THESIS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY
Biological Pre-filtration and Surface Water Treatment
Microbial barrier function and removal of
natural inorganic and organic compounds
GERALD HEINICKE
Department of Water Environment Transport
CHALMERS UNIVERSITY OF TECHNOLOGY
Göteborg, Sweden 2005
Biological Pre-filtration and Surface Water Treatment
- Microbial barrier function and removal of natural inorganic and organic
compounds
GERALD HEINICKE
ISBN 91-7291-561-7
© GERALD HEINICKE, 2005
Doktorsavhandlingar vid Chalmers tekniska högskola
Ny serie nr 2243
ISSN 0346-718x
Doktorsavhandlingar och licentiatuppsatser på Vatten Miljö Transport, nr 11
ISSN 1650-4143
Department of Water Environment Transport
Chalmers University of Technology
SE-412 96 Göteborg
Sweden
Telephone +46(0)31 772 1000
Web: www.wet.chalmers.se
Cover:
Close-up of nanofilter membrane pieces after destructive analysis. From
the left: Run with biological pre-filtration, clean membrane, run with
rapid filtration as pre-treatment.
Chalmers reproservice
Göteborg, Sweden 2005
Of beasts in Hamburg’s waterpipes
There can be found some sixteen types:
The lamprey, eel and stickleback,
Of worms-three kinds-there is no lack,
Mussels three, slow snails the same
With jolly woodlice frisk and game,
A sponge, some algae and a polyp,
Through the sieve they jump and frolic.
As corpses in the pipes are found
The mouse, the cat, also the hound;
Unfortunately lacking yetThe engineer and architect!
Rhyme in a local newspaper on the results of a biological inventory of
Hamburg’s water pipes in the late 19th century. Quoted after (Evans, 1987).
Biological Pre-filtration and Surface Water Treatment
- Microbial barrier function and removal of natural inorganic and organic
compounds
GERALD HEINICKE
Water Environment Transport
Chalmers University of Technology
ABSTRACT
Waterworks in Sweden that apply conventional chemical surface water treatment
are facing a number of challenges, including changes in raw water quality and
demands for improved particle removal. The chief objective of this work was to
evaluate biological pre-filtration for typical Swedish surface water, regarding the
removal of natural organic matter, particles, iron and manganese, and taste and
odour compounds. Pilot-scale experimental work included the investigation of
biofilters fed directly with surface water, using non-adsorptive expanded clay and
partly exhausted granular activated carbon. Process combinations with
conventional chemical treatment and nanofiltration were investigated.
Biological pre-filtration decreased the load of particles and biodegradable
organic matter to subsequent treatment processes. Peak loads of added particles
were equalised by high initial retention followed by a slow release of attached
particles. In chemical treatment with pre-filtration, the removal of µm-size
particles became less dependent on the post-sedimentation rapid filters. The
experimental study contributed with data on the microbial barrier function of
chemical surface water treatment under Swedish conditions with regard to
particle-, bacteria- and virus removal.
Simple rapid media filtration pre-treatment of surface water caused fast pressure
drop development in nanofilter membranes. Biofiltration moderated the increase
in pressure drop in comparison to rapid filtration. Destructive analysis of the
nanofiltration elements was performed to study the fouling layer on the
membrane.
Biological pre-filtration alleviated occasional episodes of dissolved manganese
and odour compounds in humic surface waters, which are difficult to control by
conventional chemical treatment alone. The mechanism of biological removal of
two biogenic odour compounds (geosmin and MIB) was found to completely
depend on metabolic activity on a non-adsorptive filter medium (expanded clay).
Within the Sustainable Urban Water Management program, a systems analysis
study was conducted. Hypothetical decentralised systems with drinking water
treatment closer to the consumer were compared to conventional centralised
treatment regarding energy consumption and microbial risk.
Keywords: biofiltration, chemical treatment, drinking water, iron and
manganese, nanofiltration, NOM, particles, taste and odour, process
combination, systems analysis
iv
Biologisk förfiltrering och behandling av ytvatten.
- Funktion som mikrobiologisk barriär och avskiljning av naturliga oorganiska
och organiska föreningar.
GERALD HEINICKE
Vatten Miljö Transport
Chalmers tekniska högskola
SAMMANFATTNING
Förändringar i råvattenkvalitet och ökande krav på behandlingseffektivitet utgör
idag utmaningar för svenska ytvattenverk. I detta arbete har alternativ till
konventionell dricksvattenbehandling utvärderats. Huvudsyftet har varit att
undersöka biologisk förfiltrering av ett typiskt svenskt ytvatten med avseende på
naturligt organiskt material (NOM), partiklar, järn och mangan, samt luktämnen.
Det experimentella arbetet har omfattat undersökningar av biofiltrering av
obehandlat ytvatten med expanderad lera (EC) och använt aktivt kol (GAC) som
filtermaterial. Biofiltreringen har kombinerats med konventionell kemisk fällning
och nanofiltrering och biofiltreringens effekter på de efterföljande processerna
har undersökts.
Den biologiska förfiltreringen minskade belastningen av NOM och partiklar för
de efterföljande processerna. Stötvis höga halter av tillsatta partiklar utjämnades,
då de fastlades i biofilterna och frisläpptes i låga halter under lång tid. I den
kemiska fällningen blev avskiljningen av råvattenpartiklar i µm-storlek mindre
beroende på den kemiska fällningens efterföljande filtreringssteg. Sammantaget
gav de utförda undersökningarna av biofiltreringens barriärfunktion information
om partikel- bakterie- och virusavskiljning under svenska förhållanden.
Snabbfiltrering som enda förbehandling för nanofiltrering resulterade i en snabb
ökning av tryckförlust över membranen. Jämfört med snabbfiltrering, minskade
biofiltrering som förbehandling tryckförlusten över membranen avsevärt. En
membranautopsi utfördes för att klarlägga foulingmekanismerna.
Biologisk förbehandling avskilde effektivt de tidvis höga halterna av luktämnen
och löst mangan som förekommer i humusrika ytvatten och som är svåra att
hantera med bara konventionell behandling. Avskiljningen av de båda
luktämnena geosmin och MIB vid biofiltrering med ett icke-adsorberande
material (EC) var helt beroende på mikrobiologisk nedbrytning.
Som ett samarbete inom forskningsprogrammet Urban Water utfördes en
systemanalys med hänsyn till miljöpåverkan och mikrobiologiska risker.
Hypotetiska decentraliserade scenarier med membranfiltrering nära
konsumenten jämfördes med konventionell dricksvattenbehandling.
Nyckelord: biofiltrering, dricksvatten, järn och mangan, kemisk fällning, lukt och
smak, nanofiltrering, naturligt organiskt kol, partiklar, process kombination,
systemanalys
v
LIST OF APPENDED PAPERS
The following papers have been appended to this thesis and are referred to in the
text by Roman numerals.
Paper I
A systems analysis comparing drinking water systems - central
physical-chemical treatment and local membrane filtration.
Westrell, T., Bergstedt, O., Heinicke, G. and Kärrman, E. (2002).
Water Science and Technology: Water Supply, 2 (2), pp. 11-18.
Reprinted with kind permission of IWA Publishing, London.
Paper II
Biological pre-treatment for improved removal of manganese in
chemical drinking water treatment.
Heinicke, G., Persson, F.,Hedberg, T., and Hermansson, M. (2000).
In: Hahn, H., Hoffmann, H. and Ødegaard, H. (eds.), Chemical
Water and Wastewater Treatment VI (pp. 201-210). Berlin, SpringerVerlag, ISBN 3-540-67574-4.
Presented at the 9th International Gothenburg Symposium on
chemical treatment of water and wastewater, Istanbul, Turkey,
October 2-4, 2000.
Reprinted with kind permission of Springer-Verlag, Heidelberg.
Paper III
Significance of dosage point for challenge tests in coagulation
treatment.
Heinicke, G., Långmark, J., Persson, F., Hedberg, T. and Storey, M.
V. (2004). In: Hahn, H., Hoffmann, H. and Ødegaard, H. (eds.),
Chemical water and wastewater treatment VIII (pp. 191-200). London:
IWA publishing. 1-84339-068-X
Presented at the 11th International Gothenburg Symposium on
chemical treatment of water and wastewater, Orlando, USA,
November 8-10, 2004.
Reprinted with kind permission of IWA Publishing, London.
Paper IV
Biological pre-filtration of surface water for removal of BOM and
regrowth potential – an investigation of filter media.
Persson, F., Heinicke, G., Uhl, W., Hedberg, T. and Hermansson, M.
in preparation (manuscript).
Paper V
Characterisation of barrier function of biofilters for pre-treatment of
drinking water with particle counts and challenge tests.
Persson, F., Långmark, J., Heinicke, G., Hedberg, T., Tobiason, J. E.,
Stenström, T. A. and Hermansson, M. (2004). Submitted to Water
Research.
vi
Paper VI
Biological pre-filtration for conventional surface water treatment.
Heinicke, G., Persson, F., Hermansson, M. and Hedberg, T. (2005).
Submitted to Aqua.
Paper VII
Removal of geosmin and MIB by biofiltration – an investigation
discriminating between adsorption and biodegradation.
Heinicke, G., Persson, F., Hedberg, T., Hermansson, M. and Uhl, W.
(2005). Submitted to Applied Microbiology and Biotechnology.
Specification of the author’s contribution to the papers:
I
In Paper I the author undertook the literature review on membrane
filtration and water budgets (Figure 1), and contributed to the choice of
treatment scenarios and writing of the paper. E. Kärrman performed the
Material Flow Analysis (MFA) and T. Westrell the Microbial Risk
Analysis (MRA).
II
In Paper II the author performed the literature study and all data
evaluation, and in cooperation with F. Persson, writing the paper. The
initial planning of the project was performed by T. Hedberg in cooperation
with a consulting company. Most of the analytical work was carried out at
the participating waterworks and commercial laboratories.
III
In Paper III the planning and execution of the experiments was
undertaken in a group, of which the author was an equal contributor. The
author was responsible for data analysis, and in cooperation with M.
Storey, writing of the paper. J. Långmark and M. Storey performed
bacteriophage preparations and analyses. The co-authors contributed with
useful commentary and discussion.
IV
In Paper V the author was an equal contributor to the planning and
executing the experimental work. The microbial- and biofilm formation
analyses were performed by F. Persson. The author performed the DOC
and BDOC analyses and evaluation of the data. The author was an equal
contributor to writing the manuscript, which was coordinated by F.
Persson. W. Uhl contributed to the discussions.
V
In Paper V the author was equal contributor to the planning of
experimental work. The spiking of microspheres and bacteriophages was
carried out by F. Persson and J. Långmark. The author performed the
continuous analysis and evaluation of particle counts, and was an equal
contributor to writing the paper, which was coordinated by F. Persson. J.
Tobiason contributed with modelling of particle removal, while the coauthors provided useful commentary and discussion on the manuscript.
VI
In Paper VI the author planned the experiments and played a leading role
in their execution. The jar tests, Total Organic Carbon (TOC) and
Biodegradable Dissolved Organic Carbon (BDOC) analyses were
undertaken by the author. The remaining parameters were analysed at the
vii
waterworks and at a commercial research laboratory. Assimilable Organic
Carbon (AOC) was analysed by J. Långmark. The author was responsible
for data analysis, figures and preparation of the paper. The co-authors
provided useful commentary and discussion on the experimental the work
and paper.
VII
In Paper VII the author planned the experiments and played a leading role
in their execution. F. Persson was responsible for the microbial sections.
Data evaluation and writing of the paper was undertaken with the
cooperation of F. Persson. W. Uhl contributed to the discussions.
Other publications by the author
Heinicke, G., Persson, F., Hedberg, T., Ekendahl, S., Thell, A.-K. and
Hermansson, M. (2000). Experiments with biological removal of iron and
manganese at four Swedish waterworks. In: Proceedings of the 4th
International Conference on Water Supply and Water Quality, (pp. 737748). Krakow, September 11-13. ISBN: 83-911077-7-9.
Palmquist, H., Norström, A., Ahlman, S. and Heinicke, G. (2003). Tokyo löser
Va-problem med avancerad teknik (In Swedish). Svenskt Vatten (No. 4,
September), pp. 49-51.
Kärrman, E., Bergstedt, O., Westrell, T., Heinicke, G., Stenström, T. A. and
Hedberg, T. (2004). Systemanalys av dricksvattenförsörjning med
avseende på mikrobiologiska barriärer, miljöpåverkan och hushållning
med naturresurser. (In Swedish). VA-forsk report 2004-12. Swedish Water
/ Svenskt Vatten. ISBN: 91-85159-18-2.
Available from www.svensktvatten.se
Kristenson, S. E., Bergstedt, O., Heinicke, G., Persson, F. and Hedberg, T.
(2005). Mikrobiologiska barriärer i vattenrening (In Swedish): VA-forsk
report. Swedish Water / Svenskt Vatten. In preparation.
Persson, F., Heinicke, G., Långmark, J., Hedberg, T., Stenström, T. A. and
Hermansson, M. (2004). Biofilters for pre-treatment of surface water in
Nordic climate conditions - a comparative study of different filter media.
Presentation at the 2nd IWA Leading-Edge Conference on Water and
Wastewater Treatment Technologies. Prague, 1- 4 June 2004. Paper IV is
based on this material.
Persson, F., Heinicke, G., Hedberg, T., Hermansson, M. and Uhl, W. (2005).
Biological pre-treatment and nanofiltration - implications for water quality
and membrane fouling. In preparation.
viii
Presentations at national conferences and seminars
Presentation titled “Biologi och membran – en optimal processkombination?” at
the drinking water seminar organised by Va-Ingenjörerna consulting engineers,
Stockholm, 2002.
Presentation titled “Nya dricksvattenmetoder och möjligheter för deras
användning i framtida dricksvattensystem” at the Swedish water and wastewater
fair (Va-mässan), Göteborg, September 2, 2002.
Presentation titled “Microbial barriers in drinking water treatment”
(Barriärfunktion i dricksvattenbehandlingen) at the Swedish food authority’s
seminar series on safe drinking water supply (Livsmedelsverkets seminarieserie:
säkrare dricksvattenförsörjning). Kalmar, April 28, Luleå, May 5, Stockholm,
October 28, 2004.
Advised MSc. theses
Moreno, C. (2002). Biodegradable Dissolved Organic Carbon (BDOC) in raw
and biologically treated water from the pilot plant at Lackarebäck
waterworks. MSc. Thesis 2002:1, WET, Chalmers University of
Technology, Göteborg.
Li, Z. (2003). Removal of Geosmin and 2-Methylisoborneol (MIB) in drinking
water - a pilot plant study of biofiltration with GAC and LECA (Lightexpanded-clay-aggregates). MSc. Thesis 2003:6, WET, Chalmers
University of Technology, Göteborg.
Crncevic, S. (2003). Odorants removal by adsorption on granulated activated
carbons. MSc. Thesis 2003:13, WET, Chalmers University of Technology,
Göteborg.
Johansson, A. and Scott, S. (2004). Den kemiska reningens barriärverkan mot
patogena mikroorganismer i dricksvattenberedningen (In Swedish). MSc.
Thesis, WET, Chalmers University of Technology, Göteborg.
ix
ACKNOWLEDGEMENTS
This work at Water Environment Transport (WET) was financed by the research
and development fund (VA-forsk) of the Swedish Water Association (Svenskt
Vatten) and MISTRA, the Swedish foundation for strategic environmental
research. Further financial support to the project was granted by the City of
Göteborg through Göteborg water and sewage works (Va-verket). All financial
contributions are gratefully acknowledged.
People at Va-verket and Lackarebäck waterworks were indispensable partners in
the planning and execution of the studies and discussion of the results. Especially
mentioned are Olof Bergstedt and Sven-Eric Kristenson, as well as Inger
Kjellberg, Henrik Rydberg and Åke Andersson at the waterworks’ lab. Thanks
also to the rest of the lab staff for taking care of samples at short notice, and the
workshop personnel for technical support and patience with clumsy PhD
students. Peter Sehn at Dow Liquid Separations gave valuable input to the
nanofiltration study.
Discussions with the fellow students in the Urban Water program provided
inspiration and perspective on water and wastewater systems. At Urban Water
headquarters, Henriette Söderberg and program director Per-Arne Malmqvist
succeeded with the ambitious task of holding together the program with nine
universities involved. With their unorthodox course themes, including the study
trips to mega cities, they made the research school an unforgettable experience.
At WET the staff and PhD students provided a good working atmosphere. I
would like to thank current and former lab personnel, Jesper Knutsson, Mona
Zanders and Evy Axén, for their analytical and practical assistance. Guest
professors John E. Tobiason of at the University of Massachusetts and Mino
Takashi at the University of Tokyo are also gratefully acknowledged for
modelling competence and hospitality during my visit to Japan. Micheal Storey at
SMI is acknowledged for good discussions and language comments.
Most of all, I would like to thank the following people:
My supervisors Prof. emeritus Torsten Hedberg at Chalmers and Prof. Wolfgang
Uhl at Dresden University of Technology, Germany for creative guidance,
treatment process competence and spending their precious time.
Co-worker Frank Persson at Göteborg University for fruitful cooperation, sharp
thinking and unlimited dedication to the project – no matter what time of day.
My parents and Oma for their loving support, and for not asking too many
questions as to how it was going at work. We will spend more time together from
now on.
My fiancée Charlotte Corfitzen who recently presented her own PhD thesis. Now
we will find time for our private projects, such as moving together and learning
each other’s languages. Tusind tak. I’ll stick to the easy words to start with.
Göteborg, December 2004
Gerald Heinicke
x
PREFACE
This work has been carried out as part of the Swedish national research
programme Sustainable Urban Water Management financed by MISTRA, the
foundation for strategic environmental research.
The objective of the Urban Water program has been to develop support for
strategic decisions on the future sustainable systems in Sweden regarding water,
storm water and wastewater systems. Five groups of criteria have been chosen,
focusing on health and hygiene, the environment, economy, socio-culture, and
technical function. Models and assessment methods have been developed and
tested for each group of criteria.
Sixteen PhD students from nine universities have been part of the program with
particular emphasis on facilitating cooperation between the different scientific
fields; engineers, microbiologists, social scientists and economists. The
interdisciplinary character has also been emphasised by a joint research school.
Within the field of drinking water, the co-operation has been between
•
Engineers at Chalmers: Gerald Heinicke and Torsten Hedberg
•
Microbiologists at Göteborg University: Frank Persson and Malte
Hermansson
•
Microbiologists and risk assessment experts at the Swedish Institute for
Infectious Disease Control, Solna: Jonas Långmark, Therese Westell,
Michael Storey and Thor-Axel Stenström.
•
Experts on systems analysis and life cycle assessment: Erik Kärrman
•
Engineers in the water industry: Olof Bergstedt, Göteborgs Va-verk.
xi
ABBREVIATIONS
AC
Activated Carbon
AOC
Assimilable Organic Carbon
BAC
Biological Activated Carbon filtration
BDOC
Biodegradable Dissolved Organic Carbon
BOM
Biodegradable Organic Matter
DBP
Disinfection By-Products
DOC
Dissolved Organic Carbon
EBCT
Empty Bed Contact Time
EC
Expanded Clay (crushed aggregates)
EPS
Extracellular Polymeric Substances
FC
Flow Cytometry
FL
Autofluorescent particles
FNU
Formazine Nephelometric Units
GAC
Granular Activated Carbon
GC-MS
Gas Chromatography – Mass Spectrometry
Geosmin
Trans-1, 10-dimethyl-trans-9-decalol
HPC
Heterotrophic Plate Count
HRT
Hydraulic retention time
ICP-MS
Inductive Coupled Plasma – Mass Spectrometry
LC-OCD
Liquid Chromatography – Organic Carbon Detection
MF
Microfiltration
MFA
Material Flow Analysis
MIB
2-methylisoborneol
MRA
Microbial Risk Assessment
NF
Nanofiltration
NOM
Natural Organic Matter
PAC
Powdered Activated Carbon
RO
Reverse Osmosis
SUVA
Specific UV-absorbance (UV254 / DOC)
THM
Trihalomethane
TOC
Total Organic Carbon
UF
Ultrafiltration
UV254
UV-absorption at 254 nm
xii
CONTENTS
1
2
3
4
5
6
INTRODUCTION .......................................................................................1
1.1
Purpose and scope of the work ..................................................................... 2
1.2
Outline of the thesis ....................................................................................... 2
LITERATURE REVIEW...........................................................................3
2.1
System considerations.................................................................................... 3
2.1.1
Sustainability in drinking water treatment .......................................... 3
2.1.2
Decentralised urban water systems and water reuse ......................... 3
2.2
Treatment objectives...................................................................................... 4
2.2.1
NOM and BOM...................................................................................... 4
2.2.2
Barrier function and particle removal.................................................. 6
2.2.3
Biogenic taste and odour ....................................................................... 8
2.2.4
Iron and manganese ............................................................................... 8
2.3
Treatment processes....................................................................................... 9
2.3.1
Chemical treatment................................................................................ 9
2.3.2
Biofiltration........................................................................................... 12
2.3.3
Membrane filtration ............................................................................. 16
INVESTIGATIONS...................................................................................21
3.1
Lackarebäck waterworks and raw water ................................................... 21
3.2
Systems studies ............................................................................................. 22
3.3
Iron and manganese removal from surface water .................................... 23
3.4
Biofilters at Lackarebäck waterworks ....................................................... 24
3.4.1
Biofilters for carrier media study........................................................ 24
3.4.2
Biofilters for odour removal study ..................................................... 26
3.4.3
Biofilters for pre-treatment study....................................................... 27
3.5
Process combinations ................................................................................... 28
3.5.1
Chemical treatment pilot plant ........................................................... 28
3.5.2
Nanofiltration........................................................................................ 30
RESULTS AND DISCUSSION...............................................................33
4.1
System studies ............................................................................................... 33
4.2
Iron and manganese removal...................................................................... 33
4.3
Geosmin and MIB removal......................................................................... 34
4.4
Evaluation of carrier materials ................................................................... 35
4.4.1
NOM and growth potential ................................................................. 36
4.4.2
Barrier function .................................................................................... 37
4.5
Biological pre-filtration and chemical treatment...................................... 37
4.5.1
NOM ...................................................................................................... 38
4.5.2
Barrier function .................................................................................... 39
4.6
Biological pre-filtration and nanofiltration ............................................... 41
4.6.1
Permeate quality................................................................................... 41
4.6.2
Fouling ................................................................................................... 43
CONCLUSIONS AND FURTHER WORK..........................................47
REFERENCES ...........................................................................................51
APPENDED PAPERS: I-VII
xiii
INTRODUCTION
1
INTRODUCTION
One of the prerequisites for an adequate drinking water supply is a sufficient
access to raw water resources of reasonable quality. In most cases this is achieved
in Scandinavia. In Sweden for example, approximately 50% of drinking water
originates from surface waters, 25% from artificial infiltration and 25% from
ground water (Swedish Water, 1996).
Swedish waterworks, typically designed thirty or more years ago however face a
number of challenges in the conventional treatment of surface water. These
challenges include more stringent regulations, demands for sustainable
production, the discovery of new chemical and microbial threats, changes in raw
water quality, decreased water consumption causing long retention times in the
distribution network, and high demands from the consumer regarding the
aesthetic quality of tap water.
New threats include the prevalence of difficult-to-remove pathogenic
microorganisms in most surface waters, which have caused outbreaks of
waterborne disease despite functional conventional water treatment. Examples of
recently-discovered chemical pollutants of interest are of both biogenic, e.g.
endotoxins (Anderson et al., 2002) and anthropogenic origin, e.g. drug residuals.
Seasonal problems with the precipitation of dissolved iron and manganese in the
distribution network and episodes of earthy-musty odour are common causes of
consumer complaints (Manwaring et al., 1986). A perceived microbial risk (Stein,
2000) as well as impaired aesthetic quality may seriously undermine the
consumer’s trust in tap water. Since bottled water is by orders of magnitude more
expensive and energy-intensive, maintenance of high quality municipal drinking
water is in fact a sustainability issue.
Since the early 1990s, changes in surface water quality were observed in parts of
Central and Northern Europe, primarily as increased colour caused by humic
substances. Conventional chemical waterworks had to react by increasing
chemical dosages (Nordtest, 2003), however this was not always sufficient to
maintain desired drinking water quality. There are examples of facilities having
to reconsider their treatment options in favour of more powerful processes.
Many Swedish waterworks are in need of renovation and many others require an
upgrading of their treatment schemes to ensure high quality tap water.
Furthermore, treatment alternatives need to be reconsidered now that membrane
filtration has become an economically feasible option.
Traditional slow sand filtration has enjoyed a reputation of producing water of
good aesthetic quality. Similarly, riverbank filtration and artificial ground water
recharge, representing processes with long hydraulic retention times and a
predominantly biological function, are known to produce feed water of high and
stable quality (Kuehn and Mueller, 2000). Despite this, minimal work has been
carried out on biological filters with higher filtration rates operated directly on
1
INTRODUCTION
surface water. The work undertaken in this thesis investigates the application of
this process to the pre-treatment of water prior to conventional separation
processes or membrane filtration.
1.1
Purpose and scope of the work
The objective of this project was to evaluate the efficacy of biological filtration in
the pre-filtration of water prior to conventional chemical treatment and
membrane filtration.
The aspects of biological pre-filtration investigated in this work were:
a)
Removal of Biodegradable Organic Matter (BOM) that may otherwise
cause downstream biological regrowth in the distribution system or
biofouling on a membrane.
b)
Removal of suspended particles and the role that processes involved in this
may play as an additional barrier to microbial pathogens.
c)
Removal of substances that may compromise the aesthetic (taste, colour
and odour) quality of water.
d)
The equalisation of variations in incoming raw water quality, limiting peak
loads to the following process.
e)
The effect on subsequent processes, chemical treatment or membrane
filtration.
During the course of this thesis, the question was addressed what the concept of
sustainability implies for drinking water supply under Swedish conditions.
Furthermore, the pilot plant constructed for the above stated studies was used to
increase knowledge on pathogen removal by conventional chemical treatment
and nanofiltration.
1.2
Outline of the thesis
The thesis is structured as follows: Chapter 2 summarises the project-related
literature. Where appropriate, the relevance of specific issues for the project is
pointed out. Chapter 3 briefly describes the experimental set-up applied in this
work. The results are summarised in Chapter 4. Project conclusions and further
research needs are outlined in Chapter 5.
2
LITERATUR REVIEW
2 LITERATURE REVIEW
2.1
System considerations
2.1.1 Sustainability in drinking water treatment
Investigations into the sustainability of urban water systems have traditionally
been focused on energy consumption and the usage of renewable and nonrenewable resources. In Europe, primary energy consumption is around 4700 W
per person, continuously. According to Imboden (2000), a sustainable level of
energy consumption should be around 2000 W per person. Wallén (1999)
conducted a life cycle analysis for the production of drinking water at
Lackarebäck waterworks in Göteborg. The dominating environmental burden
was attributed to energy consumption, which is in agreement with other studies
(Friedrich, 2002). Based on data from Wallén (1999), the energy used in the
production and supply of 1 m3 of cold drinking water in Göteborg is about 1.8
MJ, including production of chemicals, road transport, and maintenance of the
pipe network. Given a total production of 345 l per person and day (Göteborg
Water and Sewage Works, 2004), this equates to a continuous consumption of 7
W per person. Sixty percent of the energy consumption are due to the pumping of
water and therefore remains unaffected by changes in the treatment process. The
current means of supplying water to industry and the consumer would make up
less than 1% of a sustainable society’s primary energy consumption.
It may therefore be concluded that this must be acceptable for the provision of
clean drinking water to the consumer’s home. The disposal of waste from water
treatment may pose problems locally; however these problems are negligible
when compared to resource questions in other sectors of society.
2.1.2 Decentralised urban water systems and water reuse
Conventional urban water systems with centralised treatment of drinking-, stormand wastewater have been criticised for wasting resources and causing pollution.
One such example is the purification of large volumes of water to drinking
standard, of which only a small fraction is used for portable purposes. More
flexible technical solutions have been advocated that allow the local treatment of
water to the level of purity needed for a given purpose, as well as water reuse
(Weber, 2002; Wilderer, 2004).
Under conditions of water scarcity, systems that supply more than one quality of
water to consumers have evolved. This idea however is not new as such, with
researchers as early as Loll (1892) having described an in-house installation
implemented in St. Petersburg that included the reuse of grey water for toilet
flushing as part of a source-separating toilet system. Non-potable reuse where
treated wastewater is used for example in toilet flushing have been implemented
in Japan (Asano, 1996; Maeda et al., 1996; Ogoshi et al., 2000), Australia
(Anderson, 1996; Law, 1996) and the US (Okun, 1997; Okun, 2000; Thompson,
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2000). Based on projections of cost and microbial risk, Fane (2002) suggested that
a size of around 1000 consumers be preferable for non-potable wastewater reuse
systems.
Rainwater harvesting is another option to decrease the consumption of potable
water. Common in many developing countries, it is also considered for use in
cities of industrialised nations (e.g. Herrmann and Hasse, 1997; König, 2000).
However, in moderate climatic zones without pronounced water shortages, there
may be neither economical nor environmental benefits of large-scale rainwater
harvesting (Crettaz et al., 1999; Mikkelsen et al., 1999). Furthermore, due to
possible contamination with pathogens, the introduction of rainwater into the
household represents a hygienic risk (Albrechtsen, 2002).
Decentralised treatment is an option to limit the amount of water purified to
drinking water quality. There are examples of water supply systems where only
basic quality water designated for household purposes is supplied centrally, while
a part of the flow is upgraded to drinking water quality by membrane filtration
plants located in several districts of a city (Ma et al., 1998). Particularly in the US,
point of use devices have a tradition of application. Such appliances act as an
additional barrier against pathogens originating from the raw water or the
distribution system (Payment, 1998), although the issue of regrowth in the inhouse installation has been raised (Payment et al., 1991). The future of point of
use treatment and decentralised systems is subject to discussion. While expected
to spread in the US, the current development in the European Union is more
focused on holistic, catchment-to-tap improvement of centralised water supply
systems (McCann, 2004).
2.2
Treatment objectives
2.2.1 NOM and BOM
The prevalence of Natural Organic Matter (NOM) in treated drinking water may
cause both aesthetic problems of colour, taste and odour compounds, microbial
regrowth, as well as health-related problems through the formation of
disinfection by–products (DBP) such as the trihalomethanes (THM).
Bulk NOM is commonly measured as Total Organic Carbon (TOC). Humic
substances, high molecular weight products formed through the decomposition of
biota, are quantified by a water sample’s absorbance of ultraviolet radiation
(Eaton, 1995). UV absorbance is commonly measured at a wavelength of 254 nm
(UV254), which excites saturated C-C double bonds. Similarly, colour is often
used as a general indicator of the concentration of humic substances.
Free chlorine reacts with NOM, forming halogenated organic substances or DBPs
(e.g. Rook, 1977), some of which have been identified as carcinogens and
mutagens. Waterworks therefore remove a large portion of the NOM fraction
prior to chlorination.
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The fraction of NOM that can be degraded by bacteria is assigned as
Biodegradable Organic Matter (BOM). Heterotrophic microorganisms utilise
BOM for life (dissimilation) and growth (assimilation). High concentrations of
BOM can therefore give rise to bacterial regrowth in the distribution system
resulting in high planktonic bacterial numbers and biofilm formation on surfaces.
High bacterial numbers may lead to aesthetic problems in the form of
discolouration of water, bad odours and taste (e.g. O'Connor et al., 1975;
Mallevialle and Suffet, 1987), and to hygienic problems as pathogens can be
retained and survive in thick biofilms (LeChevallier et al., 1987; LeChevallier et
al., 1988; Herson et al., 1991; Camper et al., 1996; Fass et al., 1996).
Bulk parameters such as TOC and UV-absorption are used as rough measures of
organic carbon and do not give reliable information on the biological stability of
the water, i.e. the ability of the water to support microbial regrowth.
Biodegradable dissolved organic carbon (BDOC) and Assimilable Organic
Carbon (AOC) are measures of biological stability and are measured as
bioassays.
Varieties of the two bioassays have been proposed, though in principal BDOC is
quantified by the difference in DOC before and after incubation of a water
sample with autochthonous (indigenous) microorganisms (Joret and Lévi, 1986;
Servais et al., 1987; Ribas et al., 1991; Frias et al., 1992; Volk et al., 1994; Allgeier
et al., 1996; Sondergaard and Worm, 2001). AOC is quantified by the maximum
numbers of specific strains of test bacteria supported by the water, expressed as
µg C/l of a reference compound (Kemmy et al., 1989; Stanfield and Jago, 1989;
LeChevallier et al., 1993; van der Kooij and Veenendaal, 1995). Both AOC and
BOC are time-consuming bioassays that require several days to weeks to obtain a
result.
AOC and BDOC measure different fractions of NOM since no consistent
correlations could be established between the two. (Jago and Sidorowicz, 1994;
Charnock and Kjonno, 2000). While AOC is expected to comprise the rapidly
biodegradable compounds dominated by low molecular weight acids, BDOC may
contain more slowly utilisable fractions of NOM. AOC in Norwegian surface
water was associated with compounds < 1000 atomic mass units (amu) (Hem and
Efraimsen, 2001). BDOC from natural surface water in France was shown to be
dominated by low molecular weight compounds, but also contained 25% of
humic substances (Agbekodo and Legube, 1995). To assess BOM
comprehensively, it has been recommended to measure both parameters
simultaneously (Escobar and Randall, 2001a). For the assessment of bacterial
regrowth AOC is considered more suitable, whether BDOC has been
recommended to determine the chlorine demand (Huck, 1990; Kaplan et al.,
1994). In raw water from the same surface water source, the proportion of BOM
can vary seasonally. Often the proportion of the smaller, more biodegradable
fractions is higher in the warmer seasons (Klevens et al., 1996). Indicative BOM
limits for biostable water, which does not cause regrowth in the distribution
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system are an AOC concentration below 10 µg acetate-C/l (van der Kooij, 1992)
and a BDOC concentration below 0.15 mg/l (Servais et al., 1995).
2.2.1.1 Climate-induced changes in surface water quality
In the past decades, changes in surface water quality have been reported, for
example in Sweden (Abrahamsson, 2002; Hernebring, 2003; Jabur et al., 2003;
Johansson, 2003; Tilja, 2003), Norway (Eikebrokk, 2002; Liltved, 2002; Liltved
and Gjessing, 2003), and the UK (Freeman et al., 2001). The reported changes are
primarily higher average and maximum concentrations of NOM and increased
short-term variability. In addition, a superproportional increase in colour has
been observed (Nordtest, 2003). This would correspond to an elevated specific
UV absorbance (SUVA = UV254/DOC).
Humus leakage from upstream catchments into surface waters is mainly governed
by climatic factors, and is increasing with rainfall and temperature (Nordtest,
2003). High temperatures for example speed up the formation of DOC in the soil.
In Nordic catchments, which are usually characterised by a thin soil layer on
bedrock, the influent to surface waters normally occurs through the replacement
of humic-enriched pore water in the soil (Grip and Rodhe, 1994). High
precipitation particularly during the autumn causes transport of NOM into
surface waters (Hongve, 1999). Models of climatic change for the Nordic
countries generally predict higher temperatures and runoff (Rummukainen et al.,
2003), which would cause increased transport of DOC into surface waters
(Forsberg, 1992). Long-term variations of NOM in surface water of similar
magnitude to the ones observed during the past 15 years however have also
occurred during the past 100 years, and climate is known to vary through natural
cyclic processes. Predictions of future colour levels in surface waters therefore
remain uncertain (Liltved, 2002; Löfgren et al., 2003).
2.2.2 Barrier function and particle removal.
With the application of lower doses of chemical disinfectants, the importance of
physical removal of pathogens by water treatment has increased. This has been
further emphasised after waterborne outbreaks caused by chlorine-resistant
protozoa such as cryptosporidium and giardia (e.g. Miller, 1994; Fox and Lytle,
1996).
Pathogenic microorganisms causing waterborne outbreaks of disease are enteric
viruses, bacteria and parasitic protozoa. Waterborne pathogens are diverse and
normally present in raw and treated waters at very low concentrations. Given that
many analytical methods used to identify specific pathogens are time- and labourintensive, surrogate organisms are used for the routine monitoring of drinking
water quality. Swedish regulations demand the absence of bacteria that indicate
faecal contamination (Swedish National Food Administration, 2001). A number
of processes including chemical treatment, slow sand filtration and membrane
filtration, as well as disinfection are regarded as barriers against pathogenic
microorganisms (microbial barriers). Depending on the quality of the surface
6
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water, a minimum of two to three barriers is generally required. In contrast to for
example US regulations, there is no quantitative requirement regarding the
barrier efficacy of the applied processes in Sweden
For physical removal, pathogens may be simply regarded as particles.
Measurement of turbidity has long been the most common parameter for
continuous monitoring of particle removal in water treatment. In recent years,
particle counting has proven to be a more sensitive tool in waters with low
turbidity, and furthermore allows for the monitoring of relevant particle sizes in
the µm range (Ribas et al., 2000; Bridgeman et al., 2002b). Particle counters are
commonly calibrated with artificial microspheres and tend to undersize natural
particles that differ in refractive index (Bridgeman et al., 2002a).
A shortcoming of the use of particle counting to calculate removals in a chemical
treatment train is that additional particles formed in the process cannot be
distinguished from the ones derived from raw water. Naturally occurring
autofluorescent algae can be used to overcome this problem and have been used
to assess barrier functions (Bergstedt et al., 2000; Akiba et al., 2002; Bergstedt and
Rydberg, 2002). The advantage of using autofluorescent algae is that they occur
in surface waters in quantifiable numbers, they are in the same size ranges as
bacterial pathogens and parasitic protozoa, are not formed during water
treatment and can be rapidly enumerated by flow cytometry (FC).
To investigate virus removal in water treatment, bacteriophages (viruses that
infect bacteria) are preferred to human enteric viruses for reasons of safety and
ease of enumeration. Only in polluted surface waters, may their numbers be high
enough to be followed over treatment processes (e.g. Jofre et al., 1995). To allow
for quantitative assessment of the barrier function against pathogens that occur in
low numbers, challenge tests in which pathogens or surrogate particles are added
into feed waters have been conducted for research purposes.
Questions remains however, on the reliability of available quantification
methods, particularly during challenge tests when high numbers of a surrogate
particle are added to the water. Indicator bacteria and bacteriophages are
commonly enumerated by plate-count methods, in which each particle containing
one or more colony– or plague forming units results in one count. In addition to
actual removal, decreasing numbers of organisms may be caused by inactivation
and aggregation. In a study by Gale (1997), a comparison of bacterial spore
counts before and after alum coagulation and rapid gravity filtration
demonstrated that whilst drinking water treatment removed 95 to 99% of spores,
it also promoted their clustering. Through microbial clustering, consumers may
be exposed to higher doses of pathogens than they would normally encounter.
Grant (1994) modelled virus aggregation in the aquatic environment and
concluded that the prevalence of virus clustering was not probable. This may
however not hold true for the conditions encountered in chemical drinking water
treatment, where coagulation is purposely induced. With a novel method of virus
detection, Gitis (2002) showed that viruses were aggregated in the stock solution
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may disaggregate when added to the water, thereby creating the opposite effect
i.e. that removal may in fact may be underestimated by plaque counts.
The risk of waterborne disease from tap water in Sweden has been addressed by
regulations demanding a sufficient number of microbial barriers in water
treatment (Swedish National Food Administration, 2001). Recent systematic risk
assessment studies suggest that even a well-functioning conventional surface
water treatment train may result in a considerable number of waterborne
infections spread over time and space (Westrell et al., 2003). This finding is in
agreement with North American studies (Payment et al., 1997). Furthermore,
with the current conditions of catchment protection and water treatment in
Göteborg, Sweden, the theoretical risk for cryptosporidiosis caused by cattle
grazing close to the shore of a river used as a raw water source was considered to
be non-negligible (Rosén and Friberg, 2003). In addition to the fact that more
detailed data on treatment performance was needed, each of these studies
indicated that an improvement in the barrier function in surface water treatment
was advisable.
The quality of surface water varies to a certain site-specific degree. Furthermore,
the efficacy of treatment processes is variable, and may be affected by various
parameters, as well as incidents at the facilities. Routine online process
monitoring of particle removal efficacy is so far limited to turbidity and µm-size
particle counting (Carr et al., 2003). For specific pathogens and sub-µm particles
the dynamic function of treatment processes therefore remains unknown.
2.2.3 Biogenic taste and odour
Though seldom an indication of human health risk, unpleasant taste and odour in
treated drinking water may undermine consumer trust (Manwaring et al., 1986;
McGuire, 1995). In surface waters, seasonal odour episodes are a widespread
problem (Wnorowski, 1992; Suffet et al., 1996; Bruchet, 1999). Dissolved
substances produced by certain algae and actinomycetes cause earthy-musty
tastes and odours in surface waters. Two common odorous algae metabolites are
trans-1-10-dimethyl-trans-9-decalol (geosmin) and 2-methylisoborneol (MIB).
Varying odour threshold concentrations for these substances are reported in the
low ng/l range. Sensitive participants in odour panels were able to detect geosmin
and MIB at 1.3 ng/l and 6.3 ng/l respectively (Young et al., 1996). Geosmin (182
g/mol) and MIB (168 g/mol) are of relatively low molecular weight and of
markedly hydrophobic character with a high octanol-water coefficient (Pirbazari
et al., 1992). Concentrations of geosmin and MIB up to 1000 ng/l have been
reported in surface water sources (Yagi, 1988), while concentrations around 100
ng/l appear to be more common in raw waters that cause odour problems
(Nerenberg et al., 2000).
2.2.4 Iron and manganese
Iron and manganese in surface waters are generally found in their precipitated,
oxidised forms. In waters lacking oxygen (i.e. ground water) and at times in
8
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eutrophic lakes, both metals may occur in their soluble forms. When oxidised
during water distribution, these metal ions cause aesthetic and technical problems
such as colour, taste, odour as well as precipitations (e.g. Michalakos et al., 1997).
In technical systems, the traditional process for iron and manganese removal is
aeration followed by filtration and if necessary, stronger chemical oxidants may
be applied. NOM, particularly humic substances, forms complexes with divalent
metal ions. In this case, oxidation to the trivalent form may be effectively
inhibited even in the presence of elemental oxygen (Theis and Singer, 1974).
Judging from the amount of available literature, iron and manganese problems
occur predominantly in ground water. There are however examples of coloured
surface waters that posed iron and manganese problems (Knocke et al., 1987;
Hedberg and Wahlberg, 1998).
2.3
Treatment processes
In the following, treatment processes relevant for the investigations described in
Chapter 3 are characterised by their ability to fulfil the treatment objectives
described above.
2.3.1 Chemical treatment
Chemical treatment is the most common process for the purification of surface
water treatment in Sweden and internationally (Swedish Water, 1996; Volk and
Lechevallier, 2002). In the context of this work, conventional chemical treatment
is defined as chemical addition (coagulant, pH adjustment) and mixing,
coagulation, flocculation, and floc settling followed by rapid media filtration
through sand or Granular Activated Carbon (GAC).
2.3.1.1 NOM removal by chemical treatment
Chemical treatment is known to remove NOM by clearly differentiated
mechanisms that dominate under specific operational conditions. During charge
neutralisation, negatively charged organic molecules form insoluble complexes
with trivalent metal ions; a stoichiometric process that has been shown to take
place at low aluminium or iron doses without hydroxide floc formation. At
conventional coagulant doses, charge neutralisation is followed and superseded
by sweep coagulation, resulting in the formation of hydroxide flocs to which
organic molecules may adsorb, while colloids are entrapped during the process
(Dennett et al., 1996; Gregor et al., 1997). Non-polar, i.e. uncharged and aromatic
organic molecules with unsaturated C-C bonds are particularly prone to be
removed by sweep coagulation, and thus make up the main part of the NOM
removed by coagulation treatment (Owen et al., 1995). The fraction of
hydrophilic uncharged molecules is not specifically targeted by the above
mentioned processes and is accordingly almost resistant to removal by chemical
treatment, even at high coagulant doses (Chow et al., 2004). Among these
compounds are biopolymers such as bacterial and algal extracellular polymeric
substances (EPS), proteins and polysaccharides (Widrig et al., 1996; Chow et al.,
9
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1999). In low turbidity raw waters it is the content of NOM that determines the
coagulant dose required (Semmens and Field, 1980).
Ozone partly oxidises organic molecules, decreasing average molecular weight
while increasing both the polarity and charge density of humic substances. This
negatively affects removal by flocculation since higher negative charge density
increases coagulant demand for charge neutralisation and oxidation of the humic
structure lessens their affinity to both hydroxide flocs and activated carbon (AC)
surfaces (Owen et al., 1995; Becker and Omelia, 1996; O'Melia et al., 1999).
The removal of BOM by chemical treatment varies with the raw water
characteristics and design of the treatment train. Charnock (2000) reports NOM
and BOM removals from nine Norwegian waterworks (Table 1) employing
chemical treatment to raw waters similar to the water from Lake Delsjöarna,
Göteborg, treated in the study described in Chapter 3. In an investigation from
the US on less humic surface waters, waterworks applying conventional
treatment achieved lower removals of BDOC than the ones in Table 1. (Volk et
al., 2000). AOC is composed of low molecular weight, non-humic compounds that
are not amenable to removal by chemical treatment. Its removal by this process
was variable, and negatively affected by pre-chlorination (Volk and Lechevallier,
2002).
Table 1: Removal of NOM and BOM by nine chemical treatment plants (Charnock and
Kjonno, 2000)
Parameter
DOC (mg/l)
BDOC (mg/l)
AOC (µg/l)
Raw water
Drinking water
Removal
4.78
1.03
30.6
2.37
0.47
20.6
52%
55%
32%
2.3.1.2 Barrier function
Turbidity remains the method by which process performance is monitored at
most waterworks. Although no direct relation exists between turbidity and
pathogen numbers sufficient barrier function has been associated with turbidity
levels in finished water. Values below 0.1-0.2 FNU have been recommended
(Miller, 1994; Xagoraraki et al., 2004).
Since bacteria are commonly more reliably inactivated by chemical disinfection
than protozoan pathogens their removal by chemical treatment is less critical at
waterworks that apply i.e. chlorination. Gerba (2003) reported a reduction of the
added indicator organism E.coli over chemical treatment by a factor of 2.67-log.
Reported log-reductions of viruses by chemical treatment plants vary between 2
to 4 (Guy et al., 1977; Rao et al., 1988; Gerba et al., 2003). In batch tests, both
lower (i.e. log-reduction < 1 to 3) (Chaudhuri and Engelbrecht, 1970; Nasser et
al., 1995) and higher removals have been achieved (Chang et al., 1953). The large
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range may be due to differences in experimental conditions as well as applied
virus type and concentration.
The removal of cryptosporidium in chemical treatment has been investigated
extensively. While reported removals from full-scale plants were in the range of
1.4 to 2.5-log, substantially higher log-removals (typically 4 to 5-log) have been
achieved when oocysts were added in augmented concentrations (e.g. Dugan et
al., 2001; Emelko, 2003). Correct quantification of pathogen removal present at
ambient concentrations is often limited by non-detects in filtered water.
However, high concentrations of added particles do not represent natural
conditions and may overestimate the actual barrier function.
The efficacy of chemical treatment for particle removal is depending on a large
number of variables including the coagulant dose in relation to raw water NOM
and turbidity, temperature, flocculation, pH etc. Filter ripening in the beginning
of a filter run and breakthrough at the end are known to negatively effect particle
removal. Sub-optimal coagulation conditions, particularly caused by low
coagulant doses, have been shown to be detrimental to pathogen removal (e.g.
Dugan et al., 2001; Emelko, 2003). Huck (2002) developed a robustness index for
particle removal that takes into account both the average performance of a
treatment process and its variation in relation to a quantified treatment goal, for
example a maximum allowed particle concentration. Hurst (2004) applied the
robustness index to chemical treatment of a river water of variable quality.
Particularly sudden increases in raw water quality (TOC, turbidity) negatively
affected robustness of chemical treatment for particle removal. Within hours,
rainstorms events affected raw water turbidity as well as NOM concentration and
character. During such events, particle removal was seriously impaired. It was
suggested to control coagulant dosage based on online measurement of both raw
water turbidity and TOC.
2.3.1.3 Taste and odour
Dissolved geosmin and MIB are not removed by conventional chemical water
treatment processes to any great extent (Sävenhed et al., 1987; McGuire and
Gaston, 1988; Kim et al., 1997; Nerenberg et al., 2000; Bruce et al., 2002). Since
high concentrations of odorants may occur in algal cells, oxidation processes such
as pre-chlorination cause cell-bound substances to become soluble and further
aggravate the odour problem (Ashitani et al., 1988; Ando et al., 1992).
Due to their hydrophobic nature, trace concentrations of the odorants geosmin
and MIB are readily adsorbed onto AC. NOM however competes with geosmin
and MIB for AC adsorption sites (Herzing et al., 1977; Lalezary et al., 1986),
particularly NOM fractions of similar molecular weight and chemical character
(Newcombe et al., 1997; Newcombe et al., 2002; Hepplewhite et al., 2004).
Competition may impair odorant removal by AC by orders of magnitude (Chen
et al., 1997; Graham et al., 2000) and thereby increase treatment costs (Pirbazari
et al., 1993). The service time during which GAC adsorbers maintain satisfactory
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adsorption capacity for geosmin and MIB depends on process parameters such as
the empty bed contact time (EBCT) and feed water NOM (Gillogly et al., 1999),
though are commonly a few months to two years (Hattori, 1988; Ridal et al.,
2001). When powdered activated carbon (PAC) is applied, high doses are
commonly required to abate odour problems (Yagi et al., 1983; Ashitani et al.,
1988; Hattori, 1988; Gillogly et al., 1998; Jung et al., 2004) and may exceed what is
feasible for full-scale treatment facilities (Ando et al., 1992; Nerenberg et al.,
2000).
2.3.1.4 Iron and manganese
Since chemical treatment does not remove soluble ions, divalent iron and
manganese will only be minimally affected. For surface coloured waters, strong
oxidants such as permanganate are required in the process, which may induce
substantial cost (Knocke et al., 1987). The majority of Swedish surface
waterworks are currently not practising oxidation treatment other than low-level
chlorination for disinfection purposes (Swedish Water, 1996).
2.3.2 Biofiltration
Traditional biological filtration processes include riverbank filtration, artificial
recharge and slow sand filtration. While bank filtration denotes the withdrawal of
raw water near a riverbed, artificial recharge is the infiltration of surface water
into the ground. Artificial recharge and bank filtration have retention times of
one to several weeks and the water produced may have ground water
characteristics (Kuehn and Mueller, 2000). These processes necessitate
favourable geological conditions.
Slow sand filtration is the oldest engineered biological water treatment process.
The filters are operated at a low filtration rate (0.1-0.4 m/h) and skimmed when
the headloss exceeds the allowable height (e.g. Hendricks, 1991). Other materials
used as filter material in biofiltration include GAC, crushed expanded clay (EC),
porous minerals such as pumice and plastic biofilm carriers.
The extent to which biological processes account for the removal efficacy of
granular media filter is variable and the distinction between strictly
physicochemical filtration and biofiltration is gradual. The biological activity
depends on raw water characteristics such as temperature and BOM content,
retention time in the filter, the adsorption capacity of the filter media, and the
presence of chemical inhibitors of microbial activity.
The biodegradation of NOM in granular media filters can be substantially
increased by pre-ozonation. Ozonation oxidises the organic molecules of NOM,
decreasing average molecular weight and thereby making a larger part of the
NOM subject to biological degradation (van der Kooij et al., 1989). Ozonation
typically increases the AOC content of the water by 100 to 900% (e.g. van der
Kooij et al., 1989; Miettinen et al., 1998; Zacheus et al., 2000)
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The treatment combination of ozonation and rapid GAC filtration is commonly
termed Biologically Activated Carbon filtration (BAC). The BAC process is well
established in the industry and well investigated (e.g. Urfer et al., 1997; Camel
and Bermond, 1998; Uhl, 2000a). BAC filtration is commonly applied for both
NOM and particle removal after chemical treatment. To degrade the lowmolecular weight organics produced after ozonation, prolonged retention times
in media filters are required. Where this is not given, difficulties may arise to
achieve biostable water that does not cause regrowth in the distribution system
(van der Kooij et al., 1989; Huber and Frimmel, 1996; Escobar and Randall,
2001b). GAC rapid filters remove the most easily biodegradable NOM fraction,
hydrophilic compounds < 3000 amu (Klevens et al., 1996), but not the slower
degradable BDOC fractions (Carlson and Amy, 1997).
Ozonation-biofiltration using crushed EC aggregates and plastic carriers as filter
media has been specifically applied to bleach and degrade humic substances in
Nordic surface waters (Ødegaard, 1996; Melin and Ødegaard, 1999; Melin et al.,
2000; Melin and Odegaard, 2000).
Several advantages of GAC as a biofilm carrier material have been attributed to
the ability of GAC to adsorb substrates, nutrients and oxygen. In this way
bacteria living on its surfaces are supplied with a flux of factors required for their
maintenance and potential growth. This permits bacterial growth and
biodegradation even at low influent substrate concentrations (Dussert and
Tramposch, 1996). The variety of functional groups on the surface of AC
improved the attachment of microorganisms. Billen (1992) found a higher affinity
of bacteria to GAC than to sand, leading to better adsorption. For BOM removal
by ozonation-biofiltration, exhausted GAC was found to be superior to inert
porous materials such as pumice and sintered glass (Uhl, 2000b). NOM removal
in inert rapid filter materials increases with operation time due to colonisation by
bacteria, while it decreases in GAC filters due to exhaustion of adsorption
capacity. After the adsorption capacity of a GAC was largely exhausted, the
material performed somewhat better than inert materials, a finding that was most
pronounced at low water temperatures (Dussert 1996).
2.3.2.1 NOM and BOM removal
The performance of slow sand filters is site-specific due to varying prerequisites
in terms of climate, the character of raw water NOM and operational conditions
such as the filtration rate. Lambert (1995) reviewed removals of NOM
parameters by European and US-American slow sand filters (Table 2). The
removal of UV-absorbing NOM and DOC corresponded to each other at the
different plants. Collins (1992) found a superproportional removal of UVabsorbing substances and concluded considerable adsorption of high molecular
weight substances to the filter media. The removal of BDOC by slow sand filters
has therefore been assumed to occur by both adsorption and biodegradation.
Ribas (1995) found that DOC removal from a natural surface water on a non-
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adsorptive biofilter medium ceased entirely when microbial activity was
suppressed by a metabolic inhibitor.
Table 2: NOM removal by European and US-American slow sand filtration plants
(Lambert and Graham, 1995)
Parameter
DOC
UV254
BDOC
AOC
Removal range
Average
5-40%
5-35%
46-75%
14-40%
16%
17%
60%
26%
Low temperature negatively affected the ability of slow sand filters to degrade
BOM, which was partly caused by a lower biodegradable NOM fraction during
the cold season (Seger and Rothman, 1996; Welté and Montiel, 1996).
Consequently, pre-ozonation is an additional option to increase the overall NOM
removal in slow sand filters (Hendricks, 1991; Graham, 1999), the positive effect
of which has been found to be most pronounced for cold raw waters (Seger and
Rothman, 1996).
Also the effect of the filtration rate on NOM removal by slow sand filters is sitespecific. While most studies report no effect of increasing the filtration rate from
0.1 m/h to 0.5 m/h on the removal efficacy for bulk organic parameters (Lambert
and Graham, 1995; Rachwal, 1996), others find a deteriorated filtrate quality at
the higher hydraulic load (Haarhoff and Claesby, 1991; Heinicke, 1999).
During the long retention times typical for riverbank filtration, high removals (30
to 80%) of bulk NOM may be achieved (Kuehn and Mueller, 2000). The majority
of DOC removal has been reported to occur during the first metres of bank
filtration, while the high molecular weight, UV-absorbing fraction was reduced
the most during the underground passage (Ludwig et al., 1997). Similarly, Weiss
(2004) reported good removal of AOC and BDOC by bank filtration, but
observed no significant shift in NOM character.
2.3.2.2 Barrier function
The health-related benefits of slow sand filtration were well known prior to an
understanding of bacteriology. This was dramatically illustrated when the cholera
epidemic of 1892 hit Hamburg with its unfiltered surface water supply, causing
8000 casualties. The City of Altona, situated just downstream and applying slow
sand filtration since 1859, was practically unaffected by the outbreak (e.g. Evans,
1987). Low-filtration rate biofiltration processes including slow sand filtration
and artificial recharge are considered microbiological barriers according to
Swedish regulations (Swedish National Food Administration, 2003).
Virus removal in slow sand filters appears to be limited. Removal of added MS-2
bacteriophage was reported to be 2-log (Yahya et al., 1993), though found to
depend on the type of bacteriophage (Jin et al., 1997).
14
LITERATUR REVIEW
2.3.2.3 Taste and odour
Degradation of MIB as the sole carbon source by isolated strains of bacteria or
consortia of strains have been demonstrated (e.g. Izaguirre et al., 1988; Tanaka et
al., 1996; Lauderdale et al., 2004). Degradation of geosmin has been shown to
occur by co-metabolism by consortia of strains, while other BOM constituted the
primary substrate (Saito et al., 1999).
Low-hydraulic load filtration processes operated directly on surface water, such
as slow sand filtration (Yagi et al., 1983), trickling filters (Lundgren et al., 1988),
biofilters (Hattori, 1988), bank filtration (Chorus et al., 1992) and artificial ground
water recharge (Sävenhed et al., 1987) have been shown to achieve high removal
of substances causing earthy-musty odours. Also at shorter contact times,
biofilters have demonstrated substantial removals of both geosmin and MIB
present at high concentration (Sumitomo, 1992; Terauchi et al., 1995). The
removal rate depended on temperature and on the initial concentration as for a
first order reaction (Egashira et al., 1992). Other authors found a dependence on
the amount of biomass (Elhadi et al., 2004) and BOM degradation, as would be
expected for a secondary substrate.
Biodegradation as well as adsorption mechanisms contribute to the removal of
odorous compounds such as MIB and geosmin in media filters. Under realistic
biofilter operating conditions, the relative contribution of biodegradation on one
side, and adsorption to filter medium and organic matter on the other, remains so
far unresolved.
2.3.2.4 Iron and manganese
The activity of iron- and manganese bacteria is known to enhance the oxidation
of iron and manganese in water. On ground waters, an array of biological
treatment processes has been applied with success (e.g. Hässelbarth and
Lüdemann, 1971; Czekalla et al., 1985; Mouchet, 1992; Seppänen, 1992;
Michalakos et al., 1997). Terauchi (1995) studied biological pre-filtration directly
on eutrophic surface water. The filters efficiently removed iron and manganese
from the raw water.
Biologically mediated oxidation of iron and manganese occurs at considerably
higher rates than chemical oxidation (Katsoyiannis and Zouboulis, 2004), and
very high filtration rates have been applied in biological filters (Mouchet, 1992).
Biooxidation was found to be limited to certain pH- and redox conditions. Since
optimal conditions differ for iron and manganese, two separate biofilters have
been suggested (Mouchet, 1992).
2.3.2.5 High-rate biofilters prior to chemical treatment
In a limited number off applied studies, biofiltration has been investigated for
pre-treatment of eutrophic and high turbidity surface waters prior to chemical
treatment. Zhang (1998) found that biofiltration of a raw water polluted with
hydrocarbons shifted the zeta-potential towards the less negative, thereby
15
LITERATUR REVIEW
facilitating particle aggregation and decreasing coagulant demand. Similar
findings from biofilters for odour removal were reported by Terauchi (1995),
where the filters also removed iron, manganese and ammonia. Yeh (1993)
applied biological pre-filtration through coke for the nitrification of a polluted
river water.
Jabur (2003) tested biofiltration on humic-rich surface water. The filter removed
a large portion of turbidity and indicator bacteria and a few percent of NOM. The
question was raised how biological pre-filtration would affect a subsequent
chemical treatment train.
2.3.2.6 On the placement of biofilters
When biofiltration was first introduced as slow sand filtration the process was
operated directly on raw water. In the course of increasing pollution of raw
waters and technical development, slow sand filters were commonly replaced by
chemical treatment or used as a polishing step at the end of the process train.
Furthermore, rapid biofilters are commonly placed after chemical treatment to
combine biodegradation with the required solid-liquid separation and to limit the
organic NOM load to the GAC medium.
There may however be a rationale to locate biofiltation early in the treatment.
Bouwer (1988) suggested the placement of biofiltration before solid-liquid
separation to remove BOM and provide an additional barrier against particles
and microorganisms.
The prerequisites for biological activity in filters are advantageous early in the
treatment train. The growth of heterotrophic bacteria may be negatively affected
by aluminium residuals from chemical treatment (Huck et al., 1991), while the
accumulation of hydroxide floc has been reported to impair biofilter function
(Prévost et al., 1995). Also the lack of BOM and inorganic nutrients may limit
biological activity for the removal of secondary substrates such as traceconcentration odorants if removed by a process upstream. Hedberg (1998)
described an example of biofiltration for the biooxidation of manganese where
the process performed well prior to coagulation, though was inefficient when
placed after chemical treatment. Furthermore, biofiltration as the last separation
process may release elevated numbers of bacteria, either suspended in the water
phase or associated with carbon fines (Camper et al., 1986; Stewart et al., 1990).
2.3.3 Membrane filtration
Membrane filtration has in recent years evolved as a cost-effective alternative to
conventional separation processes. Common applications include low-pressure
membranes, microfiltration (MF) and ultrafiltration (UF) as particle barriers and
high-pressure membranes, nanofiltration (NF) and reverse osmosis (RO) to
remove colloidal and dissolved compounds. While UF is commonly designed as
hollow fibre membranes, high-pressure NF and RO membranes have retained the
traditional spiral-wound element shape. The pore size for MF is in the range of
16
LITERATUR REVIEW
100 nm while the molecular weight cut-off for tighter membranes is typically 500105 amu for UF, 100 to 500 amu for NF and below 100 amu for RO (Nicolaisen,
2003).
2.3.3.1 NOM and BOM removal
Particularly in Norway, NF has become a commonly used method for the removal
of humic substances from coloured surface waters (Thorsen, 1999; Ødegaard et
al., 2000). NF removes organic compounds larger than the molecular weight cutoff of the membrane by size exclusion, whereby almost complete removal is
achieved for high molecular weight fractions that cause colour (e.g. Siddiqui et al.,
2000). Low molecular weight compounds such as AOC may pass NF membranes
(Sibille et al., 1997; Hem and Efraimsen, 2001). Escobar (1999) found that fullscale NF with a molecular weight cut-off of 200 amu achieved high removals of
BDOC (96%) while AOC was not removed to a significant extent. AOC removal
by NF has been shown to depend on the membrane type and water chemistry
(pH, hardness) and charge repulsion between AOC compounds and the
membrane surface was identified as the factor governing AOC retention
(Escobar et al., 2000; Escobar et al., 2001).
2.3.3.2 Barrier function
The barrier function of membrane filters against pathogens is well documented.
Reviews have shown that the removal of microorganisms by a membrane mainly
depends on two main factors (Jacangelo, 1990; Madaeni, 1999): Membrane pore
size in relation to the size of the microorganisms and membrane integrity. In
bench- and pilot-scale challenge tests, high concentrations of bacteria or viruses
have been added. Typically, complete removal of the added surrogate was
achieved (e.g. Lipp et al., 1998; Otaki et al., 1998; Panglisch et al., 1998a).
Heterotrophic bacteria have been shown to be present in high pressure
membrane filter permeate. Their presence is commonly explained by a regrowth
on the clean side of the membrane.
In full-scale applications, barrier function has been reported to be lower than in
bench-scale testing (e.g. SINTEF, 1994). In hollow fibre UF, membrane integrity
has been seen to deteriorate with time due to fibre failure (Kruithof et al., 2002).
Kitis (2003) investigated the effect of pinholes and O-ring damages in highpressure membrane systems, concluding that 300 to 500 µm pinholes were
plugged by fouling material and re-opened by chemical cleaning. Only major Oring damage compromised the integrity of the membrane element.
During drinking water production in membrane filtration facilities, integrity
monitoring is regarded as essential to ensure the barrier function. Commonly
applied technologies include the measurement of pressure decay over lowpressure membranes and µm-size particle counting (Panglisch et al., 1998b;
Johnson, 2002; Farahbakhsh, 2003). One particle counter however can only
monitor a limited membrane surface at a given sensitivity, which causes
substantial cost for instrumentation at large facilities. More cost-effective
17
LITERATUR REVIEW
approaches for particle-based integrity monitoring are being developed (Carr et
al., 2003). To monitor NF and RO membranes for the removal of sub-µm
particles, methods remain limited. For NF membranes designed for high NOM
and low salt rejection, conductivity measurements are not useful. In the offline
mode the rejection of fluorescent dyes has been used to assess integrity. It has
however been pointed out that pinholes may permit virus passage through a
membrane, while the overall rejection of solutes remains high (Johnson, 2002).
Madaeni (1999) concluded that membrane filtration was a superior microbial
barrier, however not recommendable as sole treatment barrier.
2.3.3.3 Taste and odour
Low molecular weight substances such as geosmin and MIB are known to cause
earthy-musty taste and odours. As could be expected from the molecular weight
cut-off of the membrane, UF did not remove odorous compounds (Laine and
Glucina, 2001). In the same study odour prevailed in NF permeate although the
concentration of added odorants was decreased.
2.3.3.4 Membrane fouling and pre-treatment
The deposition of dissolved compounds and particulate matter on the membrane
surface that decrease permeability is termed fouling. Fouling has been
differentiated into the deposition of inorganic compounds (scaling), particulate
fouling, organic fouling and biofouling (e.g. Vrouwenvelder and van der Kooij,
2002).
Particulate fouling is caused by the deposition of suspended particles and colloids
on the membrane. Different methods exist to assess the particle fouling in
membrane filtration, such as the silt density index (SDI) and the Modified
Fouling Index (MFI) (e.g. Kremen and Tanner, 1998; Boerlage et al., 2000).
Organic fouling occurs through the adsorption of NOM on the membrane, part of
which is irreversible. Both adsorption and desorption of NOM were found to take
place on the membrane surface (Braghetta et al., 1998). Applying Liquid
Chromatography – Organic Carbon Detection (LC-OCD), Huber (1998)
calculated the mass balance for specific NOM fractions from natural water over a
RO membrane. It was primarily hydrophobic compounds and polysaccharides
that accumulated on the membrane. The deposition of natural biopolymers has
been associated with elevated pressure drops in RO systems (Gabelich et al.,
2004). Hong (1997) established that low pH and the presence of free or complexbound divalent cations increased the hydrophobicity of humic substances and the
density of the fouling layer. A critical flux existed below which no organic fouling
occurred.
Biofouling occurs through biofilm formation on the membrane caused by the
deposition and growth of bacteria, algae and fungi (Flemming et al., 1997). Rapid
flux decline has been linked to high biomass on the membrane and biologically
unstable feed water with AOC values above 80 µg acetate-C/l (Vrouwenvelder
18
LITERATUR REVIEW
and van der Kooij, 2001). Recently it was shown that the slimy layers usually
identified as biofouling are not necessarily grown on the membrane, but may also
result from the deposition of polysaccharides and bacteria from the feed (Uhl et
al., 2003). This would be organic fouling, which has implications for pre-treatment
strategies.
Different approaches exist for the pre-treatment of NF feed water. At Norwegian
facilities for colour removal, pre-treatment is generally simple consisting of µmsize particle removal by granular media rapid filtration. Alternatively, pretreatment by micro-strainer with a pore size of 15 to 50 µm is applied. To control
fouling, a low flux of approximately 12-18 l/(m2·h) is applied (Ericsson and
Tragardh, 1997; Ødegaard et al., 2000). This is considered an economical
alternative since the troublesome operation of a complex combination of pretreatment processes can be avoided.
In the US, NF membranes are commonly operated at a considerably higher flux
of around 40 l/(m2·h). It is assumed that chemical treatment consisting of
coagulation and filtration is the minimum pre-treatment of surface water in order
to control biofouling, and additional pre-treatment steps have been
recommended (Speth et al., 1998; Speth et al., 2000).
19
INVESTIGATIONS
3 INVESTIGATIONS
In the following, the investigations that were carried out within the thesis are
described concisely. For detailed aspects and analytical methods the reader is
referred to the appended papers.
The experimental studies were carried out on natural waters with inherent
variations. Wherever possible, investigations spanned over at least one year to
account for seasonal variation. The significance of treatment effects was
examined using the student’s t-test. To limit the influence of external variation as
much as possible, experiments for the comparison of treatments were executed in
a manner that allowed for collection of paired data. When the cumulative
removal of an added particle/tracer was quantified from grab samples, linear
integration was carried out between sample points.
3.1
Lackarebäck waterworks and raw water
The experiments that form the basis of Papers III-VII were conducted at
Lackarebäck waterworks in Göteborg, Sweden. The waterworks treats soft,
moderately humic surface water withdrawn from Lake Delsjön, which is low in
turbidity and numbers of microbial indicator organisms. The average raw water
composition is summarised in Table 3.
Table 3: Raw water characteristics for from Lake Delsjön. Average for 2003 (Göteborg
Water and Sewage Works, 2004).
Parameter
Unit
Minimum
Median
Maximum
Temperature
Turbidity
Conductivity
ºC
FNU
mS/m
1.8
0.51
10.2
6.0
0.77
11.0
21.7
1.3
12.2
pH
Alkalinity
Ca2+
Mg2+
Total Fe
Total Mn
NH4+-N
PO4--P
mmol/l
mg/l
mg/l
mg/l
mg/l
µg/l
µg/l
6.6
0.28
6.7
1.6
0.03
0.006
< 50
<4
7.1
0.30
7.4
1.8
0.06
0.14
< 50
<4
7.3
0.39
7.6
1.9
0.10
0.050
< 50
4
Colour
UV254
TOC
mg/l Pt
m-1
mg/l
10
105
3.9
20
118
4.8
20
135
5.7
cfu/100 ml
cfu/100 ml
/10 l
/10 l
<1
<1
<1
<1
2
<1
<1
<1
730
3
<2
<2
Coliform bacteria 35ºC *
E. coli 44ºC *
Giardia
Cryptosporidium
* membrane filtration method (SS 028167).
21
INVESTIGATIONS
Lackarebäck waterworks is designed for a maximum capacity output of finished
water of 5600 m3/h. Treatment consists of pH adjustment by lime, alum
coagulation, flocculation in six consecutive chambers, sedimentation and GAC
filtration followed by final pH adjustment and low-level chlorination. Prechlorination (0.3 mg/l) was applied at water temperatures above 12°C (for further
process details see Paper VI). Both by its raw water composition and long
retention times in the process steps, Lackarebäck waterworks is a typical example
of surface treatment in Sweden. The plant functions comparably well with respect
to NOM removal (Paper VI), and turbidity in drinking water is kept below 0.05
FNU (Göteborg Water and Sewage Works, 2004). As indicated in the literature
study, the microbial barrier function of this type of conventional surface water
treatment may not be fully satisfactory and an upgrade is advisable. Furthermore,
seasonal taste and odour problems and related consumer complaints have been
correlated to elevated numbers of algae in the raw water during late summer.
Although concentrations were low by international comparison, geosmin and
MIB have been found to contribute to earthy–musty odour by means of the
column-sniffing methodology described by Sävenhed (1985). For raw water
geosmin and MIB concentrations, see Paper VI. Records of process parameters
and documented incidents for Lackarebäck waterworks were made available for
research purposes. All experimental studies, with the exception of those carried
out on iron and manganese removal, were conducted in the test facility located in
the basement of the waterworks.
3.2
Systems studies
The idea for the case study concerning systems analysis described in Paper I
originated from a workshop on future drinking water supply systems held within
the Urban Water research school. Alternative water supply scenarios were
compared for the supply of water to a household in Göteborg. The aspects
studied were energy consumption and microbial risk.
The aim was to study the inherent characteristics of centralised and decentralised
water supply. For that purpose, two extreme hypothetical scenarios were devised,
based on the premise that all decentralised water treatment takes place within an
apartment building (Figure 1). Only raw water from Lake Delsjön was to be
supplied centrally. In the first decentralised scenario, the complete flow was
treated by a relatively tight UF membrane and used for all purposes in the
household. A pore size of approximately 10 nm was chosen to ensure virus
removal. A storage tank was included to compensate for the diurnal variation in
water consumption and thereby limit the required design capacity of the unit. In
the second scenario local treatment took place in two steps. The complete flow
passed a MF membrane. For drinking water and food preparation purposes a
portion of the flow was further treated by point of use RO units. For these two
scenarios, Material Flow Analysis (MFA) and Microbial Risk Assessment
(MRA) were carried out by co-workers and compared to the conventional water
supply. Since the study was of a conceptual nature other important issues such as
22
INVESTIGATIONS
cost, membrane fouling and pre-treatment, monitoring and maintenance were not
addressed quantitatively.
Chemical parameters
Alkalinity (mmol/l)
Colour (mg Pt/l)
Mercury (µg/l)
median
0.25
25
< 0.2
max
0.3
40
< 0.2
Raw water
Microbiology
E. coli
Giardia
Cryptosporidium
median
< 1 per 100 ml
< 1 per 10 l
2 per 10 l
max
7
1
3
Flocculation
Distribution network
Activated Carbon
Ultrafiltration
Microfiltration
Chlorination
Storage tank
Storage tank
Distribution network
Drinking water
Consumption 190 l/person*d
Consumption 190 l/pers *d
Food / Drink 10 l Toilet 60 l
Shower 62 l Washing up 30 l
Laundry 8 l Other 20 l
The conventional system
One-step local membrane
Drinking water
Distribution network
RO
Household
180 l/pers*d
Drinking
10l/pers*d
Two-step local membrane
Figure 1: Flowchart of the conventional drinking water system and scenarios for local
membrane treatment in one step and in two steps. Year 2000 water quality data from
Lake Delsjön. Water consumption after (Herrmann and Larsson, 1999) (Paper I).
3.3
Iron and manganese removal from surface water
In cooperation with the water industry, biofilters for the removal of dissolved
iron and manganese species were operated at four waterworks in southern
Sweden. This included three surface water sources (Växjö, Karlskrona, Sotenäs)
and one ground water (Varberg). Different reactor shapes were chosen according
to local conditions. For comparison, identical reference filters were operated at
all sites. These consisted of a polyvinylchloride pipe 20 cm in diameter filled with
a polyethylene biofilm carrier to a bed height of 1.7 m, and operated at an EBCT
of 60 minutes (Heinicke et al., 2000).
At Karlskrona and Sotenäs waterworks, biofiltration was investigated with
subsequent chemical treatment (Figure 2). At Karlskrona waterworks one of the
full-scale treatment trains was operated with biological pre-filtration. An existing
tank volume was converted into a completely mixed reactor with floating plastic
carriers. No physical filter-effect was expected and oxidised iron and manganese
were to be removed in the subsequent continuous upflow filter with a moving bed
of filter sand. At Sotenäs waterworks, a pilot-scale continuous upflow filter was
run using the same conditions (pH, coagulant dose) as the local full-scale plant.
The Sotenäs reactor was designed as a plug-flow column, which was backwashed
at intervals. Although the granular medium was coarse, some retention of
particles in the bioreactor was assumed (Paper II).
23
INVESTIGATIONS
pH adjustment
Coagulant
Wash water
Filtrate
Raw water
‘
Continuous upflow filter
Bioreactor
Figure 2: Combination of biofiltration and chemical treatment at Sotenäs and Karlskrona
waterworks (Paper II).
3.4
Biofilters at Lackarebäck waterworks
At Lackarebäck waterworks three sets of biofilter columns were operated to
serve several purposes. This included
•
The evaluation of filter media (Paper IV, Paper V)
•
NOM removal (Paper IV, Paper VI)
•
Study of barrier function and dynamics of particle removal (Paper V)
•
Taste and odour removal (Paper VI, Paper VII)
•
Effects on subsequent separation processes (Paper VI, Persson et al., 2005
in preparation)
Design parameters common to all biofilters were an EBCT of around 30 minutes
and a gravity-driven down-flow regime with constant flow rate and periodical
backwash with non-chlorinated water. Ozonation was avoided so as not to
introduce further biological instability to the water. Humic matter was to be
removed by means of separation processes, while the biodegradable fraction was
to be degraded by biological activity. The biofilters were fed with untreated
surface water withdrawn upstream of the point of seasonal pre-chlorine dosage.
3.4.1 Biofilters for carrier media study
The effects of carrier media were studied through, eight parallel down flow
biofilters (hard PVC, Ø 20 cm, bed depth 1 m on 15 cm gravel support), which
received untreated surface water from Lake Delsjön, see Figure 3 (Paper IV,
Paper V). Four different carrier materials were evaluated, each in duplicate
columns (Table 4, Figure 4). Crushed EC aggregates have been previously
applied to the biofiltration of drinking water (Melin and Ødegaard, 1999). With
their porous structure they offer favourable conditions for biofilm development
compared to other non-adsorptive media such as sand. Two different EC grain
sizes were included to investigate the effect on filter run length and biofilter
24
INVESTIGATIONS
function. An AC material was included since, even when exhausted for NOM
removal, GAC has in post-flocculation biofilters been reported as a preferable
carrier material over non-adsorbtive filter media (e.g. Uhl, 2000b). Plastic biofilm
carriers, originally designed for wastewater treatment, have been applied to
drinking water biofilters, particularly for iron and manganese removal where
sludge production was expected (e.g. Hedberg and Wahlberg, 1998; Heinicke,
1999). Among the available plastic carriers, the KMT material was chosen for its
high specific surface area.
Table 4: Biofilter media. GAC and EC as applied in (Paper IV Paper V).
1
2
Surface area1
Material
Name
Crushed EC (fine)
Crushed EC (coarse)
GAC
Plastic carriers
Filtralite2 NC 0.8-1.6 mm
Filtralite MC 2.5-4 mm
F3003
KMT4 (PVC, 1.45 g/cm3)
6667 m2/m3
2308 m2/m3
6667 m2/m3
500 m2/m3
Approximated by assuming spherical particles for granular media and a uniform diameter of d10;
Filtralite, Norway; 3 Calgon Carbon Corporation, USA; 4 Kaldnes Miljøteknologi, Norway
Openings along the column side permitted material samples to be taken at 5, 54
and 99 cm of depth. The fine EC and the GAC were backwashed with filtrate
approximately every third week when head-loss of one of the filters exceeded 70
cm. The coarse EC filter could operate for approximately six months without
backwash, while the column with plastic carriers did not accumulate any headloss. Duplicate biofilters were operated for each material to allow for invasive
investigations and material sampling without disturbing the main column. Further
details are provided in Paper IV and Paper V.
Figure 3: Photograph of the biofilters.
25
INVESTIGATIONS
Crushed EC (fine)
Crushed EC (coarse)
GAC F300
Plastic carrier KMT
Figure 4: Photograph of the filter media.
The investigations performed on these filters include the removal of NOM and
barrier function. Particles analyses included online counts using a light extinction
instrument, naturally occurring algae by FC, and removal of added microspheres
and bacteriophages (Paper V). The NOM parameters measured were UV254,
TOC, BDOC and AOC (Paper IV).
3.4.2 Biofilters for odour removal study
To study the mechanism of odorant removal in biological filters operated on
surface water, two filter columns were continuously spiked with geosmin and
MIB. The two parallel sets of 50 mm internal diameter glass columns contained a
total of 1 m of filter bed (Figure 5). Each set consisted of three columns with 20,
30 and 50 cm of material to enable for water and material sampling over the
profile. Corresponding to the pilot-scale biofilters, the EBCT was set to 30
minutes. Biofilter media were GAC (F300, Calgon) and finely crushed EC
(Filtralite NC 0.8-1.6 mm) that had previously been in operation for at least 22
months in the filters described in section 3.4.1. The temperature of the raw water
could be adjusted, and it was supplied by means of a peristaltic pump. An
odorant stock solution was added to give a final concentration of 20 ng/l, marking
the upper end of the range experienced in Lackarebäck raw water (Paper VI). To
minimise the sorption of organics, piping in contact with the odorant solution was
26
INVESTIGATIONS
made of glass and the tubing of PTFE. The columns were backwashed with raw
water three times a week.
The investigation was carried out in two steps. During the first phase, the
biofilters operated at ambient water temperature (6 to 12ºC). During the second
phase, the temperature was controlled at 15°C. At the end of the study, microbial
activity was suppressed by a metabolic inhibitor (sodium azide, 400 mg/l, 6h) to
determine the removal mechanisms. GC-MS analysis of geosmin and MIB was
complemented by an array of microbiological tests to assess the amount of
biomass and its activity (Paper VII).
Surface water
Heater
EC
GAC
Stock solution
Figure 5: Bench-scale column set-up for biological removal of geosmin and MIB (Paper
VII). Filter media were GAC and finely crushed EC.
3.4.3 Biofilters for pre-treatment study
The effect of biological pre-treatment on subsequent processes was studied
through two parallel biofilters, which consisted of stainless steel columns 60 cm in
diameter, with 2 m bed height of GAC (F200, Calgon) and an EBCT of 34
minutes. Since adsorptive removal of NOM was not the primary objective of the
biofilters, exhausted carbon was chosen that had been in full-scale use for four
years. The remaining adsorption capacity at the beginning of the study was
characterised by a methylene blue value of 30 mg/g in comparison to
approximately 250 mg/g for fresh carbon (analytical method see Paper VII). The
filtrate from the two biofilters was blended in a covered stainless steel storage
tank. A 15-minute backwash with raw water was carried out weekly at a target
expansion of 35%, and preceded by a five-minute back-pulse through a
perforated pipe embedded near the bed surface. Openings along the filter wall
permitted for sampling of material and water and were located at 25, 75 and 175
cm below the surface of the filter bed.
27
INVESTIGATIONS
3.5
Process combinations
Biofiltration was applied as pre-treatment to conventional chemical treatment
and NF (Figure 6). The chemical treatment pilot plant was fed either with raw
water or effluent from the biofilters described in section 3.4.3. Except for times of
special campaigns, the feed water was switched weekly to achieve a quasi-paired
data set and counteract seasonal influences on the comparison.
Multi-media rapid filter
Biofilter
Nanofilter 2
Nanofilter 1
Raw water
Intermittent,
change weekly
Flocculation
Sedimentation
GAC-filter
Pilot-scale chemical treatment
Figure 6: Schematic of pilot plant setup. Biofilter (section 3.4.3), multi-media rapid filter
(3.5), chemical treatment (section 3.5.1); nanofilters (section3.5.2).
Two NF pilot plants were operated in parallel. NF-pilot no. 1 (NF 1) received
biofiltrate, while NF-pilot no. 2 (NF 2) received rapid-filtered feed water. The
multi-media rapid filter consisted of a 40 cm diameter pressure vessel and
contained a 70 cm filter bed of anthracite, quartz sand in three sizes, and garnet
sand. During the experimental run, samples were taken weekly. The sampling
involved also GAC filtrate from the full-scale chemical treatment as a reference.
As with the full-scale plant, samples were taken from continuously running
sample taps for raw water, biofiltrate, feed water to chemical treatment, settled
water, GAC filter effluent, multi-media filter effluent and NF permeate.
3.5.1 Chemical treatment pilot plant
The 1 m3/h pilot-scale flocculation-sedimentation-GAC filtration train (detailed
schematic in Figure 7) was operated to closely resemble full-scale treatment.
Treatment at the pilot-scale consisted of sequential addition and static mixing of
NaOH and alum, flocculation in four chambers, sedimentation, and filtration
through AC that had previously been used in the full-scale plant for 2.5 years. For
further process parameters, see Paper III and Paper VI.
28
INVESTIGATIONS
Biofiltration
Surface Water
Flocculation
mix
Settling
GAC Filtration
mix
Alum
NaOH
intermittent
Figure 7: Chemical treatment pilot plant from Figure 6 in detail (Paper VI).
The effect of biological pre-filtration on chemical treatment was studied in terms
of NOM removal and barrier function. The NOM parameters measured were
UV254, TOC, BDOC, AOC and quantitative NOM fractionation by LC-OCD.
The effect on barrier function was characterised by FC particle analysis (Paper
VI).
To improve the sparse quantitative data existing on pathogen removal by
conventional surface water treatment under typical Swedish conditions, spiking
studies were carried out. According to local records, partial or complete failure in
coagulant dosage was the most frequent process incident. To quantify the barrier
function during such incidents, the pilot plant was subjected to episodes of
decreased coagulant dosages with 0, 31, 47 and 100% of the ordinary dose (0, 0.8,
1.2 and 2.6 mg/l of Al3+). At the same time, measurable concentrations of faecal
indicator bacteria were introduced through the addition of clarified wastewater.
Wastewater was added to a final concentration of 2% of dry weather flow
wastewater per litre of raw water over a period of two days. Samples were taken
of the wastewater, feed water, settled water and filter effluent. Addition of
wastewater and sampling commenced when the chemical treatment was in stable
operation. At one occasion, an incident of lower dose (10 mg/l of alum) was
induced during 24 hours, and thereafter the normal dose was restored . Analysis
of faecal indicator bacteria included coliforms and E.coli (most probable number
by Colilert-18TM, Idexx, USA), enterococci (EN ISO 7899-2) and Clostridium
perfringens (ISO/CD 6461-2). Two experimental runs with the addition of
hydrophilic (carboxylate-modified) 1-µm fluorescent latex microspheres
(Molecular Probes, USA) were included with and without wastewater addition.
This was done to compare abiotic particle removal to bacteria removal, and to
check for a possible effect of the wastewater addition on treatment function. The
above mentioned investigation was performed within the pilot plant project by
Johansson and Scott (2004).
In a separate study, the reduction of spiked virus surrogates (bacteriophages) was
quantified when added either before or after the point of coagulant dosage
(Paper III).
29
INVESTIGATIONS
Furthermore, a factorial design experiment was carried out on the pilot plant.
Parameters included coagulant dosage (+/0/- = 1.8/2.6/3.4 mg/l Al3+), mixing (+/= normal mixing/G-value 50% of normal in chambers 1 and 2). Parameters were
UV254, online particle counts and measurement of autofluorescent (FL) particles
by FC.
3.5.2 Nanofiltration
Each NF pilot plant (Veolia, Stockholm, Sweden) housed three commercial 4inch spiral-wound elements in series (NF 270-4040, DOW Liquid Separations,
USA), see Figure 8 to 10. The membrane was a polyamide thin-film composite
designed for high NOM removal and salt passage. A cross flow was applied by
recirculating 1 m3/h from the last to the first element. 25% of the feed flow was
discarded as concentrate. Element recovery (permeate/feed flow) was 15% and
plant recovery 75%. Automated flushing with permeate occurred at 12-hour
intervals for eight minutes at a time. Both plants were operated at a constant, low
flux of 15 l/(m2·h), resulting in feed and permeate flows of 467 and 350 l/h,
respectively. Each NF pilot included a 10 µm cut-off cartridge pre-filter (GX-1020, Osmonics Inc., USA) that was exchanged when the pressure loss exceeded
one bar. Operational data (feed and concentrate pressure, feed, permeate and
concentrate flows) were stored electronically.
Permeate
Feed
Concentrate
10 µm
Figure 8: NF pilot plant from Figure 6 in detail. Concentrate was recirculated from the last
to the first pressure vessel at 1 m3/h, and 25% of the feed flow was discarded.
Chemical cleaning of the two NF units was performed at two to four week
intervals. The 24-hour cleaning procedure consisted of high flow
recirculation/standstill-soaking cycle in 30-minute intervals at 35°C. The cleaning
solution contained 0.1 weight-% NaOH and a surfactant, 0.25 g/l sodium dodecyl
sulphate (SDS). An acid clean at pH 2 was performed at one occasion, but had no
measurable effect on the pressure drop. Since an ICP-MS scan revealed only low
concentrations of metals in the cleaning solution, acid cleaning was considered
redundant. (Persson et al., 2005 in preparation).
The effect of pre-treatment on membrane filtration performance and fouling was
investigated. For that purpose, particle numbers and NOM parameters were
monitored in the raw water, the biofiltrate, the multimedia filtrate and the NF
30
permeates. F. Persson (microbiologist co-worker) measured biofilm formation
potential in the feed- and permeate streams by means of flow cells with glass
slides that were placed into the sample streams. Particle parameters included
online particle counts and naturally occurring particles by FC. On one occasion,
bacteriophages were spiked into the feed of both NF plants (Persson et al., 2005
in preparation).
At the end of the operational phase, destructive analysis of the first membrane
element in each plant was performed in cooperation with Kiwa Water Research,
The Netherlands. The membranes were removed from the pressure vessels and
placed on ice for the transport. Destructive analysis of the elements commenced
24 hours after the end of operation. Pieces of the membrane were sampled
according to (Vrouwenvelder and van der Kooij, 2001) as well as the biofilm that
made up the fouling layer (Figures 10 and 11). The membranes and fouling layers
were analysed for chemical and microbiological parameters.
General parameters included the dry weight of fouling layer and ash content.
Metals were quantified by ICP-MS after microwave digestion of membrane
samples in HNO3. The composition of the organic matter deposited on the
membrane was investigated. A portion of the harvested fouling layer was
resuspended in a mineral salt solution by sonication and rigorous shaking. The
suspension was analysed for TOC and BDOC. AOC and LC-OCD samples were
taken both before and after the BDOC incubation.
The active biomass was determined after low energy sonication of membrane
samples. Parameters included adenosintriphosphate (ATP) and heterotrophic
plate counts (HPCR2A) Complementary analysis of extracellular polymeric
substances (EPS) was carried out by F. Persson.
31
Figure 9: Close-up of an NF element
Figure 10: One of the NF pilot plants
Figure 11: Unrolled NF element
Figure 12: Sampling of fouling material
Figure 13: Sample from NF 1 that
received biofiltrate
Figure 14: Sample from NF 2 that
received multi-media rapid filtrate
32
RESULTS AND DISCUSSION
4
RESULTS AND DISCUSSION
In the following, the obtained results are summarised, spanning the appended
publications. The biofiltration-nanofiltration investigation is being continued by
F. Persson, however results available at the time of publication of this thesis are
included. Complete results will be submitted for publication as (Persson et al.,
2005).
4.1
System studies
The systems analysis study (Paper I) elucidates some inherent characteristics of
water supply from surface sources. If access to raw water is not a limiting factor,
energy consumption is the dominating environmental burden of water supply.
Compared to the total energy consumption in society, its contribution is minimal.
Therefore the other dimensions of sustainability, i.e. social factors, such as the
prevention of waterborne infections are more important than decreasing energy
demand. There are indications that conventional treatment of surface water may
not always be sufficient to keep infection rates below internationally agreed
limits, see Paper I and Westrell et al. (2003). Possible solutions are to include
additional barriers such as UF at the waterworks. In addition, slow sand filtration
and disinfection by UV irradiation constitute barriers, although not for all groups
of pathogens alike (Kärrman et al., 2004).
The MRA was afflicted by considerable uncertainties due to limited data being
available, for example what failure rates could be expected in membrane
filtration. The results nevertheless indicate that integrating membrane filtration
in the process lowered infection risks, provided that water of lower quality was
not used, for example in showering.
Decentralisation with local treatment in the buildings had no considerable effect
on energy consumption. Its potential application however will be limited by cost
and practicality issues. Membrane filtration of surface water requires pretreatment to control fouling. Furthermore, integrity monitoring and maintenance
would have to be solved for a large number of units. Semi-decentralised systems
treating drinking water close to the consumer, but for a larger number of
households, may still overcome some of the abovementioned obstacles.
4.2
Iron and manganese removal
Biofiltration for iron and manganese removal is an established process in ground
water treatment. Surface waters were studied where episodes of dissolved iron
and manganese occurred that posed a problem in treated water. These waters
were not anoxic, but highly humic as indicated by high colour and SUVA values
(Heinicke et al., 2000).
After a start-up phase of approximately one month, manganese removal by the
biofilters could be established at all four investigated sites. The removal rate of
dissolved manganese increased with the influent concentration, so that the
33
RESULTS AND DISCUSSION
effluent concentration was kept at a low and fairly constant level. The columns
with EC media were significantly more efficient than columns containing plastic
carriers, which may be explained by the considerably larger surface area of the
EC compared to the open carriers (Heinicke et al., 2000; Paper II). Particlebound iron and manganese was seen to detach from the filter media, though was
retained by the subsequent filtration. In the investigated surface waters iron in
filtered samples remained largely unaffected by biofiltration through plastic
carriers, but was efficiently removed by coagulation treatment. This suggests that
iron was present as colloidal iron (III) (Heinicke et al., 2000). As expected, the
removal of bulk NOM (measured as UV254) was low in these filters.
With biological pre-treatment, the effluent from chemical treatment by direct
filtration consistently complied with drinking water regulations for manganese
(0.05 mg/l) (Figure 15).
Sotenäs: Mn after Dynasand
0.4
Biol. pre-treated (mg L -1 )
Biol. pre-treated (mg L -1 )
0.15
0.1
0.05
0
0
0.05
0.1
0.15
Ordinary production (mg L-1)
0.3
Karlskrona: Mn after Dynasand
Al-sulfate in both lines
FeCl3, only in bio-line
0.2
0.1
0
0
0.1
0.2
0.3
Ordinary production (mg L-1)
0.4
Figure 15: Total manganese concentrations remaining after chemical treatment at
Sotenäs and Karlskrona waterworks. Each point represents one sampling occasion for
the pilot treatment train with biological pre-filtration, and the full-scale plant (Paper II).
The pilot plant study showed that biological pre-treatment at an EBCT of 30
minutes or more provided a means to remove peaks of high manganese from
surface water feed to conventional treatment plants. For economical feasibility in
full-scale application retention times should be reduced. Goal-oriented process
optimisation would demand the investigation of predominant removal
mechanisms of manganese from humic surface water.
4.3
Geosmin and MIB removal
Since the columns for the study of carrier media (section 3.4.1) showed promising
results for geosmin and MIB removal, the removal by biofilters was further
evaluated at bench-scale. While filtration processes with long retention times are
known to remove geosmin and MIB to a high degree, it could be shown here that
at an EBCT of 30 minutes or less, near-complete removal of these substances
could be achieved from a low start concentration of 20 ng/l. Biofiltration of raw
water was shown to be more efficient for geosmin and MIB removal than the
complete full-scale chemical treatment train that includes 13-minute EBCT postsedimentation GAC filters (Paper VI, Paper VII).
34
RESULTS AND DISCUSSION
At Lackarebäck waterworks, the aim was to reduce geosmin and MIB in drinking
water to below the GC-MS detection limit of 0.5 ng/l. Values in the vicinity of the
lowest published odour threshold values correlated with consumer complaints.
Other substances than geosmin and MIB may contribute to the earthy-musty
odour in the raw water. However, the efficacy of the biofilters in improving its
aesthetic quality was confirmed in odour panel investigations (Li, 2003).
The suppression of bioactivity with sodium azide made it possible to discriminate
between removal mechanisms. Removal of geosmin and MIB completely ceased
on non-adsorptive medium (EC). Semi-exhausted GAC continued to effectively
remove geosmin, while a certain breakthrough of MIB was observed (Figure 16).
GAC Geosmin
GAC MIB
C/C0
0.5
C/C0
1.0
1.5
0.0
0
0
20
20
Depth (cm)
Depth (cm)
0.0
40
Active
60
Suppressed
0.5
Active
60
80
100
100
Suppressed
EC Geosmin
EC MIB
C/C0
C/C0
1.0
1.5
0.0
0
0
20
20
Depth (cm)
Depth (cm)
0.5
40
Active
60
1.5
40
80
0.0
1.0
Suppressed
0.5
1.0
1.5
40
Active
60
80
80
100
100
Suppressed
Figure 16: Geosmin and MIB removal by biologically active and suppressed EC and GAC
columns. Feed concentration aiming at 20 ng/l. Geosmin and MIB concentrations were
below the detection limit in the active GAC and EC samples at 100 cm (Paper VII).
4.4
Evaluation of carrier materials
In the following the results from the investigation of carrier media are
summarised. The plastic medium supported no measurable NOM removal and
considerably lower particle removal than the granular filter media. It was
therefore disregarded for further studies and results were therefore not included
in this summary.
35
RESULTS AND DISCUSSION
4.4.1 NOM and growth potential
During the investigation of carrier media the average DOC in raw water was 4.3
mg/l. BDOC in the raw water varied between 0.69 and 1.35 mg/l corresponding to
25% of the raw water DOC. AOC in the raw water varied between 23 and 68 µg
acetate-C/l (Paper IV).
Approximately 0.5 mg/l DOC was removed in the biofilter with GAC while DOC
removal with fine and coarse EC was approximately 0.25 mg/l (Table 4). The
methylene blue essay indicated a substantial adsorption capacity remaining on
GAC after 22 months of operation on surface water, while EC was nonadsorptive (Paper IV, Paper VII).
BOM parameters were reduced by approximately 30% in the biofilters. The
concentrations after the biofilters were statistically different (p < 0.05) from raw
water concentrations, although no significant differences were detected between
the biofilters. The removal of AOC or BDOC was not significantly affected by
backwashing in any of the biofilters. Several studies have observed a higher
removal of BOM in filters with GAC compared to sand and anthracite media
(e.g. Krasner et al., 1993; Wang et al., 1995) and also compared to porous media
such as pumice (Uhl, 2000b). BDOC removal correlated better to biofilter
metabolic activity than to biomass, in agreement with (Fonseca et al., 2001).
Biofilm formation potential as biomass on glass slides was measured on two
occasions during the experimental period, once during autumn-winter and once
during spring-summer. Biofilm formation was reduced by the biofilters by 80 to
93%. Also in other studies, the decrease in biofilm formation potential has been
found to be superproportional to BOM decrease (e.g. Volk, 2001). The removal
of bulk NOM in the biofilters was modest. However, an 80 to 90% decrease in
biofilm formation potential by biofilters on raw water is expected to help avoid
biofouling problems in subsequent membrane filters (Paper IV).
Table 5: NOM parameters in raw water and biofiltrates with standard deviation (Paper IV).
UV254 (n = 4), DOC (n = 6), BDOC, (n = 6), AOC (n = 8), #(n = 5), *(n = 7).
Raw water
GAC
EC fine
EC coarse
UV254 (m-1)
Removal (%)
12.5 ± 0.7
9.8 ± 0. 6
21%
11.6 ± 0.5
7%
11.8 ± 0. 6
6%
DOC (mg/l)
Removal (%)
4.31 ± 0.26
3.79 ± 0.17
12%
4.07 ± 0.15
5%
4.16 ± 0.16
5%
BDOC (mg/l)
Removal (%)
1.06 ± 0.25
0.70 ± 0.16
32 %
0.74 ± 0.15
30 %
0.76 ± 0.19#
29 %
AOC (µg/l)
Removal (%)
44.5 ± 13.6
34.5 ± 15.8
23%
28.6 ± 10.6
34%
25.4 ± 7.4*
36%
BDOC / DOC
AOC / BDOC
25 %
4%
18 %
5%
18 %
4%
18 %
3%
36
RESULTS AND DISCUSSION
4.4.2 Barrier function
The barrier function of biofilters with different carrier media was investigated by
online particle counts and FC measurements. As could be expected from
filtration without particle destabilisation by coagulants, the biofilters achieved
moderate (60 to 95%) particle removal of µm-size range particles. Raw water
derived (FL) particles were removed to a lower degree than total particles (P),
indicating a difference in properties. The concentration of particles in the
biofiltrate was found to be independent of raw water concentrations for particles
above 1 µm in size (Paper V).
The addition of microspheres revealed a dynamic character of particle retention
and detachment. Only a minor fraction of the hydrophilic (11 to 16%) and
hydrophobic (1 to 3%) microspheres were recovered in the filtrate within three
retention times. After 400 retention times, the released microspheres had
increased to 35 to 37% (hydrophilic) and 15 to 19% (hydrophobic), which is in
agreement with the magnitude of total particle removal (Paper V). After 123 days
of operation (5900 EBCTs) and three cycles of backwashing an estimated total of
11 to 12% of added hydrophobic and 9 to 14% of the hydrophilic microspheres
remained on the media. However, no microspheres were detected in the effluent
from the biofilters at that time. Added bacteriophages passed the filters rapidly,
and were not found on filter media samples. Only approximately 44 to 65% of the
added bacteriophages could however be accounted for, which may indicate losses
through inactivation or clustering.
The observed dynamics of microsphere retention and release offered an
explanation to the independence of influent and effluent particle concentrations.
An equalisation of influent particle peaks would be particularly advantageous for
waterworks that treat surface waters with short-time variation, as described by
e.g. Hurst (2004).
4.5
Biological pre-filtration and chemical treatment
The chemical treatment pilot plant was operated over a period of 15 months
covering a yearly cycle of raw water quality and allowing for the collection of a
large enough data set for statistical analysis. The pilot plant operated with a
slightly, though statistically significant higher efficacy for removal of total
particles and bulk NOM than the full-scale (p < 0.05). The absolute values
however were similar for the two plants so that the pilot plant data was
representative of the full-scale (Paper VI). Chemical treatment data with and
without pre-filtration were measured quasi-paired with a week separating the
two. For TOC, UV254, and FC particle parameters, no significant difference in the
raw water data sets could be found (paired t-test, 2-tail, p > 0.75). External bias
from systematic changes in raw water quality could therefore be excluded.
37
RESULTS AND DISCUSSION
4.5.1 NOM
The removal of NOM and its biodegradable fraction was followed over the pilot
plant. The biofilters removed on average approximately 10% of the raw water
NOM measured as TOC and UV-absorbance. As could be expected the
biodegradable fraction was removed to a higher degree than bulk NOM (Figure
17). The SUVA was not altered by biofiltration since a similar ratio of TOC and
the UV-absorbing NOM was removed. This is in agreement with findings from
slow sand filtration and riverbank filtration (Ludwig et al., 1997; Graham, 1999).
With the exception of AOC, the lower feed NOM concentration to the chemical
treatment caused by biofiltration resulted in significantly, though not
proportionally, lower effluent concentrations after chemical treatment (p < 0.05,
Figure 17).
Fractionation and quantification by LC-OCD gave further insight into how
different NOM fractions were affected by the treatment processes (Figure 18).
Particularly noteworthy was the fact that 10% of NOM did not elute from the
chromatography column and could therefore not be characterised further by this
method. Since no considerable particulate NOM was present, the fraction must
consist of natural hydrophobic matter. The hydrophobic fraction was removed by
chemical treatment to less than 50% and passed a subsequent ultrafilter
unaffected. Although hydrophobic molecules are expected to easily adsorb to
uncharged hydroxide floc formed in flocculation, low molecular weight
hydrophobic like geosmin and MIB demonstrate similarly low removal rates over
chemical treatment. The hydrophobic fraction made up more than half of the
DOC removed by the biofilters, either trough biodegradation or adsorption.
100%
Removal
80%
60%
40%
20%
0%
TOC
UV-254
Biofilter
Chemical treatment with biofilter
BDOC
AOC
Chemical treatment alone
Full-scale
Series5
Figure 17: Removal of bulk NOM and BOM by biofiltration and pilot-scale chemical
treatment. Error bars are standard deviations.
38
RESULTS AND DISCUSSION
Humics Building
blocks
Acids and
LMW humics
Polysaccharides
Neutrals
Raw
water
After
biofilter
After chemical
treatment
Figure 18: LC-OCD chromatogram of raw water, biofiltrate, and after chemical treatment.
Batch testing of flocculation-sedimentation-filtration gave results that were
consistent with those from the pilot-scale (Paper VI). Chemical treatment with
biological pre-filtration achieved a similar removal of UV-absorbing substances
with an approximately 25% lower coagulant dose than ordinary treatment.
4.5.2 Barrier function
4.5.2.1 Chemical treatment
The reduction in the concentration of faecal indicator bacteria added with
wastewater is summarised in Table 6 as log-reduction, -log (Cout/Cin).
Table 6: Reduction of bacteria and microsphere concentrations added to the pilot plant.
The values are log-reduction, -log (Cout/Cin), if not stated otherwise. n.m.: not measured.
Modified from Johansson and Scott (2004)
Parameter
Coliforms
E. coli
Enterococci
C. perfringens.
Microspheres
Coagulant dose (mg/l Al3+)
2.6
1.2
0.8
Feed (ml-1)
1140 - 1700
160 - 255
20 - 135
4 - 120
7700 - 9000
(n = 5)
(n = 5)
(n = 5)
(n = 5)
(n = 2)
3.7 - 3.8 (n = 3)
3.5 - 4.2 (n = 3)
3.3 - 4.1 (n = 3)
2.7 - 3.3 (n = 3)
3.3 - 3.5 (n = 2)
39
2.6
3.0
3.9
2.7
n.m.
1.0
1.3
0.9
1.5
n.m.
0
0.2
0.1
0.2
0.4
n.m.
RESULTS AND DISCUSSION
Although variations of wastewater composition were largely compensated for by
adjusting the volume that was added, the above comparison is affected by
changes in raw water composition and temperature. Absolute values should
therefore be regarded as indicative. The minimum coagulant dose for floc
formation was between 0.8 and 1.2 mg Al/l (i.e. 0.17 and 0.26 mg Al per mg/l
TOC). At low coagulant dosage, the removal of particles by flocculation/
sedimentation deteriorated and increased the load to the rapid media filter
resulting in higher numbers of faecal indicator bacteria in first filtrate after
backwash. Figure 19 shows the time-series of a simulated disturbance of
coagulant dosage lasting 24 hours. The vertical lines mark the beginning and end
of the 0.8 mg/l dosage. The formation of visible floc ceased, as did the effect of
flocculation/sedimentation on coliform numbers.
-1
Coliform bacteria (ml )
10000
1000
100
10
1
Feed
Settled
Filtrate
0.1
0.01
-2 0 2 4 6 8 10 12 14 16 18 20 22 24 26 28 30 32 34 36 38 40 42 44 46 48 50
Time (hours)
Figure 19: Coliform bacteria in the chemical treatment pilot plant during a simulated
disturbance of coagulant dosage. Most probable number by Colilert-18™. Vertical lines
delimit the beginning and end of the phase with 0.81 mg/l Al dosage (31% of normal). The
last three data points in the filtrate are directly prior to a backwash, directly after, and
one hour after. Modified from Johansson and Scott (2004).
The factorial design experiment of process parameters indicated no conclusive
effects. The process was particularly insensitive to mixing speed in flocculation –
a complete stop of mixing in chambers 2 and 4 caused by a thunderstorm
overnight had no measurable effect on effluent online particle counts (data not
shown).
The reduction of added concentrations of virus surrogate (bacteriophages)
depended both on the bacteriophage type and the point of addition in the
treatment train (Paper III). The log-reduction of MS-2 bacteriophages was higher
than for φX174 bacteriophages and reduction took place mainly in the initial
stages of the treatment train, i.e. coagulant addition and mixing. The φX174
bacteriophage appeared to be a more reliable conservative surrogate for human
virus behaviour in chemical treatment than MS-2. Over the current treatment
train a 3.8-log reduction in φX174 was achieved.
Conventional treatment appeared to be a robust though limited barrier against
particles in the size range of pathogenic bacteria and parasitic protozoa. Barrier
40
RESULTS AND DISCUSSION
function appeared robust against changes in process parameters. The experienced
robustness is somewhat contradictory to reported failures in barrier function
during sub-optimal filtration (e.g. Emelko, 2003), and may be due to the
conservative design of the treatment train investigated, with long retention times.
In case an additional particle barrier was to be included in the process train (i.e.
UF), coagulant dose may be designed to satisfy only colour removal, which would
allow for lower doses.
4.5.2.2 Chemical treatment with biological pre-filtration
To study barrier function in chemical treatment quantitatively, the parameter of
total particles was not meaningful, since their concentration was increased by
coagulation (Paper VI). Instead, naturally occurring algae were used as a tracer
throughout the process. Over the period of 15 months (n=22), the biofilters
removed FL particles on average by 56% (FL 0.4-1 µm) and 63% (FL 1-15 µm).
This is in agreement with the more limited earlier data set included in Paper V.
The percentage removal increased with raw water particle concentration so that
the filters had an equalising effect on the feed to the subsequent chemical
treatment. With pre-filtration the lower particle load to chemical treatment
resulted in significantly lower concentrations of FL particles in GAC-filtrate
(p<0.05). The average ratio of raw water particles passing the complete treatment
train decreased from 10 to 5% (FL 0.4-1 µm) and 1.1 to 0.3% (FL 1-15 µm)
(Paper VI). The equalisation of variability in raw water particle concentration
increased the robustness of the barrier function in chemical treatment, which will
have implications for MRA. Similar to the results obtained for NOM removal,
the coagulant dose could be lowered by 25% to achieve a similar filtrate particle
concentration as in the process without pre-filtration.
4.6
Biological pre-filtration and nanofiltration
4.6.1 Permeate quality
The performance of the NF was in agreement with the literature summarised in
section 2.3.3 as high removal of NOM was achieved. Low-molecular weight
organic acids were the only fraction passing the membrane to an appreciable
extent (Figure 20). Average AOC concentrations in the permeate were 5 to 7 µg
acetate-C/l (n=9), comparable to the levels found after chemical treatment (Paper
VI).
41
RESULTS AND DISCUSSION
Humics Building
blocks Acids and
LMW humics
Raw
water
Polysaccharides
Neutrals
After
biofilter
After
nanofilter 1
Figure 20: LC-OCD chromatogram of raw water, biofiltrate, and after nanofiltration. The
signal height for nanofilter permeate is enlarged by a factor of 10.
The barrier function of the NF was investigated by the removal of naturally
occurring particles and added bacteriophages. During the challenge test
quantifiable concentrations of bacteriophages occurred in NF permeate. A 8-log
virus retention however confirmed the integrity of the membrane elements
(Table 7). Rapid media filtration did not remove bacteriophages while the low
reduction by biofilters was in agreement with the results presented in Paper V.
Table 7: Log-reduction of added bacteriophage by NF pilot plants and pre-treatment
processes.
Process
Biofilter
Multi-media rapid filter
NF 1 (bio pre-filtration)
NF 2 (rapid filter)
φX174
Bacteriophage
MS-2
0.31
0.02
7.9
7.8
0.09
-0.02
8.3
7.7
The fractions of raw water particle concentrations remaining in NF permeate are
given in Figure 21. The removal of total particles (P) was considerably lower than
of autofluorescent (FL) particles. Since a selective retention of FL particles
appear unlikely this finding suggests that particles were formed on the clean side
of the membrane for example through proliferation of microorganisms.
According to the measurements, large FL particles were retained to a lesser
degree than small ones. Even for FL particles in the size 0.4 to 1 µm log-removals
were 4.3-log, which was considerably lower than the removal of nm-size
bacteriophages. It may therefore be assumed that the occasional detection of
singular FL particles in NF permeate was at least partly due to artefacts in the FC
42
RESULTS AND DISCUSSION
method. With the concentration of FL particles in the raw water, it was not
possible to measure a higher log-removal.
C/C0 (raw water)
4%
NF 1 (bio pre-filtration)
3%
NF 2 (rapid filter)
2%
1%
0%
P 0.4-1 µm
P 1-15 µm FL 0.4-1 µm FL 1-15 µm
Figure 21: Removal of total (P) and autofluorescent (FL) particles by the NF pilot plants
including respective pre-treatment (n=48, n=52).
4.6.2 Fouling
4.6.2.1 Feed water and pressure drop
Both nanofiltration pilot plants were operated in parallel for 19 months prior to
destructive analysis of one membrane element from each plant. The elements in
both plants developed comparably rapid pressure drop over the membrane that
indicate fouling problems. Pressure drop development was particularly rapid in
the elements fed multi-media filtrate where permeability could not be fully
recovered by alkaline cleaning (Table 8). In the water industry, chemical cleaning
of NF elements is performed when the pressure drop has increased by about 30%
from the original value. Likewise, cleaning intervals of less than around one
month and membrane life times below 3 years are not considered economically
feasible. Thus membrane filtration with the investigated NF element type at a
flux of 15 l/(m2·h) was not feasible on the applied surface water with multi-media
rapid filter pre-treatment. Although the pressure drop was mediated by
biofiltration, it was still relatively high compared to the demands of the water
industry. Further investigations were undertaken to determine the causes for the
difference in fouling rate between the pre-treatment options.
Table 8: Pressure drop development (bar) during the last cleaning cycle of NF pilot plant,
17 days.
Membrane
NF 1
NF 2
After cleaning
(bar)
Prior to autopsy
(bar)
Increase
(bar)
Increase
(%)
1.7
3.1
2.3
5.1
0.6
2.0
35 %
65 %
Parameters of NOM in biofiltrate (feed to NF 1) and multi-media filtrate (feed to
NF 2) are summarised in Table 9. Biofiltrate feed to NF 1 contained significantly
lower concentrations of TOC and BDOC, while AOC values were similar. At
one occasion, the feed waters were characterised by LC-OCD. Both waters
43
RESULTS AND DISCUSSION
contained measurable concentrations of polysaccharides that have previously
been shown to accumulate on NF membranes.
Table 9: NOM characteristics of feed water to NF pilot plants. t-test between feed waters.
NF 1 (feed from biofilter)
NF 2 (feed from rapid filter)
2-tail paired t-test
TOC
(mg/l)
(n=55)
BDOC
(mg/l)
(n=8)
AOC
(µg C/l)
(n=9)
Polysaccharides*
(µg C/l)
(n=2), (n=1)
4.30 ± 0.30
4.67 ± 0.33
p < 0.01
0.86 ± 0.22
1.06 ± 0.21
p < 0.01
28 ± 9
26 ± 5
p = 0.5
187±3
215
n.a.
* by LC-OCD; n.a.: not applicable
The concentrations of FL particles in the two feed waters are shown in Table 10.
Biofiltrate contained in average two to three times less particles and FL particles
than rapid filter effluent. Additionally the concentrations of total iron and
manganese were significantly lower in biofiltrate than in the effluent of the multimedia filter (33 ± 28 compared to 53 ± 29 mg/l Fe and 2 ± 1 compared to 6 ± 3
mg/l Mn, n=61).
Table 10: Particle content (flow cytometry) in feed water to NF pilot plants (ml-1), n=55.
P: particles, FL: autofluorescent particles.
NF 1
NF 2
t-test, paired
P 0.4-1µm
P 1-15µm
FL 0.4-1
FL 1-15µm
1.2x105 ± 5.0x104
3.6x105 ± 1.3x105
p < 0.001
2.5x103 ± 1.2x103
1.0x104 ± 3.0x103
p < 0.001
2.9x103 ± 2.5x103
6.4x103 ± 6.3x103
p < 0.001
1.1x103 ± 0.7x103
2.6x103 ± 1.9x103
p < 0.001
Biofilters operated at an EBCT of 30 minutes were superior to multi-media
filtrate pre-treatment of NF regarding NOM and inorganic parameters. Further
microbial results (EPS, biofilm formation potential) will be included in the
journal publication (Persson et al., 2005 in preparation).
4.6.2.2 Destructive analysis of membrane elements
Visually, the fouling layer was evenly distributed over the length of the
membrane sheets (Figures 11 and 12). The membrane harvested from NF 2,
which had been operating on rapid filtrate, was more distinctly coloured than the
membrane operating on biofiltrate (Figures 13 and 14, and title page).
Biofouling was evaluated by the concentration of active biomass present on the
membrane (Table 11). The average ATP concentration over the length of the
element was 50% higher for NF 2 than for NF 1. Levels of active biomass (ATP
and HPCR2A) were within the range of concentrations found at previous
autopsies (Vrouwenvelder and Van der Kooij, 2002). Biomass concentrations
above 2000 pg ATP/cm2 have been associated with increased pressure drops in
NF and RO systems (Vrouwenvelder et al., 1998). However, the difference in
biomass level between the two membranes was modest and cannot alone explain
44
RESULTS AND DISCUSSION
the high pressure drops observed over NF 2 operated on multi-media filtered
water.
Table 11: Concentrations of active biomass as ATP and HPCR2A,
spacer, membrane and product spacer.
Average over length (n=10)
ATP (pg/cm2)
NF 1
NF 2
10 d, 25°C
including feed
Mid-length of element (n=1)
ATP (pg/cm2)
HPCR2A (cfu/cm2)
1800 ± 700
2700 ± 800
3.3x105
2.1x105
2300
3200
ICP-MS analyses of metals on the membrane surfaces showed that iron and
aluminium were present at the highest concentrations (Table 12). The level of
inorganic material on the membrane was moderately elevated in comparison to
previous investigations (Vrouwenvelder and Van der Kooij, 2002) and may have
contributed to the observed pressure drop.
Table 12: Concentrations of selected elements on the membrane surface (mg/m2). Sum of
feed spacer, membrane and product spacer. New: New membrane as blank values.
NF 1 (n=4)
NF 2 (n=4)
New (n=2)
Al
Zn
K
Ca
Mg
Mn
Fe
32.7±1.6
71.3±3.6
0.2±0.1
0.9±0.1
1.9±0.2
0.1±0.0
11.4±0.4
27.6±1.3
0.5±0.1
5.8±0.3
13.1±1.1
0.8±0.2
2.9±0.1
9.9±0.9
0.1±0.0
2.3±0.1
7.6±0.8
0.2±0.0
43.7±1.3
134.7±17.3
1.7±0.0
The NOM contained within the fouling layer was investigated in detail (Table 13)
by means of LC-OCD fractionation, BDOC and AOC analysis.
Chromatographable DOC (CDOC) constituted 11% of the fouling material dry
weight in NF 1 and 8% in NF 2. Polysaccharides that represent only a small
fraction of raw water NOM (Paper VI) made up more than half of the CDOC on
the membrane. Mass balance calculations by Huber (1998) have shown a
deposition of polysaccharides on membrane surfaces. During BDOC incubation,
53 to 60% of the polysaccharides were removed from the water phase.
Experience from the operation of the pilot plant provided further evidence for a
certain biodegradability of the fouling layer on the membrane. On occasions
when the pilot plants were shut down due to technical problems or power
failures, a two-day standstill resulted in a considerable increase in permeability.
This phenomenon may be considered a sort of “biological cleaning”.
Table 13: Characterisation of NOM on the membrane surface by LC-OCD, BDOC and AOC
Membrane
NF 1
NF 2
CDOC*
(µg C/m2)
Polysaccharides*
(µg C/m2)
BDOC/CDOC
(---)
AOC
(µg C/ m2)
248 ± 21
475 ± 14
161 ± 7
272 ± 10
54%
36%
13
28
* by LC-OCD (n=2) CDOC=chromatographable DOC.
45
RESULTS AND DISCUSSION
The NF plant receiving rapid-filtered surface water had 50 to 100% higher
concentrations of most investigated parameters on the membrane. Although no
single parameter convincingly explained the large difference between the two
pilot plants, together these factors have contributed to the increased pressure
drop.
46
CONCLUSIONS AND FURTHER WORK
5 CONCLUSIONS AND FURTHER WORK
Within this work, part of the Sustainable Urban Water Management program,
options for surface water treatment were investigated. This included a systems
analysis study on options for decentralised water treatment under Swedish
conditions. Biological pre-filtration of surface water was investigated for the
removal of NOM, particles, iron and manganese and taste and odour compounds.
The barrier function of a treatment facility typical for surface water supply in
Sweden was quantified. From the results obtained in this thesis and taking into a
consideration the available literature, the following conclusions can be drawn:
Systems aspects
Under conditions where raw water sources were not a limiting factor,
environmental impact was not a major issue for the development of water supply
systems as it represents a minor contribution to society’s total energy
consumption. Other factors such as minimising infection risk are therefore more
important. Hypothetical decentralised systems with treatment closer to the
consumer did not create an increased energy demand and therefore appear
unrestricted by that factor. Reasons of practicability however favour more largescale technology.
Iron and manganese
Biological pre-treatment was shown to be a viable process to remove dissolved
manganese also from difficult-to-treat humic surface waters. Dissolved iron was
not consistently removed by the biofilters, but constituted no problem for
subsequent granular media filtration. The mechanism of manganese removal in
the biological pre-filters catalysed chemical oxidation on MnO2 or biological
oxidation could not be determined in this applied study. In highly coloured
waters where NOM affects iron and manganese speciation through complex
formation optimal conditions and achievable reaction rates may differ from
experiences with ground waters. The elucidation of the mechanism of manganese
oxidation in humic surface waters and the influence of process parameters would
be helpful to optimise the process primarily with regard to shorter retention
times.
Barrier function
Risk assessment carried out within the Urban Water group indicated a possibility
for a non-negligible frequency of waterborne disease from conventionally treated
surface water. These results are however afflicted with considerable uncertainty.
The experimental study contributed data on the microbial barrier function of the
most common surface water process under Swedish conditions with regard to
47
CONCLUSIONS AND FURTHER WORK
particle and virus removal. Bacteria removal was quantified under simulated
malfunction of coagulant dosage. The data may be used to achieve more accurate
and precise MRA predictions. Conventional treatment constituted a mediocre
barrier against particles in the size of pathogenic bacteria and parasitic protozoa.
Barrier function however appeared robust against moderate changes in process
parameters. Biological pre-filtration was shown to lower the load of raw water
derived particles to the subsequent processes by approximately 80%. Peak loads
of added particles were equalised by high initial retention followed by a slow
release of attached particles. Protection against peak loads increases the
robustness of the microbial barrier and should decrease microbial risk in
conventional drinking water treatment. The alleviation of particle peaks in the
feed also protects chemical treatment trains that are operated at constant
coagulant dosage from sub-optimal operating conditions. Further work should
include more detailed investigation of treatment variability in chemical treatment
and its effect on microbial barrier function. For waterworks, the development of
rational guidelines would be useful, regarding the significance of specific
incidents in the treatment process. Most urgently, existing knowledge about the
variability of raw water quality and treatment processes, as well as technology for
online process control, need to be applied in the water industry on a wider scale.
NOM
Without pre-ozonation, the effect of biofiltration with an EBCT of 30 minutes on
bulk NOM was low (approximately 10%). The biofilters however reduced a
fraction of hydrophobic NOM that otherwise passed the subsequent chemical
treatment (and an ultrafilter) to a large extent. Removals of BOM (AOC,
BDOC) was higher than bulk for NOM (approximately 20%). Biofilm formation
potential in biofilter effluent was decreased to a higher degree than measures of
BOM in the water phase. In the future, the effect of treatment processes on
emerging organic contaminants need to be investigated such as drug residuals.
Trace compounds causing earthy-musty odour
Two compounds (geosmin and MIB) that cause earthy-musty odour episodes in
surface water were effectively removed by biofiltration at an EBCT of 15
minutes. The removal mechanism was investigated on semi-exhausted GAC and
on non-adsorptive EC media. With regard to EC, removal was entirely through
biological activity, while on GAC an adsorption capacity remained when
metabolic activity was suppressed. Therefore also GAC that has been in
operation for several years for the pre-treatment of surface water added
robustness to the removal of hydrophobic and biodegradable trace organics that
may occur intermittently in the raw water. Further evaluation of the results is
under way, with the modelling of reaction rates of geosmin and MIB. For
accurate quantification of adsorption, it would be advantageous to determine
48
CONCLUSIONS AND FURTHER WORK
adsorption isotherms in the relevant concentration range for the activated carbon
used in this study.
Membrane fouling
Simple rapid filtration pre-treatment of typical Swedish surface water caused a
rapid in pressure drop in NF membranes operated at a comparably low flux. In
comparison to pre-treatment by rapid media filtration, biofiltration significantly
moderated the development of pressure drop. The feed water originating from
the biofilter contained lower concentrations of particles and biodegradable
organic matter than rapid filtrate. Destructive analysis of the NF elements was
carried out for inorganic, organic and microbial parameters. An interesting
finding was that polysaccharides, only a small fraction of feed water NOM,
constituted more than half of the chromatographable dissolved organic carbon
material in the fouling layer. Although no single parameter convincingly
explained the large difference between the two pilot plants, it is likely that
together these factors have contributed to the improved operation of NF
membranes fed with biofilter effluent. Since the majority of the NOM
resuspended from the fouling layer sample were polysaccharides, it would be
advantageous to optimise biofilter pre-treatment with regard to removal of these
compounds.
Applications
Biological pre-filtration has been applied in full-scale at one of the waterworks
involved in the project, and is considered at two more waterworks. The technique
may be particularly interesting for the treatment of raw waters with episodes of
taste and odour or dissolved iron and manganese species. Furthermore, biological
pre-treatment offers – as a sort of bank filtration “light” – an additional barrier
against peak loads of particles and NOM in surface waters. Integrating a new
process into an existing conventional treatment train implies considerable
investment costs. In each specific case, the benefits need to be weighed against
cost and compared to other options of process upgrade, such as ozonationbiofiltration, UF for increased particle removal, or nanofiltration.
49
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