
and pyrolysis, depending on the assumptions for the sales of the recovered products. For
example, in the best case for pyrolysis, its profit increases by almost a nine-fold, from
£29 to £256/1,000 kg DM, but in the worst case it costs a total of £198. The best case for
incineration (complete recovery of products) is comparable to the average costs of wet
air oxidation, with costs of ~ £66/1,000 kg DM. Assuming the least favourable condition
for composting, it can operate at a profit of £128, lower than costs of wet air oxidation in
the same conditions (£149). For the maximum recovery of electricity and fertilizer value
from digested sludge, the plant breaks reach profits of £73/1,000 kg DM.
0%
10%
20%
30%
40%
50%
60%
70%
80%
90%
100%
Granular activated
carbon
Nanofiltration Solar photo-Fenton Ozonation
Construction & infrastr.
replacements
Fresh GAC
Regenerated GAC
NF membranes
Electricity
EDTA+NaHO
Iron sulphate
Hydrogen peroxide
Other chemicals
Aluminium sulphate
Waste disposal
Transport
175
In case of no sales of the products from sludge treatment, pyrolysis is the most
expensive option at £198/1,000 kg DM, followed by wet air oxidation at £149/1,000 kg
DM. Anaerobic digestion is the cheapest alternative, with a total LCC of £91/1,000 kg
DM. The contribution of construction and infrastructure replacements are expressive (20-
40% of the total operating cost) for all alternatives. As indicated in Figure 52, electricity
is an important cost factor for all the alternatives, contributing from 10% in digested
sludge system to 52% in wet air oxidation. The only exception to this is incineration
where it contributes around 5% to the total. The costs of natural gas for sludge drying are
significant for pyrolysis, contributing nearly 15%. In anaerobic digestion, pyrolysis and
wet air oxidation, around 45-65% reduction in their costs are obtained if mean recovery
of products is maintained during their life cycle (Figure 52). Pyrolysis is the most
expensive plant to build (£40/1,000 kg DM), followed by wet air oxidation (£32).
Transport is the most significant for the digested sludge and compost, contributing in ~
20%, respectively, mainly due to their transport to the farm. It also contributes 18% to
the costs of incineration because of the transport of ash to disposal.
Figure 51 - Life cycle cost of sludge treatment techniques showing the contribution of different stages (the
data labels represent the costs for the average and the error bars for the minimum and maximum values of
the parameters in Table 26.). Values for transport in pyrolysis includes waste management of non-recovery
resources
14
.35
14
.35
27
.41
-82.2
0
34
.80
8.7
2
13.7
0
13.7
0 65.8
8
-21.2
5
35.0
9
107.1
2
20.8
7
20.8
7
51.0
1
-23.9
0
20
.30
89.1
6
39.7
9
39.7
9
95.4
9
-215.3
3
11.3
5
-28
.90
31
.96
31
.96 8
2.6
3
-80
.25
2.3
2
68.6
2
-450
-350
-250
-150
-50
50
150
250
Construction
costs
Infrastructure
replacement costs
Operating costs
(Fixed)
Revenue
(Resource
recovery)
Transport costs Total
Agricultutral application of anaerobically digested sludge
Agricultutral application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Lif
e c
ycle
co
sts
(£ /
1,0
00
kg o
f dry
matt
er)
176
Figure 52 - Contribution of different life cycle stages to the costs of sludge treatment techniques for the
mean resource recovery (for the latter, see Table 26)
7.4. Sensitivity analysis
7.4.1. Energy costs
The costs of energy were varied for all the options between the minimum and
maximum values, and the results can be seen in Figure 53. For the advanced wastewater
treatment techniques, the greatest effect on the total LCC was found for ozonation, which
in the best case decreased by 16% from £112 to £97/1,000 m3 and, in the worst, increased
by 12% to 127/1,000 m3 (Figure 53a). This is due solely to the electricity cost as no other
forms of energy are used in this process. Still, this alternative remains overall the cheapest
wastewater treatment techniques. Likewise, the costs of NF, also reliant on electricity,
are affected by the variation in the electricity prices, ranging from £136 to £152/1,000
m3, compared to the average value of £144. However, the effect on the costs of the other
two alternatives is negligible (<0.5%). To explore the influence of the variation in the
energy costs on sludge treatment.
-60%
-40%
-20%
0%
20%
40%
60%
80%
100%
Agricultural
application of
anaerobic digested
sludge
Agricultural
application of
composted sludge
Incineration Pyrolysis Wet air oxidation
Construction &
infrastr. replacements
Electricity
Diesel
Natural gas
Heavy fuel oil
Polymer
Sodium hydroxide
Calcium hydroxide
Ammonia
Waste disposal
Transport
Products recovered
177
The results in Figure 53 show that the greatest effect is on pyrolysis due to its
dependence on both electricity and natural gas. Its LCC varied from -£6 to -£51/1,000 kg
of DM, showing that at high energy price this alternative show little profit potential at
average recovery of products. The total costs of composting vary approximately 10% and
wet air ozonation 20% according energy costs. However, the ranking of the options
remained the same as before. Therefore, these results suggest that the energy costs do not
affect the ranking of the options is preserved across the range of the cost values (at the
mean operating parameters).
Figure 53 – Influence of energy costs on the life cycle costs of advanced wastewater (a) and sludge (b)
treatment techniques (The vertical bars show the average LCC costs and the error bars the minimum and
maximum costs of energy given in Table 29)
7.4.2. Costs of chemicals and other materials
Like the high-energy users, the alternatives relying heavily on chemicals are most
influenced by their price variation. This is particularly noticeable for the activated carbon
system which uses a significant amount of aluminium sulphate, the price of which varies
widely. Figure 54 shows that the LCC of the GAC treatment range from £121 (imports
from China) to £289/1,000 m3 (UK production) at its average operating requirements,
representing a variation in the total costs of 30%-70%. This means that in the best case,
GAC is comparable to ozonation, which is on average the best option, and it becomes
cheaper than NF. However, at the highest costs of chemicals, it is by far the most
expensive option. NF and SPF are also sensitive to the costs of chemicals, varying from
23%-40% and 16%-24%, respectively, or by around £41/1,000 m3. This is mostly due to
the cost of carbon dioxide and calcium hydroxide used in the former and sodium
0
50
100
150
200
250
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Lif
e c
ycl
e co
sts
(£/1
,00
0 m
3o
f se
con
dar
y e
fflu
ent)
(a)
-60
-30
0
30
60
90
120
Agricultural
application of
anaerobically
digested sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Lif
e c
ycle
cost
s
(£/1
,00
0 k
gd
ry m
att
er)
(b)
178
hydroxide and sulphuric acid in the latter method. Ozonation is not affected by the costs
of chemicals.
In the sludge treatment techniques, incineration is the only option affected by the
costs of chemicals either, however in a small degree (~8%), mostly due to sodium
hydroxide used to balance the acid effluent from air pollution control (Gottschalk et al.
1996). Given that the costs of activated carbon contribute 25% to the total LCC of GAC,
the effect of the costs of fresh and regenerated carbon on the LCC of GAC is considered
here, using the cost data in Table 28. In addition, the costs of the membrane modules used
in nanofiltration are considered because of their significant variation (see Table 28). The
results in Figure 55 suggest that the total LCC of GAC are not affected significantly by
the variation in the costs of activated carbon, changing only by 4%. A similar outcome
was found for the total costs of NF, which varied by 2% with the costs of membranes.
Figure 54 – Influence of the costs of chemicals on the life cycle costs of advanced wastewater (a) and sludge
(b) treatment techniques (The vertical bars show the average LCC costs and the error bars the minimum
and maximum costs of chemicals given in Table 27)
Figure 55 – Influence of the costs of activated carbon and membranes on the life cycle costs of granular
activated carbon and nanofiltration (The vertical bars show the average LCC costs and the error bars the
minimum and maximum costs of these materials given in Table 28)
0
50
100
150
200
250
300
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Lif
e cy
cle
cost
s
(£ /
1,0
00
m3
of
trea
ted s
eco
nd
ary e
fflu
ent)
(a)
-60
-30
0
30
60
90
120
Agricultural
application of
anaerobically
digested sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Lif
e c
ycle
co
sts
(£ /
1,0
00
kg
dry
mat
ter)
(b)
0
50
100
150
200
250
Granular activated carbon Nanofiltration
Lif
e c
ycl
e co
sts
(£ /
1,0
00
m3
of
seco
nd
ary
eff
luen
t)
179
7.5. Economic feasibility of wastewater reuse and resource recovery from sludge
The advanced wastewater and sludge treatment techniques were assumed to be
coupled with a MBR in a conventional WWTP. MBRs are being increasingly adopted in
Europe due to their efficiency in treating wastewater and also for enabling wastewater
reclamation when combined with advanced treatment methods (Cases et al. 2011; Laera
et al. 2012; Alturki et al. 2010; Melin et al. 2006). Thus, arguably, the total costs of
wastewater reclamation should include both the MBR and advanced treatment costs.
Therefore, this section considers these total costs and compares them to the costs of
potable water produced in conventional potable water treatment plants to gauge if such
systems are economically feasible. In addition, the feasibility of different sludge
treatment techniques is also discussed.
The costs of R treatment at a medium to large scale (≥ 19,000 m3/d) are
estimated at approximately £300/1,000 m3 of urban effluent including capital and
operating costs, being electricity the major contributor (Hai et al. 2014; Côté et al. 2005).
If this cost is added to the costs of the advanced wastewater treatment estimated here, the
total costs range from £412 to £515/1,000 m3. As shown in Figure 56, these costs are
higher than the consumer costs of potable water in the UK, which range from £160 to
£240/1,000 m3 (South West Water 2015). However, in regions relying on desalination as
a source of freshwater, wastewater reuse through advanced treatment is an attractive
alternative, since the desalination costs are currently in the range of £450-850/1,000 m3
(Ghaffour et al. 2013). Therefore, the combination of MBR and advanced treatment
methods could be considered economically feasible and could in the future compete with
desalination facilities for potable water production. However, in addition to the costs,
other aspects must also be considered, including technical reliability of the advanced
treatment methods, potential generation of hazardous by-products and social acceptance
of wastewater reuse (Tchobanoglous et al. 2011; Moran & Dann 2008; Salgot et al. 2006;
Urkiaga et al. 2006).
180
Figure 56 – Comparison of costs estimated in this work for the production of potable water from wastewater
with water and sewage costs in the UK and costs of desalination worldwide (*Membrane bioreactor coupled
with one of the advanced wastewater treatment techniques operating at the average operating requirements;
distribution of the reclaimed wastewater to the end user not included)
The results of this work also demonstrated that the economic viability of some
sludge handling alternatives is highly dependent on the recovery potential and sales of
their products. Assuming the best-case scenario with all the outputs sold, anaerobic
digestion and pyrolysis could potentially be more profitable than the other methods.
However, for the later, the products are highly variable (both quality and quantity), which
hinders their use in most regions. Taking this into account, anaerobic digestion could be
considered more economically feasible than pyrolysis because the market for the digested
sludge and electricity from biogas is well established and its use is widely practiced.
7.6. Chapter conclusions
This study considered the life cycle costs of advanced wastewater treatment
methods aimed at recovery of potable water and sludge handling techniques for recovery
of resources. Among the wastewater treatment options considered, ozonation is the least
expensive, averaging £112 per 1,000 m3 of treated secondary effluent. Solar photo-Fenton
has the highest costs (£215/1,000 m3), followed closely by granular activated carbon
(£205/1,000 m3). However, the costs vary significantly with the operating parameters.
For example, in the best case the costs of granular activated carbon are comparable with
the top range of the ozonation costs.
100
300
500
700
900
Wastewater
reuse*
Potable water Desalinated
water
Co
st (
£/m
3 )
181
Similarly, for the most favourable conditions, Solar photo-Fenton is competitive
with nanofiltration. Nevertheless, ozonation is by far the cheapest option assuming its
best performance, costing only £47/1,000 m3. These costs are currently lower than
desalination costs and could be also competitive with conventional potable water in the
future. However, consumer acceptance of reusing wastewater as potable water may be a
significant barrier that should be explored further.
For the resource recovery from sludge, pyrolysis is the best option with an average
net profit of £29/1,000 kg dry mater. Anaerobic digestion is the next least costly
alternative with net costs of £8.7, followed by wet air oxidation at £69. With the cost
estimated at £107/1,000 kg, composting is the most expensive method for sludge
treatment. However, there is a significant variation in the costs, depending on the
assumptions for the sales of the products. For instance, the profits from pyrolysis would
increase by a factor of nine if all the outputs are sold, but if there is no recovery or
products, its overall costs approach £200/1,000 kg. Assuming the most favourable
conditions, incineration can operate at a profit of £65 but in the worst-case scenario, its
costs exceed the maximum costs of anaerobic digestion (£91/1,000 kg). Therefore, the
economic viability of the sludge treatment options is highly dependent on the recovery
rates and the revenue from the recovered resources and should be assessed carefully on a
case-by-case basis as most methods are site specific.
182
8. INTEGRATED SUSTAINABILITY ASSESSMENT OF
WASTEWATER AND SLUDGE TREATMENT METHODS
This chapter presents the results of an integrated sustainability assessment of the
wastewater and sludge treatment options using multi-criteria decision analysis (MCDA)
to help identify most sustainable options. All three dimensions of sustainability are
considered in MCDA– environmental, economic and social. First, as a reminder, a
summary of the life cycle environmental impacts and life cycle costs is given in sections
8.1 and 8.2, respectively, based on the results discussed in Chapters 5-7. This is followed
in section 8.3 by the discussion of the social sustainability of the options considered,
based on the methodology detailed in section 3.3.3. in Chapter 3. Finally, the integrated
sustainability assessment of the treatment techniques is presented in section 8.4 and the
conclusions in section 8.5.
8.1 Summary of life cycle environmental impacts
8.1.1 Wastewater treatment techniques
As discussed in Chapter 5, at the mean operating conditions, nanofiltration (NF)
has the lowest life cycle environmental impacts for 10 out of 18 categories (Figure 39).
Granular activated carbon (GAC) is the next best alternative, with its six impacts being
the lowest, including climate change (together with solar photo-Fenton); however, it has
the highest marine eutrophication. SPF is the best technique for the latter and for fossil
depletion, in addition to climate change. However, it is the least sustainable for seven
other impacts. Nevertheless, ozonation can be considered the worst option overall, with
10 impacts higher than for any other alternative. However, most impacts from SPF and
ozonation vary widely with the operating parameters and, when considering their ranges
rather than the mean values, for some impacts they become comparable to the other two
alternatives. These include climate change, ozone depletion, eutrophication, acidification
and photochemical oxidants, where the minimum values for ozonation are lower than the
respective mean values for GAC. These results can be consulted again in Figure 57.
183
Figure 57 –Potential environmental life cycle impacts of the advanced wastewater treatment techniques for
the mean operating conditions. Results per 1,000 m3 of secondary effluent
8.1.2 Sludge treatment methods
The LCA results discussed in Chapter 6 and summarized in Figure 58 suggest that
agricultural application of anaerobic digested sludge (ADG) has the lowest environmental
impacts for 13 out of 18 categories. Wet air oxidation (WAO) and composting are the
worst alternatives for the mean and maximum recovery of products, with the highest
impacts for seven and eight categories, respectively. Pyrolysis is the best option for four
impacts at the maximum recovery potential; however, at the lower recovery it becomes
less competitive in comparison to the other techniques. ADG has the lowest climate
change potential but high freshwater ecotoxicity because of heavy metals. Nevertheless,
composting is the worst option for this impact. The impacts are sensitive to the
assumptions on the recovery of resources in the case of incineration and pyrolysis,
affecting the ranking of the options. For the maximum resource recovery, they are the
best option for three impacts. However, at no resource recovery, incineration is the best
for four impacts and pyrolysis in nil impacts.
0
200
400
600
Cli
mate
chan
ge
[kg
CO
2-E
quiv
.]
Foss
il d
eple
tion
[kg
oil
Eq
uiv
.]
Meta
l deple
tion
[kg
Fe E
quiv
. x 0
.1]
Wat
er
dep
leti
on
[m3
x10]
Ozo
ne
deple
tion
[mg
CF
C-1
1 E
quiv
. x 0
.1]
Fre
shw
ate
r eu
troph
icati
on
[g P
Equiv
.]
Mari
ne
eutr
oph
icati
on
[g N
Equ
iv.]
Ter
rest
rial
acid
ific
atio
n[k
g S
O2 E
qu
iv. x
0.0
1]
Ion
izin
g r
adia
tion
[kg
U235
Eq
uiv
.]
Fre
shw
ate
r ec
oto
xic
ity
[kg
1,4
-DB
Eq
uiv
. x 0
.1]
Mari
ne
eco
toxic
ity
[kg
1,4
-DB
Eq
uiv
. x 0
.1]
Ter
rest
rial
ecoto
xic
ity
[kg
1,4
-DB
Eq
uiv
. x 0
.01
]
Hum
an t
ox
icit
y[k
g 1
,4-D
B E
quiv
.]
Nat
ura
l la
nd t
ransf
orm
atio
n[m
2 x
0.0
01]
Urb
an land
occ
upati
on
[m2
a x 0
.1]
Agri
cult
ura
l la
nd
occ
upati
on
[m2
a x 0
.1]
Part
icu
late
mat
ter
form
ati
on
[kg
PM
10 E
quiv
. x 0
.01
]
Pho
tochem
ical
oxid
ants
form
atio
n[k
g N
MV
OC
x 0
.01]
Granular activated carbon
Nanofiltration
Solar photo-Fenton
Ozonation
184
Figure 58 - Potential environmental life cycle impacts of the sludge treatment techniques at the mean
operating conditions. Results per 1,000 kg of dry matter
8.2 Summary of life cycle costs
8.2.1 Wastewater treatment options
Among the wastewater treatment options considered, for the average operating
conditions, ozonation is the least expensive, averaging £112 per 1,000 m3 of treated
secondary effluent (Figure 59). Solar photo-Fenton (SPF) has the highest costs
(£215/1,000 m3), followed closely by granular activated carbon (£205/1,000 m3).
However, the costs vary significantly with the operating parameters. For example, in the
best case the costs of GAC are comparable with the top range of the ozonation costs.
Similarly, for the most favourable conditions, SPF is competitive with nanofiltration.
Nevertheless, ozonation is by far the cheapest option assuming its best performance,
costing only £47/1,000 m3.
-400
-200
0
200
400
600
Cli
mate
chan
ge
[kg
CO
2-E
quiv
.]
Foss
il d
eple
tion
[kg
oil
Eq
uiv
.]
Meta
l deple
tion
[kg
Fe E
quiv
. x 0
.01
]
Wat
er
dep
leti
on
[m3
]
Ozo
ne
deple
tion
[mg
CF
C-1
1 E
quiv
.]
Fre
shw
ate
r eu
troph
icati
on
[g P
Equiv
.]
Mari
ne
eutr
oph
icati
on
[g N
Equ
iv.]
Ter
rest
rial
acid
ific
atio
n[k
g S
O2 E
qu
iv. x
0.0
1]
Ion
izin
g r
adia
tion
[kg
U235
Eq
uiv
.]
Fre
shw
ate
r ec
oto
xic
ity
[kg
1,4
-DB
Eq
uiv
. x 0
.1]
Mari
ne
eco
toxic
ity
[kg
1,4
-DB
Eq
uiv
. x 0
.1]
Ter
rest
rial
ecoto
xic
ity
[kg
1,4
-DB
Eq
uiv
. x 0
.10
]
Hum
an t
ox
icit
y[k
g 1
,4-D
B E
quiv
.]
Nat
ura
l la
nd t
ransf
orm
atio
n[m
2 y
r x
0.0
01]
Urb
an land
occ
upati
on
[m2
yr
x 0
.1]
Agri
cult
ura
l la
nd
occ
upati
on
[m2
yr
x 0
.1]
Part
icu
late
mat
ter
form
ati
on
[kg
PM
10 E
quiv
. x 0
.01
]
Pho
tochem
ical
oxid
ants
form
atio
n[k
g N
MV
OC
Eq
uiv
. x 0
.01
]
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
-9,8
10
-54
9
185
Figure 59 – Life cycle costs of the advanced wastewater treatment techniques showing the contribution of
different stages
8.2.2 Sludge treatment options
As can be seen in Figure 60, pyrolysis is the best option economically, with an
average net profit of £29/1,000 kg DM. Anaerobic digestion is the next least costly
alternative with the net cost of £9, followed by wet air oxidation at £69. At £107/1,000
kg, composting is the most expensive method for sludge treatment. However, there is a
significant variation in the costs, depending on the assumptions for the recovery and sales
of the products. For instance, the profits from pyrolysis would increase by a factor of nine
if all the outputs are sold, but if there is no recovery or products, its overall costs approach
£200/1,000 kg DM. Assuming the most favourable conditions, incineration can operate
at a profit of £65 but in the worst-case scenario, its costs exceed the maximum costs of
anaerobic digestion (£91/1,000 kg).
0.4
1
0.9
3
53.9
0
132.8
5
0.3
9
16.5
3
205.0
0
3.2
1
2.7
8
61.6
7
70.4
9
0.0
3
5.8
0
143.9
8
0.7
8
2.3
5
109.7
0
76.8
4
1.6
1
23.2
0
214.4
9
1.8
6
5.3
5
75.0
0
24.8
0
4.6
4
111.6
5
0
50
100
150
200
250
300
Construction
costs
Infrastructure
replacement costs
Operating costs
(Variable)
Operating costs
(Fixed)
Waste
management costs
Transport costs Total
Granular activated carbon
Nanofiltration
Solar photo-Fenton
Ozonation
Lif
e c
ycle
co
sts
(£ /
1,0
00
m3
of
seco
ndar
yef
flu
ent)
186
Figure 60 – Life cycle cost of sludge treatment techniques showing the contribution of different stages
8.3. Social life cycle impact assessment
The social sustainability assessment was carried out using the indicators defined
in section 3.3.3. Chapter 3. The results are summarized in (Table 32) for the wastewater
treatment methods and in Table 33 for the sludge handling options and are discussed
below at the national, supplier and consumer levels, respectively.
14
.35
14
.35
27
.41
-82.2
0
34
.80
8.7
2
13.7
0
13.7
0 65.8
8
-21.2
5
35.0
9
107.1
2
20.8
7
20.8
7
51.0
1
-23.9
0
20
.30
89.1
6
39.7
9
39.7
9
95.4
9
-215.3
3
11.3
5
-28
.90
31
.96
31
.96 8
2.6
3
-80
.25
2.3
2
68.6
2
-450
-350
-250
-150
-50
50
150
250
Construction
costs
Infrastructure
replacement costs
Operating costs
(Fixed)
Revenue
(Resource
recovery)
Transport costs Total
Agricultutral application of anaerobically digested sludge
Agricultutral application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Lif
e c
ycle
co
sts
(£ /
1,0
00
kg o
f dry
matt
er)
187
Table 32 – Social sustainability assessment of the advanced wastewater treatment techniques (per 1,000 m3 wastewater)
Level Social issue Indicator Granular
activated carbon Nanofiltration Solar photo-Fenton Ozonation
Unit per 1,000 m3
wastewater
National
Water securitya Water stress 0.4 0.4 0.4 0.4 -
Net water useb (min/mean/max) -690/-642/-553 -657/-531/-406 -188/150/480b -227/300/780 m3
Energy securitya Net energy use (min/mean/max) 21/22/23 270/412/554 0.42 150/750/1300 kWh
Food securitya Agricultural land use
(min/mean/max) 2.4/5.5/9.8c 1.6/5.5/7.1 1.8/5.7/7.5 6.7/10.7/16.8 m2.year
Suppliers Product adoption and the market Potential for product utilization Low Low Low Low -
Consumers
Human health
Damage to human health
(min/mean/max) 7.44/9.85/21.40 13.40/19.20/25.10 12.40/17.80/23.10 15.30/39.80/62.40 DALY x 10-4
Emerging contaminants and heavy metals
Very low Moderate Very high Moderate -
Product acceptance Wastewater reuse acceptance Low Low Low Low -
The rebound effect Moderate Moderate Moderate Moderate -
a The issues and indicators used to evaluate the impact on the energy-water-food nexus.
b Negative values denote the amount of freshwater saved by the treatment.
c Although the study concerns only UK, this impact includes the fresh granular activated carbon production in Germany.
188
Table 33 – Social sustainability assessment of the sludge treatment techniques (results per 1,000 kg of dry matter)
Level Social issue
Indicator
Agricultural
application of
anaerobically
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Unit per 1,000 kg of
dry matter
National
Water securitya Water stress 0.4 0.4 0.4 0.4 0.4 -
Net water use (min/mean/max)b -549/-175/200 476/490/504 -183/19/220 209/314/419 581/640/698 m3
Energy securitya
Net energy use (min/mean/max)b -598/-201/196 534/534/534 -316//-77/162 -626/707/2040 -1503/-353/797 kWh
Imported fossil fuel avoided
(min/mean/max) 0/35/70 0/0/0 0/21/42 0/132/264 0/90/180 koe
Diversity of outputs High Very low High Very high Very low -
Food securitya
Agricultural land useb (min/mean/max)
-7.0/-2.4/2.3 5.9/6.0/6.1 -1.6/0.9/3.4 -982.0/-488.6/4.8
8.7/8.8/8.9 m2.year
Synthetic fertilizer avoided
(min/mean/max) 0/7.5/15 0/3.8/15 0 0 0 kg of P
Suppliers Product adoption and
the market
Potential for product utilization High High Moderate Moderate High -
Public opposition to the treatment Moderate Low High Moderate Low -
Consumers
Human health Damage to human health
(min/mean/max) -27/-4.5//18 22/23/24 -10.7/-2.3/8.4 17.9/21.2/24.5 29/30.9/32.8 DALY x 10-4
Product acceptance
Similarity to traditional products High High Moderate Low Very high -
Presence of harmful substances Moderate Moderate Low Very low Very low -
The rebound effect Low Low Low Low Very low -
a The issues and indicators used to evaluate the impact on the energy-water-food nexus.
b Negative values denote the amount of freshwater, energy or land saved by the treatment.
189
8.3.1. Advanced wastewater treatment techniques
8.3.1.1 National level: Energy-water-food nexus
Following the methodology outlined in section 3.3.3.1.4. in Chapter 3, the EWF
integrates the indicators related to water, energy and food security to consider the effect
of different advanced wastewater treatment methods on the nexus. The results are
summarised in Figure 61 and discussed below in turn, followed by the integrated impact
on the nexus.
i) Water stress index
Given that this indicator refers to the water stress at the national level, there is no
difference between the waste treatment methods, with each assigned the same water stress
index (WSI) corresponding to the UK WSI. The latter is equivalent to 0.4 (Table 32)
which indicates a water stress slightly below the average (Smakhtin et al. 2004; Alcamo
et al. 2003). This suggests that the water availability in the UK as a whole is not critical
but is not abundant. This would suggest that at present wastewater reuse enabled by the
advanced treatment methods is not critical for a country such as the UK but would be
useful and it may become more important in the future with the advent of climate change.
However, some regions are already more water-stressed and would already benefit from
wastewater reuse, including Greater London, which has the WSI greater than 0.9.
ii) Net water, energy and land use
As can be seen from Figure 61, GAC is the best option for the net water use across
all the operating conditions considered, saving from 553-690 m3 of freshwater water per
1,000 m3 of wastewater treated. This is followed closely by NF with 406-657 m3.
Ozonation and SPF use more water than they produce at the mean and worst operating
conditions. For the net energy use, SPF is the best option across the operating parameters
and ozonation is the highest energy consumer at the mean and maximum operating
requirements; however, at the minimum, NF is the worst alternative. On the other hand,
the latter is the best option for the agricultural land use at all operating conditions
considered and ozonation is the worst.
190
iii) Impact on the EWF nexus
The integrating of values for the water, energy and food indicators in Table 32
according the steps in section 3.3.3.1.4 gives the results in Figure 61 and Table 34. Note
that the larger the area in Figure 61 the higher the negative impact on the nexus and vice
versa. As can be seen, SPF has the smallest nexus influence (0.0078, see Table 34) but
the second highest nexus homogeneity (0.3949) at the minimum operating parameters.
Nevertheless, it scored as the overall best alternative (nexus score 0.0129), followed
closely by GAC (nexus score 0.0198). At the mean operating parameters, GAC has the
preference in nexus influence and score as the lowest in nexus homogeneity (0.0024 and
0.1335 respectively), consequently scoring also as the best option in nexus score, with
0.0028. NF is the second-best alternative, with a nexus score of 0.0655. Although OZO
had the second-best nexus homogeneity score (0.1732), this alternative is the least
preferable in the nexus (nexus score of 0.9669) operating at the mean operating parameter.
Comparing the alternatives at their maximum operating requirements, GAF and NF
showed similar nexus influence score, of 0.0377 and 0.0355 respectively. Yet, these
results are still over 3 times higher than the obtained for SPF (0.0103) (see Figure 61).
The later scored the highest value in nexus homogeneity (0.3279), but still maintained its
overall preference on the nexus in comparison with the other techniques (nexus score of
0.0153), followed by GAC and NF, with nexus scores of 0.0454 and 0.0452 respectively.
Thus, it can be concluded from results in Table 34 that SPF is the preferred option
regarding EWF nexus impacts at minimum and maximum operating requirements, and
GAC at mean operating requirements.
Table 34 – Results for energy-water-food nexus impacts of the advanced wastewater treatment techniques
Operating
requirements Nexus impact category
Granular
activated
carbon
Nanofiltration Solar
photo-Fenton Ozonation
Minimum
Nexus influence
(Anexus) 0.0186 0.0776 0.0078 0.5262
Nexus homogeneity
(SDnexus) 0.0620 0.5232 0.3949 0.2325
Nexus score
(Nscore) 0.0198 0.1628 0.0129 0.6856
Mean
Nexus influence
(Anexus) 0.0024 0.0475 0.0088 0.7994
Nexus homogeneity
(SDnexus) 0.1355 0.2747 0.3471 0.1732
Nexus score
(Nscore) 0.0028 0.0655 0.0135 0.9669
Maximum
Nexus influence
(Anexus) 0.0377 0.0355 0.0103 0.7992
Nexus homogeneity
(SDnexus) 0.1704 0.2144 0.3279 0.1732
Nexus score
(Nscore) 0.0454 0.0452 0.0153 0.9666
191
Figure 61 – Impact of the advanced wastewater treatment techniques on the energy-water-food nexus
(integration of nexus indicators) for the minimum, mean and maximum operating requirements
0.0
1.0Water
EnergyFood
Minimum operating
requirements
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
0.0
1.0
Water
EnergyFood
Mean operating
requirements
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
0.0
1.0
Water
EnergyFood
Maximum operating
requirements
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
192
8.3.1.2 Suppliers: Potential for product utilization
This indicator, relevant to water suppliers, considers the readiness for the adoption
and distribution of treated wastewater as potable water. At present, there are no facilities
in the UK with the advanced treatment of wastewater that would enable its reuse as tap
water although plans are underway in the London area (IERP 2013). Furthermore, there
is no infrastructure for distribution of reused waste water to consumers. Therefore, the
current potential for the utilization of wastewater as potable water in the UK is considered
low (Table 32, see topic 2.6.2.1). As this indicator refers to the suppliers rather than the
individual technologies, there is no distinction between them with respect to the potential
for product utilization.
8.3.1.3 Consumers: Damage to human health
As indicated in Table 32 ozonation has the highest impact on human health, for
the mean operating parameters estimated at 39.8 DALY x 10-4/1,000 m3. This is
approximately twice as high as the values for NF and SPF and over four times higher than
that from GAC. The main contributors to this indicator are ionizing radiation from
electricity generation and for GAC hard coal burning.
8.3.1.4 Consumers: Emerging contaminants and heavy metals
This indicator refers to the presence of heavy metals and the emerging
contaminants such as PPCP compounds in potable water, affecting human health. GAC
is expected to be the most efficient advanced wastewater treatment method in removing
these contaminants. Furthermore, it does not generate transformation products during the
treatment, which favours it further over the other treatment methods (see section 2.53.1).
The least efficient method is SPF (Wang et al. 2005; Silva et al. 2012; Lofrano 2012; Lee
et al. 2009; Gogate & Pandit 2004a). NF and ozonation are considered to have a moderate
capacity for removing PPCP compounds and heavy metals (see section 2.53.1).
193
8.3.1.5 Consumers: Water reuse acceptance
Based on previous studies in Australia and the US, reuse of wastewater as potable
water faces moderate opposition due to consumer perceptions (Russell & Hampton 2006;
Marks 2006; Hartley 2006). It is not known how the UK public would react to the
proposals for wastewater reuse as no studies have been carried out yet. However, given
that wastewater is currently not reused as potable water and that the UK is not a water-
stressed country, the consumer opposition may be high. Therefore, the acceptance is
assumed to be low across the wastewater treatment techniques (Table 32).
8.3.1.6 Consumers: The rebound effect
The rebound effect takes into account that adoption of a ‘green’ product can lead
to a rise in consumption of the same or other products. A typical example are the low-
energy bulbs which people leave on for longer because they consume less energy. In this
case, the rebound effect refers to a potential increase in water consumption because the
consumer may consider that the water comes from wastewater and therefore does not
waste freshwater resources. However, there are no studies that would confirm or dispute
this supposition. Therefore, a conservative value has been assumed for this indicator,
suggesting a moderate rebound effect across all the technologies.
8.3.2. Sludge treatment techniques
The indicators considered for the EWF nexus for the sludge treatment techniques
are listed in Table 33 and are discussed in turn below.
8.3.2.1 National level: EWF nexus
i) Water stress index
This indicator is the same as for the wastewater treatment discussed in section
194
8.3.1.1 and is therefore not repeated here.
ii) Net water, energy and land use
ADG has the lowest water consumption, saving 175-549 m3/1,000 kg DM for the
mean and highest recovery of resources; however, at no recovery, it uses 200 m3/1,000
kg DM more than it treats. WAO is the worst option at all products recovery potentials
considered, consuming up to 700 m3/1,000 kg DM more water than it treats. However,
WAO consumes the least energy, saving between 350 and 1500 kWh/1,000 kg DM at the
mean and maximum recovery of resources. Pyrolysis is the second best option for energy
use at the highest recovery of its outputs but the worst at no recovery. In the best case,
pyrolysis also avoids the use of almost 1,000 m2.year of agricultural land while WAO, the
worst alternative for this indicator, requires around 9 m2.year. ADG is the second best
option for the minimum and mean recovery of outputs.
iv) Imported fossil fuels avoided
For this indicator, which is related to the national security of energy supply,
pyrolysis is the best option. Assuming a total recovery of its outputs, this option
potentially saves 264 koe/1,000 kg dry matter (DM), four times more than anaerobic
digestion and six times more than incineration.
v) Diversity of products
This indicator measures the contribution of the options to water, energy and food
production through the diversity of their products. Pyrolysis is also among the best
options for the diversity of products since it produces different types of fuels which can
be used in different applications. Incineration has a moderate score for this indicator –
while it produces electricity and heat, the latter cannot be distributed easily in the UK due
to a lack of infrastructure. ADG also has a moderate score as it produces fertilizers and
energy, but the latter also requires adequate infrastructure. The composted sludge and
WAO score very low since they only recovery one type of product.
195
vii) Synthetic fertilizer avoided
This indicator is related to the avoidance of social impacts from synthetic
fertilizers, such as human health and depletion of resources, in particular phosphorous,
which is becoming a scarce resource and may not be available for future generations, thus
affected intergenerational equity (Childers et al. 2011; Cordell et al. 2009; Cordell &
White 2008). Digested sludge and compost are the only options that avoid the use of
synthetic fertilizers, with the former displacing on average 7.5 kg of phosphorus and the
latter half that amount.
viii) Impact on the EWF nexus
Integrating the values for the water, energy and food indicators in Table 33
according the steps in section 3.3.3.1.4 and eqns. (20)-(27) gives the results in Figure 62
and Table 35. Note that the larger the area in Figure 62 and the higher value in Table 35,
the higher the negative impact on the nexus and vice versa. For maximum resource
recovery ADG had clear advantage over the other techniques for nexus influence. The
option had a score of 0.1075, over two times lower the second-best option in this impact
category, obtained for PYR (0.2491) and 5 times smaller than the worst ranked option,
WAO, with 0.5506. In relation to nexus homogeneity, PYR had the lowest score,
followed by COM (of 0.0295 and 0.0935 respectively). INC and WAO were the worst
alternatives in this due their high scores for their food indicator, as demonstrated in Figure
62. The nexus score indicated that ADG is the best option in the nexus impact, with great
advantage over the second in rank, PYR (0.1260 and 0.2567 respectively). For maximum
resource recovery potential, COM and INC had comparable scores in this category (of
0.4021 and 0.4306) while WAO is the lest recommended alternative.
Comparing the alternatives at their mean resource recovery potential, the results
in Table 35 suggested ADG again as the preferred option for sludge treatment. Although
scoring among the highest in nexus homogeneity (0.3195), the low nexus influence score
obtained for this category, of 0.1037, resulted in a nexus impact of 0.1524. This result
was over 2.5 times lower than for PYR (0.3757) and up to 5 times lower in comparison
to the worst alternative in the nexus at this resource recovery potential, WAO (0.7447).
196
If the alternatives are assessed when not recovering products, ADG and INC have
comparable nexus influence (0.0542 and 0.0685 respectively). The other alternatives had
nexus impact influence over 3 times these values. In nexus homogeneity, PYR is the
preferred alternative (0.1844), but closely followed by COM (0.2005) and ADG (0.2125).
Thus, the calculation of the nexus score indicated that ADG and INC with clear advantage
over the other alternatives, having scores of 0.0688 and 0.0927 (see Table 35). Thus, these
results suggest that ADG is the best option for sludge handling with respect to the EWF
nexus, followed by PYR when maximum resource recovery is possible and INC for no
products recovery.
Table 35 – Results for the energy-water-food nexus impacts of the sludge treatment techniques
Resource
recovery Nexus impact categories
Agricultural
application of
anaerobically
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Maximum
Nexus influence
(Anexus) 0.1075 0.3645 0.2692 0.2491 0.5506
Nexus homogeneity
(SDnexus) 0.1468 0.0935 0.3748 0.0295 0.2245
Nexus score
(Nscore) 0.1260 0.4021 0.4306 0.2567 0.7100
Mean
Nexus influence
(Anexus) 0.1037 0.5186 0.2298 0.3442 0.5597
Nexus homogeneity
(SDnexus) 0.3195 0.1362 0.4156 0.0838 0.2485
Nexus score
(Nscore) 0.1524 0.6004 0.3932 0.3757 0.7447
No recovery
Nexus influence
(Anexus) 0.0542 0.3035 0.0685 0.2225 0.4857
Nexus homogeneity
(SDnexus) 0.2125 0.2005 0.2607 0.1844 0.2773
Nexus score
(Nscore) 0.0688 0.3796 0.0927 0.2728 0.6721
197
Figure 62 - Impact of the sludge treatment techniques on the energy-water-food nexus (integration of
nexus indicators) for the maximum, mean and no recovery of resources
0.0
1.0Water
EnergyFood
Maximum products
recovery
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
0.0
1.0Water
EnergyFood
Mean products
recovery
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
0.0
1.0Water
EnergyFood
No products
recovery
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
198
8.3.2.2 Suppliers: Potential for product utilization
The potential for utilization of digested and composted sludge is high as they are
already being utilized (Milieu et al. 2010). For incineration and pyrolysis, a moderate
utilization potential of its products is expected – for the former due necessity of landfilling
of toxic wastes and infrastructure for heating distribution; for the latter as the technology
is still not fully developed and deployed; furthermore, use of fuel may require adaptation
of combustion engines (Fytili & Zabaniotou 2008; Kim & Parker 2008; Werle & Wilk
2010; Fonts et al. 2009). The outputs from WAO have a high utilization potential because
it can be used in the denitrification process in wastewater treatment plants (Wang et al.
2005; Lund et al. 2010; DEFRA 2013; Chauzy 2010).
8.3.2.3 Suppliers: Public opposition to the treatment
The public opposition to composting is deemed low due to small vector (e.g.
mosquito) attraction and pathogens content (L. Wang et al. 2008). These concerns are
somewhat higher for ADG (L. Wang et al. 2008; Giusti 2009) and thus a moderate score
was assigned to this option. On the other hand, public opposition to incineration is high
in the UK (L. Wang et al. 2008; European Commission 2001a; DEFRA 2015b; Wang et
al. 2005; Fytili & Zabaniotou 2008; European Commission 2006). For pyrolysis, a
moderate opposition is expected as this technology is sometimes confused by the public
with incineration. However, as it produces biochar and biofuel without generating dioxins
(the main objection to incineration), it may be more acceptable than incineration (Stehlík
2009). Finally, WAO is expected to be acceptable to the public as the product would be
utilized by the wastewater treatment plant without affecting consumers.
8.3.2.4 Consumers: Damage to human health
ADG and incineration are the preferred alternatives for this indicator as they both
avoid damage to human health by -27 and -10.7x10-4 DALY/1,000 kg DM at the
maximum recovery of resources. Wet air oxidation has the worst impact for this category
(~30x10-4 DALY/1,000 kg DM).
199
8.3.2.5 Consumers: Similarity to traditional products
For this indicator, WAO has a very high score since substitution of methanol
during denitrification process is optimized process (Luck 1999; Kolaczkowski et al. 1999;
Levec & Pintar 2007). The digested and composted sludge have a high score since these
products have a long history of successfully substituting synthetic fertilizers, despite some
drawbacks (X. Wang et al. 2008; Kilbride 2014; European Commission 2001b).
Incineration scores as moderate because of the lack of a heating distribution network
which makes it difficult to substitute traditional heating infrastructure in the UK (Which?
2015; Pöyry Energy 2009). Finally, pyrolysis was assigned a low score because its
products may differ from the traditional products due to the great dependency on the
pyrolytic process and sludge quality (Pokorna et al. 2009; Smith et al. 2009;
Thipkhunthod et al. 2006; Agrafioti et al. 2013).
8.3.2.6 Consumers: Presence of harmful substances
The presence of harmful substances in the products from pyrolysis and WAO is
very low and from incineration is low. ADG and composting have a moderate score due
to the concerns related to heavy metals (Hwang et al. 2007; Dabrowska & Rosińska 2012;
Mantovi et al. 2005; Park et al. 2010).
8.3.2.7 The rebound effect
There are no data on the potential for the rebound effect for the products from
sludge treatment so the discussion here is speculative. It could be argued that for most
products recovered from sludge the rebound effect could be expected to be low. Digested
sludge and compost are already in use and farmers generally optimize their use on land.
For electricity from incineration, the consumers would not be able to distinguish between
the different sources of electricity once fed into the grid so the rebound effect for this type
of treatment is less relevant. Similar applies to the liquid fuel from pyrolysis which would
be blended with conventional fuels. In the case of biochar, the product is similar to the
conventional charcoal with which the consumer is familiar; however, this would also
depend on the price of biochar compared to the conventional product. Finally, the product
from WAO would be used by wastewater treatment plants rather than consumers, with
200
the process operated for an optimum consumption of resources so that the rebound effect
is deemed very low.
8.4. Integrated sustainability assessment
This section presents the results of the integrated sustainability assessment of the
wastewater and sludge treatment options obtained considering environmental, economic
and social aspects simultaneously through MCDA (for the methodology, see Chapter 3).
The results are first discussed for the wastewater, followed by the sludge treatment
techniques. The results refer to the minimum, mean and maximum operating parameters
as specified in Chapters 5-7. The MCDA was carried out assuming that all the
sustainability indicators are of equal importance and therefore have the same weights. In
the base case, it was also assumed that the environmental, economic and social
sustainability dimensions have the same importance. However, this assumption was
tested through a sensitivity analysis to explore how the ranking of the options may change.
An arbitrary increase in the importance of five times for each sustainability dimension
was assumed in turn for this purpose. Note that the lower the score, the more sustainable
the option.
8.4.1. Sustainability assessment of advanced wastewater treatment techniques
As can be seen in Figure 63, for the equal importance of the environmental,
economic and societal dimensions of sustainability, ozonation is the best option for the
minimum operating requirements, scoring overall 0.38, around 30% lower than the
second-best options, GAC and NF. SPF is the least sustainable alternative. For the mean
operating requirements, the ranking changes and NF becomes the best alternative, scoring
0.28, followed by GAC (0.46) and ozonation (0.58); SPF remains the worst option. At
the maximum operating conditions, the ranking of the options remains the same. For
most options, the main contributor to the overall scores is their cost, followed by the
environmental impacts. The exception to this is ozonation, which is affected mainly by
the poor environmental and social performance.
201
Figure 63 – MCDA results for the advanced wastewater treatment techniques with the equal weights for
the sustainability indicators and environmental, economic and social dimensions of sustainability: (a)
minimum operating requirements; (b) mean operating requirements; and (c) maximum operating
requirements
0.11 0.12
0.25
0.14
0.33
0.22
0.20
0.00
0.10
0.17
0.19
0.24
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
0.540.51
0.64
0.38
Equal criteria weights
Minimum operating requirements
Su
stain
ab
ilit
y s
core
(a)
0.07 0.02
0.190.27
0.30
0.10
0.33
0.00
0.09
0.16
0.18
0.26
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
(b)
0.46
0.28
0.70
0.53
Equal criteria weights
Mean operating requirements
Su
stain
ab
ilit
y s
core
0.110.01
0.17
0.28
0.25
0.00
0.33
0.01
0.10
0.13
0.16
0.26
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
(c)
0.46
0.14
0.66
0.55
Equal criteria weights
Maximum operating requirements
Su
stain
ab
ilit
y s
core
202
If the environmental impacts are assumed to be five times more important than
the other two dimensions of sustainability, then all the techniques are comparable at the
minimum operating requirements (Figure 64a). The exception is SPF which is the least
sustainable wastewater treatment technique. However, at the mean and maximum
operating conditions, NF is the most sustainable option, followed by GAC; ozonation is
the worst option. If the economic dimension is five times more important than the other
options, then ozonation is the best alternative at the minimum and mean conditions and
is the second-best alternative after NF for the maximum operating conditions (Figure 65).
SPF is the worst option for the mean and maximum and GAC for the minimum operating
conditions.
If there is a strong preference for the social criteria (Figure 66), GAC is the best
option at the low and mean operating requirements, although only with a slight advantage
over NF for the mid-range conditions (scores of 0.35 and 0.39, respectively). At the
highest operating requirements, NF is the preferable technique. Ozonation is the least
sustainable for the mean and maximum conditions and SPF for the minimum operating
requirements.
Figure 64 – MCDA results for the advanced wastewater treatment techniques with the environmental
dimension of sustainability five times more important: (a) minimum operating requirements; (b) mean
operating requirements; and (c) maximum operating requirements
Figure 65 - MCDA results for the advanced wastewater treatment techniques with the economic
dimension of sustainability five times more important: (a) minimum operating requirements; (b) mean
operating requirements; and (c) maximum operating requirements
0.24 0.26
0.53
0.31
0.14 0.09
0.09
0.00
0.040.08
0.08
0.10
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
0.42 0.43
0.70
0.41
5x environmental preference
Minimum operating requirements
Su
stain
ab
ilit
y s
core
(a)
0.16
0.05
0.41
0.58
0.13
0.04
0.14
0.00
0.04
0.07
0.08
0.11
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
(b)
0.33
0.16
0.63
0.69
5x environmental preference
Mean operating requirements
Su
stain
ab
ilit
y s
core
0.24
0.03
0.37
0.590.11
0.00
0.14
0.04
0.06
0.07
0.11
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
(c)
0.39
0.09
0.58
0.70
5x environmental preference
Maximum operating requirements
Su
sta
ina
bil
ity
sco
re
0.05 0.050.11
0.06
0.71
0.470.44
0.00
0.04
0.080.08
0.10
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
0.80
0.600.63
0.16
5x economic preference
Minimum operating requirements
Su
stain
ab
ilit
y s
core
(a)
0.03 0.010.08
0.12
0.65
0.22
0.71
0.00
0.04
0.07
0.08
0.11
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
(b)
0.72
0.30
0.87
0.23
5x economic preference
Mean operating requirements
Su
stain
ab
ilit
y s
core
0.05 0.010.07
0.12
0.53
0.00
0.71
0.02
0.04
0.06
0.07
0.11
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
(c)
0.62
0.07
0.85
0.25
5x economic preference
Maximum operating requirements
Su
sta
ina
bil
ity
sco
re
203
Figure 66 - MCDA results for the advanced wastewater treatment techniques with the social dimension of
sustainability five times more important: (a) minimum operating requirements; (b) mean operating
requirements; and (c) maximum operating requirements
The above results are summarised in Figure 67, showing how the ranking of the
options changes with the assumptions on the preferences for the different sustainability
dimensions, ranging from equal to five times higher preference for each in turn. For
example, it can be seen from Figure 67a that the only significant change in the ranking of
the options occurs between ozonation and SPF, whereby the latter becomes a better option
at the mean operating conditions when the environmental criteria are 4.5 times more
important than the others and at the maximum conditions when the environment is 2.5
times more important. For the economic and social criteria, the change in the ranking of
the options is more pronounced across the operating conditions (Figure 67b-c). At the
minimum operating conditions, GAC becomes the worst option when the costs are
assumed two times more important than the other sustainability aspects. At the mean
operating conditions, ozonation is a better option than GAC if the costs are 1.3 times more
important and more sustainable than NF if they are around 3.5 times more important. For
the maximum requirements, the only reversal of the ranking is found for ozonation and
GAC if the costs are around 1.5 times more important.
For the social criteria being 2.5 times more important, GAC becomes the best
option at the minimum operating requirements. At the mean conditions, if the social
criteria are 4.5 times more important, there is a reversal in the ranking between ozonation
and SPF in favour of the latter and between NF and GAC also in favour of the latter.
Finally, at the maximum operating conditions, the only rank reversal is found between
SPF and ozonation if the social criteria are 2.5 times more important.
0.05 0.050.11
0.06
0.140.09
0.09
0.00
0.21
0.37
0.40
0.51
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
0.40
0.51
0.600.57
5x social preference
Minimum operating requirements
Su
stain
ab
ilit
y s
core
(a)
0.03 0.010.08
0.12
0.13
0.04
0.140.00
0.19
0.34
0.390.55
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
(b)
0.35
0.39
0.610.67
5x social preference
Mean operating requirements
Su
sta
ina
bil
ity
sco
re
0.05 0.010.07
0.12
0.11
0.00
0.140.00
0.21
0.28
0.34 0.55
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
Granular
activated carbon
Nanofiltration Solar photo-
Fenton
Ozonation
Environmental
Economic
Social
(c)
0.37
0.29
0.55
0.67
5x social preference
Maximum operating requirements
Su
stain
ab
ilit
y s
core
204
Environmental criteria preference: (a) minimum operating requirements; (b) mean operating
requirements; (c) maximum operating requirements.
Economic criteria preference: (d) minimum operating requirements; (e) mean operating requirements; (f)
maximum operating requirements.
Social criteria preference; (g) minimum operating requirements; (h) mean operating requirements; (i)
maximum operating requirements.
Figure 67 – Sensitivity analysis for the advanced wastewater treatment techniques for different weights of
importance for the sustainability dimensions
8.4.2. Sustainability assessment of sludge treatment techniques
As shown in Figure 68a&b, for the maximum and mean recovery of outputs,
anaerobic digestion and pyrolysis are the most sustainable options and composting the
least. If no products are recovered, anaerobic digestion remains the best option but
pyrolysis is now the worst alternative (Figure 68c). At the maximum and mean recovery
of products, the economic cost is the most important contributor to the total sustainability
of most options, except for pyrolysis where the environmental and social criteria
dominate.
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Environmental 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(a)
Minimum operating requirements
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Environmental 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(b)
Mean operating requirements
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Environmental 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(c)
Maximum operating requerements
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Economic 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(d)
Minimum operating requirements
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Economic 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(e)
Mean operating requirements
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Economic 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(f)
Maximum operating requirements
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Social 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(g)
Minimum operating requirements
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Social 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(h)
Mean operating requirements
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Social 5x
Granular activated carbon Nanofiltration
Solar photo-Fenton Ozonation
Su
sta
ina
bil
ity sc
ore
(i)
Maximum operating requirements
205
Figure 68 - MCDA results for the sludge treatment techniques with the equal weights for the sustainability
indicators and environmental, economic and social dimensions of sustainability: (a) maximum resource
recovery; (b) mean resource recovery; and (c) no product recovery
0.07
0.30
0.18 0.180.26
0.19
0.31
0.33
0.00
0.250.15
0.20
0.24
0.25
0.02
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(a)
0.41
0.81
0.43
0.53
Equal criteria weights
Maximum products recovery
Su
stain
ab
ilit
y s
core
0.75
0.06
0.28
0.16 0.190.26
0.09
0.33
0.29
0.00
0.24
0.20
0.22
0.21
0.14
0.07
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(b)
0.35
0.83
0.33
0.57
Equal criteria weights
Mean products recovery
Su
stain
ab
ilit
y s
core
0.66
0.07
0.22
0.11
0.25 0.250.00
0.12
0.07
0.33
0.18
0.19
0.13
0.25
0.18
0.10
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(c)
0.28
0.47
0.76
0.53
Equal criteria weights
No products recovery
Su
stain
ab
ilit
y s
core
0.43
206
If the preferences change so that either the environmental or costs criteria are
much more important (5x), the best and the worst options remain the same as for the equal
importance of all three sustainability dimensions (Figure 69 and Figure 70). However, if
the social impacts are five times more important than the environmental and economic,
the ranking changes (Figure 71). Now, WAO is the best option for all the products
recovery rates. Although for the mean and no recovery it is comparable with pyrolysis
and composting, respectively, for the maximum products recovery its sustainability score
is half that of the second-best option, anaerobic digestion. Further details on the change
in the ranking of the options for the preferences ranging from equal to five times greater
can be found in Figure 72. As can be seen, the ranking remains pretty much the same with
the change in the importance of the environmental and economic criteria but changes
more significantly if the social sustainability is most important. This is pronounced for
WAO which swings from the third or the fourth place to being the most sustainable option
(Figure 72c).
Figure 69 - MCDA results for the sludge treatment techniques with the environmental dimension of
sustainability five times more important: (a) maximum resource recovery; (b) mean resource recovery; and
(c) no products recovery
Figure 70 - MCDA results for the sludge treatment techniques with the economic dimension of
sustainability five times more important: (a) maximum resource recovery; (b) mean resource recovery; and
(c) no products recovery
0.15
0.64
0.40 0.38
0.55
0.08
0.13
0.14
0.00
0.11
0.06
0.08
0.10
0.11
0.01
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(a)
0.29
0.85
0.49
0.67
5x environmental preference
Maximum products recovery
Su
sta
ina
bil
ity
sco
re
0.64
0.13
0.60
0.340.41
0.56
0.04
0.14
0.12 0.00
0.10
0.09
0.09
0.09
0.06
0.03
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(b)
0.26
0.83
0.47
0.69
5x environmental preference
Mean products recovery
Su
stain
ab
ilit
y s
core
0.55
0.16
0.46
0.23
0.53 0.53
0.00
0.05
0.03
0.140.08
0.08
0.06
0.11
0.08
0.04
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(c)
0.24
0.57
0.75
0.65
5x environmental preference
No products recovery
Su
sta
ina
bil
ity
sco
re
0.37
0.030.13
0.08 0.08 0.11
0.41
0.670.71
0.00
0.54
0.06
0.08 0.10
0.11
0.01
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(a)
0.50
0.88
0.19
0.66
5x economic preference
Maximum products recovery
Su
stain
ab
ilit
y s
core
0.89
0.030.12
0.07 0.08 0.11
0.20
0.71
0.62
0.00
0.51
0.09
0.09
0.09
0.06
0.03
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(b)
0.32
0.92
0.14
0.65
5x economic preference
Mean products recovery
Su
stain
ab
ilit
y s
core
0.78
0.030.09
0.050.11 0.110.00
0.25
0.15
0.71
0.39
0.08
0.06
0.11
0.08
0.04
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested
sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(c)
0.11
0.40
0.90
0.54
5x economic preference
No products recovery
Su
stain
ab
ilit
y s
core
0.31
207
Figure 71 - MCDA results for the sludge treatment techniques with the social dimension of sustainability
five times more important: (a) maximum resource recovery; (b) mean resource recovery; and (c) no
products recovery
Environmental criteria preference: (a) maximum recovery; (b) mean recovery; (c) no recovery.
Economic criteria preference: (d) maximum recovery; (e) mean recovery; (f) no recovery.
Social criteria preference: (g) maximum recovery; (h) mean recovery; (i) no recovery
Figure 72 - Sensitivity analysis for the sludge treatment techniques for different weights of importance for
the sustainability dimensions
0.030.13
0.08 0.08 0.110.08
0.130.14
0.00
0.11
0.32
0.42 0.51
0.54
0.03
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(a)
0.43
0.68
0.62
0.25
5x social prpefenrece
Maximum products recovery
Su
stain
ab
ilit
y s
core
0.73
0.030.12
0.07 0.08 0.110.04
0.14
0.120.00
0.10
0.44
0.46
0.44
0.30
0.16
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(b)
0.51
0.72
0.38 0.37
5x social preference
Mean products recovery
Su
stain
ab
ilit
y s
core
0.63
0.03 0.090.05
0.11 0.11
0.050.03
0.140.08
0.410.28
0.53
0.39
0.21
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Agricultural
application of
anaerobic
digested sludge
Agricultural
application of
composted
sludge
Incineration Pyrolysis Wet air
oxidation
Environmental
Economic
Social
(c)
0.440.42
0.64
0.40
5x social preference
No products recovery
Su
stain
ab
ilit
y s
core
0.61
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Environmental 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(a)
Maximum products recovery
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Environmental 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(b)
Mean products recovery
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Environmental 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(c)
No products recovery
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Economic 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(d)
Maximum products recovery
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Economic 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(e)
Mean products recovery
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Economic 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(f)
No products recovery
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Social 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(g)
Maximum products recovery
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Social 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(h)
Mean products recovery
0.00
0.10
0.20
0.30
0.40
0.50
0.60
0.70
0.80
0.90
1.00
Equal preference Social 5x
Agricultural application of anaerobic digested sludge
Agricultural application of composted sludge
Incineration
Pyrolysis
Wet air oxidation
Su
sta
ina
bil
ity sc
ore
(i)
No product recovery
208
8.5. Chapter conclusions
The main findings of this chapter demonstrate that nanofiltration has the upper
hand concerning environmental impacts for advanced secondary effluent treatment,
followed by granular activated carbon. From the economic standpoint, ozonation has
clear advantage at lower operating requirements but is surpassed by nanofiltration at
mean-maximum requirements. For equal sustainability criteria nanofiltration followed by
granular activated carbon are the prefer options for advanced wastewater treatment.
Concerning sludge handling, anaerobic digestion is the best alternatives from the
environmental standpoint at any products utilization potential. Economically pyrolysis
has clear advantage among the other options, unless when not recovering products.
Regarding social aspects anaerobic digestion, pyrolysis and wet air oxidation have similar
results depending of the product utilization and preference given to the social criteria.
Yet, anaerobic digestion has preference at equal sustainability criteria preference.
Remarks concerning updates of the UK electricity grid
As commented in the conclusions of Chapters 5 and 6, updates in the UK
electricity grid in the last decade may influence the results obtained in this assessment. In
relation to the advanced wastewater treatment techniques, changes in the UK electricity
grid is not expected to promote shifts in the ranking of the alternatives, but accentuate
ozonation as the most sustainable alternative for minimum operating requirements (while
becoming more competitive to granular activated carbon at mean operating requirements,
especially for social preference – see Figure 67), and nanofiltration as the prefer option
(i.e. more sustainable) at mean-maximum operating parameters.
The present UK electricity grid may potentially affect the results for the most
sustainable sludge treatment techniques as following: (i) wet air oxidation can became
more competitive in relation to incineration and pyrolysis for the mean-maximum
products recovery in relation to environmental preference (see Figure 72a&b); and (ii)
this same option can be amongst most sustainable for strong social preference, close to
agricultural application of anaerobic digested sludge, at minimum-mean resource
recovery (Figure 72h&i).
209
9. CONCLUSIONS, RECCOMENDATIONS AND FUTURE WORK
The main findings of this research are first summarized for the target PPCP
compounds, followed by the sustainability evaluation of the wastewater and sludge
treatment techniques and its contextualization at the UK level. Finally, recommendations
are made for future research.
9.1. PPCP COMPOUNDS IN WWTPs
The conclusions discussed below refer to the findings detailed in Chapters 2 and
4-6.
9.1.1. The assessment methodology
Given that WWTPs are the main source of PPCP compounds in the environment,
it is important to quantify the amounts released in the treated wastewater and sludge. This
would provide a clearer picture of their contribution to the environmental concentrations
of these substances. Thus, in addition to offering an overview of this subject, the
methodology proposed in this work is expected to be useful for the following purposes:
determining discharge of PPCP compounds proportional to the consumption of
PPCPs;
estimation of expected concentrations of PPCP compounds in influents, effluents
and sludge from conventional WWTPs;
prediction of possible concentrations of PPCP compounds in freshwater bodies in
the proximity of WWTPs, providing initial data for environmental monitoring and
risk assessment;
provision of data for ecotoxicological tests for future evaluations of synergistic
effects of PPCP compounds;
definition of levels at which advanced wastewater treatment techniques can be
considered to remove effectively PPCP substances from effluents; and
development of policies, regulations and guidelines to control environmental
concentrations of PPCP compounds.
210
9.1.2. Specific findings for the target PPCP compounds
The findings for each of the 14 target PPCP compounds considered suggest the
following:
Acetaminophen has by far the highest concentration in influents and among the
highest in the sludge. Despite the very high removal rates which vary little among
the conventional treatments, its concentration in secondary effluents is still one of
the highest.
Diclofenac also has one of the highest influent concentrations. Its removal rate in
conventional treatments is moderate and highly variable. This combination results
in high concentrations in secondary effluents, albeit significantly lower than
acetaminophen. In sludge, this compound is amongst the ones with the highest
concentrations.
Ibuprofen is also among the compounds with the highest concentrations in
influents and sludge. However, its high removal rates by conventional WWTPs
means that its final concentrations in secondary effluents are significantly lower
than those of the other two analgesics;
Trimethoprim is present in moderate concentrations in influents and low
concentration in the sludge; its has moderate but highly variable removal rates in
the conventional WWTPs. This combination results in moderate concentrations
in secondary effluents.
Erythromycin has the highest concentration in influents among the antibiotics and
it also has one of the highest concentrations of the compounds assessed here.
Together with its low to negligible removal rates in conventional treatment plants,
this means that its concentrations in the secondary effluents are higher, on
average, than the analgesics. In sludge, this compound is present in high
concentrations, similar to acetaminophen.
Sulfamethoxazole has moderate to low influent concentrations and moderate
removal rates by conventional WWTPs. Accordingly, its concentration in
secondary effluents is among the lowest of all the compounds assessed here. In
sludge, it is present at low-range concentrations.
Metoprolol has one of the lowest concentrations in influents, and although its
removal rates by conventional treatments are low or negligible, its concentrations
in the secondary effluents and sludge are still among the lowest.
211
Gemfibrozil is present in moderate to low concentrations in influents and
moderate albeit highly variable removal rates in conventional treatments.
Therefore, its presence in secondary effluents is in the low range. In sludge, this
compound is present at low-range concentrations.
Bezafibrate: this compound is present in mid-range concentrations in influents
and exhibits moderate removal rates in conventional treatment. Thus, its
concentration in secondary effluents is expected to fall within the middle to low
range of the target compounds.
Carbamazepine is found in high concentrations in influents, comparable to those
of analgesics. Its removal rates vary significantly in conventional treatments,
showing recalcitrant behaviour. Hence, its concentrations in secondary effluents
are the highest, on average, among the target compounds. In sludge, this
compound is present in the mid-range concentrations, similar to trimethoprim.
Oestrone has the second lowest concentrations in influents and the most variable
removal rates by conventional treatments, exhibiting a considerably recalcitrant
behaviour. Therefore, its average concentration is expected to be at the upper limit
of the low-concentration compounds in secondary effluents. Its concentration in
sludge is in the mid-range.
17β-oestradiol has the lowest concentrations in influents and in sludge, and high
removal rates by conventional wastewater treatments. Consequently, it has the
lowest concentrations in secondary effluents.
Triclosan has a moderate concentration in influents, but the highest in sludge. Its
moderate-to-high removal rates by conventional treatment means its
concentration in secondary effluents is low.
Caffeine has the second highest concentrations in influents and sludge. Due to its
very high removal rates by secondary treatment, its concentration in secondary
effluents is among the lowest assessed here.
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9.1.3. Removal of PPCP compounds by advanced wastewater and sludge treatment
techniques
The main findings of this part of the work were as follows:
Granular activated carbon: this is the preferred alternative in terms of removal
potential of PPCP compounds from secondary effluents, with removal rates of all
the target compounds typically exceeding 90%, regardless of their
physicochemical properties, except for compounds with a very low Kow.
Nanofiltration: its removal potential of PPCP compounds is moderate (and the
least effective of all the alternatives discussed here), rarely removing more than
70% of compounds from secondary effluents. Diclofenac, hormones and triclosan
are particularly difficult to remove by nanofiltration.
Solar photo-Fenton: this alternative shows a significantly higher removal potential
of PPCP compounds from secondary effluent than nanofiltration, but considerably
lower than granular activated carbon, being particularly ineffective for diclofenac
and carbamazepine. Moreover, harmful by-products can also be generated during
treatment.
Ozonation: this technique has the second most promising potential for the removal
of PPCP compounds, only slightly lower than granular activated carbon.
However, its potential to remove ibuprofen and triclosan seems to be low.
Moreover, it also generates harmful by-products during treatment.
Agricultural application of anaerobic digested sludge: the removal potential of
PPCP compounds by anaerobic digestion of sludge is unknown/highly variable.
Agricultural application of composted sludge: the removal potential of PPCP
compounds by composting is also unknown/highly variable.
Incineration: this sludge disposal method removes completely the PPCP
compounds.
Pyrolysis also removes the PPCP compounds completely.
Wet air oxidation: like incineration and pyrolysis, it removes completely PPCP
compounds.
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9.1.4. Ecotoxicological potential of PPCPs in wastewaters and sludge
Current water quality regulations, such as the Water Framework Directive, WFD
2000/60/EC, impose strategies against water pollution by establishing a list of 33 priority
substances considered to pose a significant ecotoxicological threat to or through aquatic
environments (see topic 2.4.2). The findings of this research could be helpful in guiding
the development of future regulations for the inclusion of PPCP compounds among
contaminants to be monitored more closely, and perhaps future inclusion in the list of
priority substances. Regulations on sewage sludge handling routes are also becoming
stricter, and in some regions of Europe, PPCP compounds are already included among
substances for closer monitoring in sludge.
As discussed in Chapter 2, there are still gaps in the ecotoxicological assessment
of PPCP compounds with respect to their synergetic effects, bacteria resistance, and other
issues. However, the conclusions of this study about the potential ecotoxicity of the target
PPCP compounds in freshwater bodies at the concentrations released by WWTPs,
according the USEtox methodology, suggest the following risk levels for the target PPCP
compounds:
negligible risk: diclofenac, ibuprofen, trimethoprim, sulfamethoxazole;
low to negligible risks: carbamazepine, oestrone;
considerable risk: erythromycin, triclosan; and
very high risks: 17β-oestradiol.
As demonstrated in Chapter 5, freshwater ecotoxicity generated by the advanced
wastewater treatment techniques are similar to or higher than the freshwater ecotoxicity
of the PPCP compounds in the treated effluent. Therefore, their use solely for controlling
the levels of PPCP compounds in WWTPs effluents is not recommended. Likewise,
PPCPs in sludge have negligible freshwater ecotoxicity potential compared to the
potential freshwater ecotoxicity generated by the sludge treatments and heavy metals
(Chapter 6).
214
9.2. SUSTAINABILITY ASSESSMENT
The conclusions in this section refer to the findings presented in Chapters 5-8.
9.2.1. Life cycle environmental impacts
9.2.1.1. Advanced wastewater treatment techniques
The results in Chapter 5 suggest that nanofiltration and granular activated carbon
have the lowest life cycle environmental impact. The former is the best option for 10 and
the latter for six impacts out of 18 impact categories considered. Furthermore, they are
the only techniques that do not generate harmful by-products and have considerable
removal rates of heavy metals from effluents, thus being more suitable for indirect and
direct potable reuse of treated wastewater (see section 2.5.3.1). In this regard, preference
is given to granular activated carbon because it can reduce more efficiently the potential
ecotoxicity of PPCP compounds released into freshwater bodies.
9.2.1.2. Sludge treatment techniques
Agricultural application of anaerobic digested sludge is the best option for 13 out
of 18 impact categories for all the resource recovery potentials considered in this work.
At a recovery of its products above 50%, pyrolysis is the second-best alternative, with
four impacts lowest than any other option. At lower recovery rates, incineration is the
second-best option.
9.2.2. Life cycle costs
9.2.2.1. Advanced wastewater treatment techniques
Based on the findings in Chapter 7, the lowest cost alternatives are ozonation and
nanofiltration with the average costs estimated at £112 and £144 per 1,000 m3 of treated
secondary effluent. Granular activated carbon would only fall within this range if the
coagulant (iron sulphate) prices were within the lower cost range considered in this
215
evaluation, while solar photo-Fenton appears to be economically unattractive across all
the conditions considered.
9.2.2.2. Sludge treatment techniques
The pyrolysis is economically the most attractive alternative, generating profits of
£29/1,000 kg of dry matter at the average recovery of the products. This is followed by
the anaerobically digested sludge, with an estimated cost of around £9/1,000 kg of dry
matter for the average recovery of the products, significantly lower than the cost of
composting and incineration (£117 and £89, respectively). The recovery rates of their
respective products radically change the ranking. For example, the agricultural
application of anaerobic digested sludge has the costs comparable to incineration at the
lowest recovery rates. Pyrolysis achieves a profit of £256/1,000 kg of dry matter at the
complete recovery and sales of their products, although these are currently not realistic
suppositions. Composted sludge and incineration seems not to be economically feasible
compared to the other alternatives even at a considerable recovery of the products.
9.2.3. Social sustainability assessment
9.2.3.1. Advanced wastewater treatment techniques
As discussed in Chapter 8, social impacts at the national level indicate that,
generally, granular activated carbon has the lowest and ozonation the highest impact. For
water suppliers, all the advanced technologies are equivalent, but for consumers, granular
activated carbon is again the preferred alternative, especially given its lowest human
health concerns (e.g. lowest DALY and PPCP compounds concentration). Overall, in
terms of the potential social impacts of the wastewater treatment techniques evaluated
here, the options can be ranked from best to worst as follows: granular activated carbon,
nanofiltration, solar photo-Fenton and ozonation.
216
9.2.3.2. Sludge treatment techniques
Wet air oxidation has the lowest and incineration the highest social impacts across
all the recovery rates. Composting and pyrolysis are comparable to incineration at the
mean and maximum recovery rates of products, respectively. From the supplier’s
standpoint, the agricultural application of sludge and wet air oxidation are the most
practical alternatives. For costumers, anaerobic digestion and incineration offers a slight
advantage over the other techniques because of their low threat to human health and the
fact that the products resemble traditional products. However, incineration suffers from
strong public opposition mainly because of the perception that it is damaging to health,
despite the scientific findings to the contrary (European Commission 2001a; European
Commission 2001b).
9.2.4. Integrated sustainability assessment
9.2.4.1. Advanced wastewater treatment techniques
The integrated sustainability assessment of the treatment options carried out via
MCDA in Chapter 8 suggests that:
at equal weights for the environmental, economic and social aspects, ozonation is
the most sustainable option when secondary effluents are deemed of good quality.
When secondary effluents are of medium quality, nanofiltration and granular
activated carbon are the best alternatives. For the low effluent quality,
nanofiltration becomes clearly preferable over all other alternatives;
increasing the preference for the environment over the other two sustainability
aspects by (an extreme) five times, makes nanofiltration the most suitable
alternative, unless the secondary effluent is of a high quality, in which case
granular activated carbon and ozonation are comparable to nanofiltration;
if the economic aspect is five times more important, ozonation is the best option
for the high to medium quality of secondary effluent, while nanofiltration is
preferable for the low quality; and
if the social impacts are most important, granular activated carbon is the most
sustainable option unless the effluent is of a low quality (maximum operating
requirements) in which case nanofiltration is the best alternative.
217
9.2.4.2. Sludge treatment techniques
Also, based on the findings in Chapter 8, the following conclusions can be drawn
on the overall sustainability of the sludge treatment techniques:
at equal criteria weights, agricultural application of anaerobic digested sludge and
pyrolysis are the best alternatives for the complete and mean recovery of the
products. At no recovery, the former is most sustainable while pyrolysis becomes
the worst option;
if the environmental impacts are five times more important, anaerobic digestion
remains the best option at any recovery potential of the products. Composting is
the worst option for the maximum and mean recovery rate and pyrolysis is again
the worst option;
assuming a five times greater preference for the economic costs, pyrolysis is by
far the best alternative unless there is no recovery of products, in which case it is
the worst option and anaerobic digestion is most sustainable; and
if the social impacts are most important, wet air oxidation is the best option for all
the assumed rates of products recovery. Composting and incineration are the least
preferred options for the high and mean recovery of products while at no recovery,
pyrolysis and incineration are the worst alternatives.
218
9.3. RECOMMENDATIONS
This section considers the implication of the findings of this work at the UK level
and makes recommendations for possible future implementation.
9.3.1. Implications for advanced wastewater treatment in the UK
9.3.1.1. Freshwater ecotoxicity potential from PPCPs compounds
From roughly 2.74x107 m3/d of wastewater generated in the UK (assuming q =
428 L/d), the adoption of advanced wastewater treatments techniques could potentially
avoid the discharge of over 71 tonnes/year of PPCP compounds to the environment
(calculated using data from Table 10). However, the results from this research suggest
that:
The adoption of advanced wastewater treatment methods solely to control PPCP
compounds and their impact on freshwaters would create similar or greater
freshwater ecotoxicity potential than their removal, particularly if solar photo-
Fenton or ozonation are used. Hence, their use exclusively for this purpose is not
recommended.
Irrigation of agricultural soils with secondary effluent would have a negligible
impact on freshwater ecotoxicity from PPCP compounds compared to the impact
generated by the advanced treatments. Therefore, their use to reduce the
concentration of PPCP compounds in irrigation water is also not recommended;
Consequently, advanced wastewater treatment techniques are only recommended
for application in densely populated areas for safer water reuse and, ultimately,
reuse as potable water. This requires further research as discussed in the future
work section.
219
9.3.1.2. Wastewater reuse potential
In near future WWTPs could play a key-role in sustainable development by
promoting rational, efficient and reliable use of water at the basin level. Although there
is perception of abundance of this resource in the UK, forthcoming uncertainties such as
climate change necessitate consideration of a wastewater reuse potential. The first attempt
at a large-scale wastewater reuse (for indirect purposes) in the UK is currently under
review for the Thames river basin surrounding London due to the need to balance water
deficit in the region. Reports commissioned by Thames Water suggest that granular
activated carbon is the most suitable alternative from the technical perspective (IERP
2013), agreeing with this work’s suggestion. The findings of this work also suggest that
the costs of advanced wastewater treatment for producing potable water are significantly
lower than water desalination and could be competitive in the future with conventional
potable water production.
9.3.2. Implications for resource recovery from sludge in the UK
9.3.2.1. Nutrients recovery
If all 1.4 million tonnes of sewage sludge (dry matter) generated annually in the
UK was used in agriculture after anaerobic digestion, the production of 21,200 tonnes of
phosphate could potentially be avoided annually (assuming the content in the digested
sludge of 15 kg/1,000 kg dry matter adopted in this work). From these numbers, this work
estimate that around 12,700 tonnes on nutrients are nowadays avoided in the UK by
considering anaerobic digestion in sludge handling, equivalent to 6.4% of the amount of
nutrients consumed and potentially reaching 10.6% of UK’s phosphorous requirements
(DEFRA 2015a). Given that phosphorous is becoming a scarce resource, these savings
are significant. In addition, recovering this nutrient from sludge improves national self-
sufficiency and food security and it is thus recommended for further expansion.
220
9.3.2.2. Energy recovery
Based on the above-mentioned amount of sewage sludge generated annually in
the UK, this work estimates that if the maximum amount of electricity is recovered in
anaerobic digestion from all the sludge generated in the UK, 1.12 TWh of electricity
would be generated annually. This is less than 0.4% of the currently consumed electricity
figure, of around 310 TWh/year (DECC 2014), and therefore would not significantly
contribute to the UK electricity grid. The same can be said for incineration.
If all the sludge was pyrolyzed, 325,000 tonnes of charcoal and 56,000 tonnes of
fuel oil could potentially be avoided at the maximum recovery of pyrolysis products.
Assuming the energy content of 18 and 30 MJ/kg for these products, respectively, this
would represent a saving of around 0.18 Mtoe annually, or less than 0.3% of the UK
domestic consumption of fuels (Park et al. 2008; Bridle & Pritchard 2004; DECC 2015).
Yet, since pyrolysis is still at the very beginning of commercial application, these figures
are far from feasible estimative for the next decade or so. Consequently, anaerobic
digestion is the route with the highest potential positive impact on energy recovery and
is, therefore, recommended for improving energy security in the UK.
9.3.2.3. Minimization of impacts in the EWF nexus
This work has proposed a novel methodology for assessing the impacts in the
EWF nexus at the national level which is generic and applicable beyond the technologies
considered here. The specific application of the methodology to the wastewater and
sludge treatment options suggests that:
granular activated carbon has the lowest negative impacts on the EWF nexus at
the mean operating requirements; at the minimum and maximum requirements,
solar photo-Fenton is the prefer technique; and
agricultural application of digested sludge could contribute to improving food and
energy security besides contribute to water security, and is recommended as the
best sludge treatment option to minimize the impact on the EWF nexus.
221
9.4. FUTURE WORK
9.4.1. Further work on PPCPs in WWTPs
Water pollution is a topic of growing concern worldwide and due to a stricter
control of priority pollutants, emerging and unregulated compounds, such as those from
PPCPs, are starting to be monitored, at least in developed nations. Further work on the
following topics would contribute substantially to a better understanding of the behaviour
of these substances in conventional wastewater treatments and their contribution to
freshwater ecotoxicity:
more data on the frequency of detection and increased sampling of influents and
effluents of WWTPs, especially during different climatic conditions;
more accurate and range definition of Kd values to better understand the sorption
behavior of these substances and further studies over their sorption and
degradation at different operating parameters;
consolidation of knowledge regarding the most recalcitrant and toxic compounds
from the over 3,000 currently available; and
quantification and characterization of metabolites and transformation products in
secondary effluents (e.g. more precise mass balances).
9.4.2. Improvements in life cycle assessment
The following future work is needed to improve life cycle assessments of
advanced wastewater and sludge treatment techniques:
characterization factors for the most recalcitrant and/or hazardous PPCP
compounds and other ECs to enable estimations not only of freshwater ecotoxicity
potential but also terrestrial and marine ecotoxicity and human toxicity;
study of heavy metals in influents and effluents, and estimation of their
ecotoxicity and human toxicity to be used in future life cycle assessments
(calculation of characterization factors);
evaluation of different combinations of the advanced treatment techniques to
optimize removal of emerging contaminants and the operating parameters;
sustainability evaluation of the integrated conventional and advanced treatments
and comparison to water treatment plants;
222
market research on the potential for commercialization of products from advanced
sludge treatment techniques; and
studies on public acceptance of each advanced treatment and their products.
9.4.3. Future role of advanced wastewater and sludge treatment plants
Ultimately, advanced wastewater and sludge treatment techniques can be used as
means for reclaiming water, energy and nutrients from wastewaters, helping towards
achieving the goals of sustainable development (see Figure 73). To promote their use and
understand better their future role, the following further research is recommended in the
UK and elsewhere:
compilation and evaluation of the quality of secondary effluents and thickened
sludge in the UK (temporally and spatially), enabling better contextualization of
this work results;
definition of potential WWTPs/regions showing possibility / necessity of indirect
and direct wastewater reuse, e.g. higher effluent quality, favourable surrounding
topography, expected increased water demand, lack of new fresh water sources,
affected by climate change, etc.;
definition of potential WWTPs/regions that show possibility / necessity of
resource recovery from sludge, e.g. limited disposal sites, presence of
infrastructure, expected increased energy demand, etc.;
initiation of schemes and data collection for implementation of wastewater reuse
at large scale for assuring its sustainability, e.g. regional effluent distribution
networks, evaluation of human health risks posed by direct wastewater reuse,
climate change threats to infrastructure, etc.;
further evaluation of potential reduction of impacts from WWTPs on freshwaters
and coastal areas; and
further studies concerning the potential contribution of resources recovered from
sludge for achieving targets for renewable energy sources.
223
Figure 73 – Concept of the ultimate role of advanced wastewater and sludge treatment techniques in the
rational use of resources in the EWF nexus
Water
EnergyFood
Natural environment
“ ”
Ev
entu
al
224
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261
11. SUPPLEMENTARY INFORMATION
11.1. Chapter 4 supplementary information
Analgesics/anti-inflammatories
Acetaminophen
Acetaminophen is a steroidal anti-inflammatory substance utilized for the relieve
all sorts of pain and reduce fever. It is a major ingredient in numerous cold and flu
remedies, and in combination with opioid analgesics can also be utilized in the
management of more severe kinds of pain (Larson et al., 2005). This substance is listed
as a “core medicine” by the World Health Organization, as necessary drug to meet the
minimum medical needs of a basic healthcare system (WHO, 2013). It is one of the top
three drugs prescribed in Europe, with a per capita consumption around 15.683 g/year in
the UK and around 4.456 g/year in Germany (An et al., 2009; Roig, 2010). The expected
unchanged human excretion for this compound is 3% (Khan and Ongerth, 2004).
Diclofenac
Diclofenac is a nonsteroidal anti-inflammatory drug (NSAID), administered to
reduce inflammations but also as an analgesic for soften pain in certain conditions. It is
available as generic drug in several formulations, and its exact mechanism of action is not
yet entirely known. Renal failure is a side effect of diclofenac overdoses in mammals
(Kallio et al. 2010). This substance is commonly sold over- the-counter (OTC) in many
countries and one of the most widely used pharmaceutical drugs, although having lower
consumption compared to other analgesics. Its per capita consumption revolves around
0.538 g/year in the UK and 0.506 g/year in Poland. The unchanged human excretion for
this compound can reach 15% (Carballa et al. 2005; Roig 2010).
Ibuprofen
Ibuprofen is also a NSAID used for pain relief, fever reduction and against
swelling, typically acting as a vasoconstrictor. It has an antiplatelet effect though
relatively mild and somewhat short-lived compared with aspirin or other prescription
antiplatelet drugs (Buser et al. 1999). This substance it is listed as a “core medicine” by
the World Health Organization (2013). Its per capita consumption is believed to be around
2.843 g/year in the UK and 3.425 g/year in France, although presumably higher since it
262
is also sold OTC. The expected unchanged human excretion for this compound varies
from 8 to 16% (Carballa et al. 2005; Roig 2010).
Antibiotics
Trimethoprim
Trimethoprim is a bacteriostatic antibiotic used in the prophylaxis and treatment
of urinary tract infections (Dodd & Huang 2007). The per capita consumption of this
compound revolves around 0.140 g/year in the UK and Germany. The expected
unchanged human excretion usually varies from 43 up to 73% (Khan & Ongerth 2004;
Roig 2010).
Erythromycin
Erythromycin is a macrolide antibiotic with antimicrobial spectrum similar to
penicillin, active against a wide diversity of bacteria that cause an extensive variety of
infections of the upper or lower airways, soft tissues, eyes or ears. It may also be used to
treat certain sexually-transmitted infections, oral and dental infections and in prevention
of infections caused by surgeries or burns (Göbel et al. 2005). Conversely to most
antibiotics, erythromycin is highly metabolized after intake by the organism. The per
capita consumption of this compound is around 0.536 g/year in the UK and 0.257 g/year
in Germany. The expected unchanged human excretion ranges between 4 and 10%
(Carballa, Omil, et al. 2008; Roig 2010).
Sulfamethoxazole
Sulfamethoxazole is a sulphonamide bacteriostatic antibiotic. It is most often used
combined with trimethoprim in a 5:1 ratio due synergistic effects, reducing the
development of bacterial resistance seen when either drug is managed separately. It is
commonly used to treat urinary tract infections, but can also be applied as an alternative
to amoxicillin-based antibiotics to treat sinusitis or toxoplasmosis (Beltrán et al. 2008;
Dantas et al. 2008). The per capita consumption of this compound is around 0.018 g/year
in the UK and 0.652 g/year in Germany. Its expected unchanged human excretion
revolves around 20 to 30% (Carballa et al. 2005; Roig 2010; Khan & Ongerth 2004).
263
Cardiovascular beta-blocker
Metoprolol
Metoprolol is a selective β1 receptor blocker used in treatment of several diseases
of the cardiovascular system, especially hypertension. It can also be used for a number of
conditions, including acute myocardial infarction, tachycardia, heart failure and adjunct
in treatment of hyperthyroidism (Huggett et al., 2002; Maurer et al., 2007). The per capita
consumption of this compound is around 0.036 g/year in the UK, 0.148 g/year in France
and 0.995 g/year in Germany (Roig, 2010). The expected unchanged human excretion for
this compound is 7% (Khan and Ongerth, 2004).
Lipid regulators
Gemfibrozil
Gemfibrozil is the generic name for an oral drug used to lower lipid levels in
organisms. It belongs to a group of drugs known as fibrates, acting as an activator of
nuclear receptors involved in the metabolism of carbohydrates and fats, as well as adipose
tissue differentiation (Rubins et al., 1999). The per capita consumption of this compound
is around 0.017 g/year in the UK and 0.072 g/y in Germany (Roig, 2010). The expected
unchanged human excretion for this compound is 76% (Khan and Ongerth, 2004).
Bezafibrate
Bezafibrate belongs to the fibrates drug class, used for the treatment of
hyperlipidaemia. It helps to lower low-density lipoprotein and triglyceride in the blood,
and increase high-density lipoprotein (Kyrklund et al., 2000). The per capita consumption
of this compound is around 0.146 g/year in the UK and 0.382 g/year in Germany (Roig,
2010). The expected unchanged human excretion for this compound is 50% (Carballa et
al., 2008a).
Psychiatric drug
Carbamazepine
Carbamazepine is a substance for reducing abnormal electrical activity in the
brain through chronic administration at high doses (Kosjek et al. 2009). It can be
prescribed singly or in combination to control certain types of seizures in patients with
epilepsy among other nerve pains or indicated to treat mania or mixed episodes in patients
with bipolar disorder. Its per capita consumption revolves around 0.767 g/year in the UK,
264
0.614 g/year in France, 0.838 g/year in Poland and 0.983 g/year in Germany. The
expected unchanged human excretion for this compound ranges from 2 up to 31% (Khan
& Ongerth 2004; Carballa et al. 2005; Roig 2010).
Estrogenic hormones
Oestrone
Oestrone is an estrogenic hormone (one of many natural ones although the least
abundant of the three main hormones in human body) administered for diverse reasons or
naturally secreted by the ovary as well as adipose tissue. It is an odourless white solid
crystalline powder, relevant to organism function because converts to estrone-sulphate, a
long-lived derivative which acts as a reservoir that can be converted when the more active
17β-oestradiol (Baronti et al. 2000; Vandenberg et al. 2012).
17β-oestradiol
The 17β-estradiol is a hormone with two hydroxyl groups in its molecular
structure, about ten times more potent than estrone in its estrogenic effect. Except during
the early follicular phase of the menstrual cycle, its serum levels are somewhat higher
than estrone during the reproductive years of the human female. It is also present in males
as a metabolic product of testosterone. The behaviour and presence in wastewater
treatment plants are like the ones related to estrone.
Antiseptic
Triclosan
Triclosan is a halogenated phenol with broad antimicrobial spectrum. It is an
ingredient of many disinfectants, soaps, detergents, plastic additives, and innumerable
veterinary, industrial and household products. It is effective against the propagation of
many types of bacteria and certain types of fungi. Unlike some other organic chlorine
compounds, its use is not yet a regulated compound (Dann & Hontela 2011; Katz et al.
2013). Potential human health issues surrounding the use of triclosan includes microbial
resistance, skin irritations, endocrine disruption, increasing rates of allergies and the
formation of carcinogenic by‐products (Aranami & Readman 2007; Yazdankhah et al.
2006).
265
Stimulant
Caffeine
Caffeine is a bitter and white xanthine alkaloid. It is found in the seeds, leaves,
and fruit of some plants, where it acts or as a natural pesticide that paralyzes and kills
certain insects feeding on the plants, or as enhancing the reward memory of pollinators.
In humans, caffeine acts as central nervous system stimulant, temporarily warding off
drowsiness and restoring alertness (Buerge et al., 2003). This compound is an important
component in many pharmaceuticals, since enhances the effect of certain analgesics.
However, it is most commonly consumed by humans in infusions extracted from the seed
of the coffee plant and the leaves of the tea bush, as well as from various foods and drinks
containing products derived from the kola nut. It is the world's most widely consumed
psychoactive drug (average 26 g/year), but unlike many other psychoactive substances, it
is legal and unregulated in nearly all parts of the world (Buerge et al., 2003; Kendler et
al., 2007).
266
Figure 74 - Box plots for IMinf,i values showing the ranges of data found in the literature for different world
regions. White dots represent estimated mean values, horizontal lines median values and small red dots the
outliers
Europe
Asia Europe
North America
Asia
Europe
North America
Asia
Europe
Australia
Asia
EuropeAsia Europe
Australia
Asia
Europe
Asia Europe
Asia
Europe
Asia
Europe
Asia
Europe
North AmericaAsia
North America
Asia North AmericaAsia
267
𝑁𝐼𝑀𝑖𝑛𝑓,𝑖 =𝐼𝑀𝑖𝑛𝑓,𝑖 −𝐼𝑀𝑖𝑛𝑓,𝑖(min)
𝐼𝑀𝑖𝑛𝑓,𝑖 (max)−𝐼𝑀𝑖𝑛𝑓,𝑖 (min) (31)
NIMinf,i– normalized value of influent concentration of PPCP compound i
IMinf,i – data point for influx of PPCP compound i into WWTP found in the literature
(mg/inhab.year)
IMinf,i (min) – minimum value of IMinf,I (mg/inhab.year)
IMinf,i (max) – maximum value of IMinf,I (mg/inhab.year)
𝑊𝑟𝑒𝑔𝑖𝑜𝑛 =∑ 𝑁𝐼𝑀𝑖𝑛𝑓,𝑖 (𝑟𝑒𝑔𝑖𝑜𝑛)
𝑛𝑖
𝑛𝑟𝑒𝑔𝑖𝑜𝑛 (32)
Wregion – total score for a region
NIMinf,i (region) – normalized value of IMinf,i in a region
nregion – total number of data points in a region.
Table 36 - Estimated annual per-capita influx into WWTPs of target PPCP compounds (dataset A)a
Annual per-capita discharge of target PPCP compounds, IMinf,i (mg/inhab.year)a
Sourceb Acetami-
nophen
Dic
lo-
fen
ac
Ibu-
prof
en
Trimet-
hoprim
Eryth-
romycin
Sulfame-
thoxazole
Meto-
prolol
Gem-
fibrozil
Beza-
fibrate
Carbama
-zepine
Oest
rone
17β-
oestradi
ol
Tricl
osan
Caff
eine
Thomas & Foster
(2004) 9.9
9
201.
97 63.78
931.
20
Atkinson et al.
(2012) 9.80 9.80
Behera et al.
(2011) 1,122.38
22.
45
329.
23 29.93 13.47 0.75 29.93 14.97 7.48 0.60 82.31
374.
13
Sim et al. (2010) 1,056.15 1.3
2
132.
02 99.01 2.64 39.61 792.
11
Nakada et al.
(2006) 173.
47 17.35 8.67 4.34
130.1
0
Nakada et al.
(2007) 53.9
6 10.79 5.40 2.70 74.19
Leung et al.
(2012) 28.23 141.13 14.11
Xu et al. (2012) 154.71 8.99
Gulkowska et al.
(2008) 30.16 78.98
Zhou et al. (2012) 12.17 6.08
Sui et al. (2010) 46.
01 39.43 13.14 5.26 5.26 19.72 788.
67
Gracia-Lor et al.
(2012) 4,022.30
38.
69
106
5.80 7.30 32.85 15.33 5.84
Tauxe-Wuercsh et
al. (2005) 280
.42
413.
24
Lindqvist et al.
(2005) 109
.85
146
1.02 54.93
Kasprzyk-Hordern
et al. (2009) 25,088.85
8.3
2
199.
76 260.40 191.44 3.57 9.51 49.94 200.95
Zhou et al. (2009) 391
.14 71.84 730.40
Zorita et al. (2009) 30.
53
915.
82 2.65 0.40
Lindberg et al.
(2005) 1.98 3.25
Baronti et al.
(2000) 8.93 2.23
Watkinson et al.
(2007) 24.82 26.28
268
Table 37 - Estimated removal rates for the target PPCP compounds (dataset B)a.
Removal rate, Rrate,i (%)
Sourceb Acetami-
nophen
Dicl
o-
fena
c
Ibu
-
pro
fen
Trime-
thoprim
Eryth-
romycin
Sulfame-
thoxazole
Meto-
prolol
Gemfi-
brozil
Beza-
fibrate
Carbama
-zepine
Oest
rone
17β-
oestradi
ol
Tricl
osan
Caff
eine
Gao et al. (2012) 90.91 -100.00 98.33
Conkle et al.
(2008) 99.95 99.
19 92.42 90.48 -10.30 -50.00 99.8
8
Thomas & Foster
(2004) 99.9
8
99.
79 97.33
99.9
1
Batt et al. (2007) 96.71 77.50
Yang et al. (2011) 99.94 95.4
5 99.45
54.10 20.59 83.85 -8.70 95.74 99.9
1
Lishman (2006) 5.00 95.
50 44.44 66.67 100.0
0
Atkinson et al.
(2012)
-
100.0
0
94.00
Behera et al.
(2011) 99.87
86.6
7
93.
18 80.00 0.00 20.00 90.00 20.00 60.00 99.75 81.82
99.2
0
Sim et al. (2010) 100.00 0.00 99.
99 80.00 99.50 33.33 99.6
7
Choi et al. (2008) 99.97 77.27 69.23 60.87 98.8
3
Nakada et al.
(2006) 98.
75 37.50
-
25.00 50.00 83.33
Nakada et al.
(2007) 97.
50 62.50 50.00 90.00 78.18
Hashimoto (2007) -
33.33 83.33
Leung et al.
(2012) 5.00 0.00 30.00
Xu et al. (2012) 13.95 40.00
Gulkowska et al.
(2008) -9.52 7.27
Zhou et al. (2012) 85.00 95.00
Sui et al. (2010) 42.8
6 66.67 10.00 25.00 75.00 20.00 99.8
3
Gracia-Lor et al.
(2012) 100.00
35.8
5
100
.00 10.00 88.89 -133.33 25.00
Carballa et al.
(2004) 64.
05 56.90
-
100.0
0
Radjenovic et al.
(2009) 98.89
20.4
5
98.
11 40.00 34.15 77.78 25.00 0.00 79.80 0.00
Santos et al.
(2006) 88.
42 -66.67 42.8
6
Tauxe-Wuercsh et
al. (2005) 0.00
78.
57
Maurer et al.
(2007) 33.33
Lindqvist et al.
(2005) 65.0
0
91.
73 34.00
Kasprzyk-
Hordern et al.
(2009)
94.45
-
42.8
6
84.
52 47.49 13.66 66.67 12.50 45.24 -47.93
Jones et al. (2007) 95.00 87.
50
Zhou et al. (2009) 91.8
4 83.33 54.10
Roberts &
Thomas (2006) 99.99
65.3
1
44.
83 -53.85 -81.82
Zorita et al.
(2009)
-113.
04
98.
70
-250.0
0
16.67
Lindberg et al.
(2005) 12.00 53.66
Baronti et al.
(2000) 25.00 80.00
Watkinson et al. (2007)
85.29 25.00
269
Table 38 - Estimated daily influx for the target PPCP compounds for a TP serving a population “p”
Compound αrange,i (g/day)
αmin αmean αmax
Acetaminophen 2.9397E-03 x p 2.8623E-02 x p 5.4307E-02 x p
Diclofenac 2.7397E-05 x p 2.2192E-04 x p 4.1644E-04 x p
Ibuprofen 4.4658E-04 x p 1.5288E-03 x p 2.6110E-03 x p
Trimethoprim 3.2877E-05 x p 6.7123E-05 x p 1.0137E-04 x p
Erythromycin 2.4384E-04 x p 3.5890E-04 x p 4.7397E-04 x p
Sulfamethoxazole 1.3425E-05 x p 4.9178E-05 x p 8.4932E-05 x p
Metoprolol 2.0548E-06 x p 1.8836E-05 x p 3.5616E-05 x p
Gemfibrozil 8.2192E-06 x p 3.9726E-05 x p 7.1233E-05 x p
Bezafibrate 1.4247E-05 x p 8.1096E-05 x p 1.4795E-04 x p
Carbamazepine 4.1096E-05 x p 2.9589E-04 x p 5.5068E-04 x p
Oestrone 1.3699E-05 x p 2.0548E-05 x p 2.7397E-05 x p
17β-oestradiol 1.6438E-06 x p 9.0411E-06 x p 1.6438E-05 x p
Triclosan 1.8082E-04 x p 2.5205E-04 x p 3.2329E-04 x p
Caffeine 1.3096E-03 x p 1.8822E-03 x p 2.4548E-03 x p
p – population served by the WWTP
Table 39 - Estimated ranges for the removal of the target PPCP compounds in WWTPs
Compound i Removal rate range Rrange,i (%)
Rmin Rmean Rmax
Acetaminophen 98.99 99.49 99.99
Diclofenac 0 40 80
Ibuprofen 88 93.5 99
Trimethoprim 10 42.5 75
Erythromycin 0 12.5 25
Sulfamethoxazole 45 62.5 80
Metoprolol 10 22.5 35
Gemfibrozil -5 30 65
Bezafibrate 35 55 75
Carbamazepine -50 -5 40
Oestrone -100 -25 50
17β-oestradiol 65 80 95
Triclosan 82 89.5 97
Caffeine 99 99.45 99.9
270
11.2. Chapter 5 supplementary information
Table 40 - Operating data for GAC, NF and SPF considered in the study
Granular activated carbon (Bonton
et al. 2012)
Nanofiltration (Bonton et
al. 2012)
Solar photo-Fenton (Ortiz
2006)
Type Modelled plant Full-scale plant Industrial-scale plant
Location Canada Canada Spain Influent flow 2,000 m3/d 2,000 m3/d 6.8 m3/d
Inhabitants 3,140 3,140
Specifications
Activated carbon density: 500 kg m3
Carbon usage rate: 0.076 kg/m3 Current servicing time: 91 days
No regenerations
Pre-coagulation with Alum Empty bed contact time: 20 min
Filtration rate: 4.5 m/h
Thin polyamide membrane
Porous size: 0.2 μm Pressure applied: 620 kPa
Spiral-wound modules: 90
Number of modules: 270
Parabolic concentrators
UV radiation ~ 30 W/ m2 Hydraulic retention time > 15
min Panel surface area: 4.16 m2
Borosilicate tubes
Tubes internal diameter: 0.05 m. Batch mode
Influent characteristics
pH 6.90 6.90
Total organic carbon (mg/L)
9.70 9.70 20
Dissolved organic
carbon (mg/L) 9.20 9.20
Alkalinity
(mg CaCO3)/L 6.50 6.50
Temperature (°C) 7.7 7.7 30
Effluent characteristics
pH 7.5 7.5 3.0
Total organic
carbon (mg/L) 0.90 0.90
Dissolved organic
carbon (mg/L) 0.90 0.90
Alkalinity (mg CaCO3)/L
40.0 40.0
Temperature (°C) 8.0 8.0 30
Table 41 - Spiral wound modules inventory modules (Bonton et al. 2012)
Spiral wound modules Amount
(kg/1,000 m3)
Polyester resin, unsaturated, at plant 0.14
N,N-dimethylformamide at plant 0.12 Polyphenylene sulfide, at plant 0.0014
Polyvinyl chloride, at regional storage – permeate tube 0.05
Epoxy resin, liquid, at plant 0.03 Isopropanol, at plant 0.017
271
Table 42 - Freshwater ecotoxicity potential of effluents discharged to freshwaters estimated according to the USEtox methodology
Compound
Freshwater ecotoxicity potential (CTUe/1,000 m3)
Effluent before advanced treatment
Effluent after advanced treatment
Granular
activated carbon Nanofiltration
Solar
photo-Fenton Ozonation
Min Mean Max Min Mean Max Min Mean Max Min Mean Max Min Mean Max
Diclofenac 0 1.308 2.59 0 0.0433 0.0857 0 0.6822 1.3505 0 0.6887 1.363 0 0.1308 0.259
Ibuprofen 0.0021 0.0773 0.1526 0 0.0008 0.0015 0.0004 0.0135 0.0265 0.0005 0.0179 0.0352 0.0012 0.0445 0.0877
Trimethoprim 0.0095 0.0569 0.0995 0.0011 0.0064 0.0113 0.0048 0.0286 0.0501 0.0029 0.0176 0.0308 0.0005 0.0028 0.005 Erythromycin 10.71 19.17 27.64 0.1071 0.1917 0.2764 1.449 2.595 3.741 0.1071 0.1917 0.2764 0.1071 0.1917 0.2764
Sulfamethoxazole 0.0299 0.1794 0.3289 0.0034 0.0203 0.0372 0.0112 0.0671 0.1229 0.0099 0.0593 0.1087 0.003 0.0179 0.0329
Carbamazepine 0.0512 0.845 1.648 0.0005 0.0085 0.0165 0.0157 0.2588 0.5045 0.0306 0.5042 0.9828 0.0005 0.0085 0.0165 Oestrone 0.428 1.498 2.782 0.0043 0.015 0.0278 0.1869 0.6542 1.215 0.107 0.3745 0.6955 0.0856 0.2996 0.5564
17β-oestradiol 0 1840 1840 0 18.4 18.4 0 816.5 816.5 0 460 460 0 368 368
Triclosan 1.06 7.42 14.84 0.0597 0.4176 0.8352 0.5309 3.717 7.433 0.2808 1.966 3.931 0.5085 3.559 7.119
TOTAL 12.29 1871.00 1890.00 0.18 19.10 19.69 2.20 824.50 830.90 0.54 463.80 467.40 0.71 372.30 376.40
Table 43 - Freshwater ecotoxicity potential of effluents discharged to agricultural soils estimated according to the USEtox methodology
Compound
Freshwater ecotoxicity potential (CTUe/1,000 m3)
Effluent before advanced treatment
Effluent after advanced treatment
Granular
activated carbon Nanofiltration
Solar
photo-Fenton Ozonation
Min Mean Max Min Mean Max Min Mean Max Min Mean Max Min Mean Max
Diclofenac 0.0000 0.0515 0.1019 0.0000 0.0017 0.0034 0.0000 0.0268 0.0531 0.0000 0.0271 0.0536 0.0000 0.0051 0.0102
Ibuprofen 0.0000 0.0014 0.0027 0.0000 0.0000 0.0000 0.0000 0.0002 0.0005 0.0000 0.0003 0.0006 0.0000 0.0008 0.0015
Trimethoprim 0.0004 0.0023 0.0040 0.0000 0.0003 0.0005 0.0002 0.0012 0.0020 0.0001 0.0007 0.0012 0.0000 0.0001 0.0002 Erythromycin 1.3420 2.402 3.463 0.0134 0.0240 0.0346 0.1816 0.3252 0.4688 0.0134 0.0240 0.0346 0.0134 0.0240 0.0346
Sulfamethoxazole 0.0020 0.0117 0.0215 0.0002 0.0013 0.0024 0.0007 0.0044 0.0080 0.0006 0.0039 0.0071 0.0002 0.0012 0.0021 Carbamazepine 0.0008 0.0124 0.0241 0.0000 0.0001 0.0002 0.0002 0.0038 0.0074 0.0004 0.0074 0.0144 0.0000 0.0001 0.0002
Oestrone 0.0004 0.0014 0.0025 0.0000 0.0000 0.0000 0.0002 0.0006 0.0011 0.0001 0.0003 0.0006 0.0001 0.0003 0.0005
17β-oestradiol 0.0000 2.5500 2.5500 0.0000 0.0255 0.0255 0.0000 1.132 1.1320 0.0000 0.6375 0.6375 0.0000 0.5100 0.5100 Triclosan 0.0020 0.0140 0.0280 0.0001 0.0008 0.0016 0.0010 0.0070 0.0140 0.0005 0.0037 0.0074 0.0010 0.0067 0.0134
TOTAL 1.347 5.047 6.198 0.0138 0.0537 0.0683 0.1839 1.501 1.686 0.0153 0.7049 0.7571 0.0147 0.5483 0.5729
272
11.3. Chapter 7 supplementary information
Figure 75 - The exchange rate of British Pounds (£) in the period 2006 - 2015 against the US dollar (US$)
and Euro (€), taking into account the inflation in the UK in the same period
1.00
1.05
1.10
1.15
1.20
1.25
1.30
1.35
1.40
1.45
1.50
0.8
1.0
1.2
1.4
1.6
1.8
2.0
2005 2007 2009 2011 2013 2015
Dollars (US$)
Euros (€)
UK inflation depreciation
Exch
an
ge
rati
o t
o B
riti
sh P
ou
nd
(£)
Year
Dep
reci
ati
on
of
Bri
tish
Pou
nd