Set-up MSc report

Struvite recovery from digested sludge
Struvite recovery from digested sludge
At WWTP West
Delft University of Technology
Faculty of Civil Engineering and Geosciences
Department of Water Management
Section of Sanitary Engineering
Stevinweg 1
2628 CN Delft
www.drinkwater.tudelft.nl
Bart Bergmans
Bart Bergmans
February 2011
Struvite recovery from digested sludge
At WWTP West
Bart Bergmans
for the degree of:
Master of Science in Civil Engineering
Date of submission: February 10, 2011
Date of defense: February 24, 2011
Committee:
Prof.dr.ir. L.C. Rietveld
Prof.dr.ir. J.B. van Lier
Prof.dr.ir. M.C.M. van Loosdrecht
Ing. A.M. Veltman
Delft University of Technology
Sanitary Engineering Section
Delft University of Technology
Sanitary Engineering Section
Delft University of Technology
Environmental Biotechnology Section
Waternet
Afdeling Onderzoek & Projecten
Sanitary Engineering Section, Department of Water Management
Faculty of Civil Engineering and Geosciences
Delft University of Technology, Delft
Abstract
The implementation of biological nutrient removal processes in wastewater treatment plants
(WWTPs) has led to higher concentrations of phosphate in the excess sludge. Part of the
phosphate is released into the liquid phase during anaerobic digestion. Under specific
conditions, this phosphate reacts with magnesium and ammonium to form a precipitate called
‘struvite’.
Over the past few years, struvite has caused scaling problems in pumps, pipes and
dewatering facilities of the sludge line at WWTP West (Waternet). A process called ‘AirPrex’ is
available for the controlled formation and removal of struvite directly from the digested
sludge. It is claimed that the AirPrex process is not only effectively preventing scaling
problems, but is also improving the sludge dewaterability. On top of that, the recovered
struvite can be used as a phosphate fertilizer.
In the AirPrex process the digested sludge is led through a reactor tank where air is supplied
and magnesium is added as magnesium chloride (MgCl2). Air is supplied for raising the pH
value (by CO2 stripping) and for mixing the sludge and the magnesium chloride. The formed
struvite is intermittently tapped from the conical reactor bottom. In a second tank, smaller
crystals are allowed to settle.
During a pilot at WWTP West in April 2010, it was found that the AirPrex process could
effectively reduce the orthophosphate concentration of the digested sludge. Besides that, the
solid content of the dewatered sludge improved from 22% to 25%. This improvement in
dewaterabilty could significantly reduce the sludge disposal costs, since it reduces the sludge
volumes that need to be transported and incinerated. During the pilot, the effects of pH and
magnesium variations on phosphate removal and sludge dewaterability were not determined.
In this MSc thesis, these effects were investigated by conducting laboratory experiments in a
batch reactor.
Based on the experimental results, a Matlab model was made to describe the relations
between pH, phosphate removal and magnesium dosage. A strong correlation was found
between free PO43- and free Mg2+ concentration. It was concluded that assuming a constant
NH4+ concentration is allowable within the concentration ranges that were studied. From the
model it seemed that even at low pH values (pH≈7.2) the larger part of the phosphate could
be removed using a reasonable magnesium dosage.
Filtration tests showed that higher pH values (pH>7.5) led to a decrease in dewaterabilty
when no magnesium chloride was dosed. Dosing of magnesium chloride led to an increase in
dewaterability, regardless the pH value. The best dewaterability result was found with a slight
magnesium over-dosage (Mg:PO4=1.2) while stirring the sludge instead of aerating it.
Unfortunately, the struvite that was formed during the experiments was not successfully
separated from the sludge. It was concluded that the struvite crystal size was too small. This
was probably caused by high local magnesium over-dosage, stimulating the nucleation of new
crystals instead of crystal growth. Future research should focus more on crystal growth and
the removal of struvite from digested sludge. The struvite that was found during the
experiment was probably already present in the digested sludge and could have been formed
in the anaerobic digesters.
Next to the AirPrex process, several other reactor alternatives seem suitable for
implementation at WWTP West, such as a stirred tank reactor or a fluidized bed reactor.
Another alternative is to adapt the digested sludge buffer tank (USB) in such a way that
struvite can be directly recovered from it. To be able to make a balanced choice, these
alternatives should be further analyzed on performance and costs. Besides that, more
research is needed on struvite crystal growth in digested sludge and on struvite/sludge
separation, since struvite is a valuable product that should not be wasted.
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Preface
This report is the final result of my graduation research project. I performed this project at
Waternet, the watercycle company of Amsterdam, to complete my MSc studies on Sanitary
Engineering at the TU Delft.
The subject of this research, ‘Struvite recovery from digested sludge’, has a somewhat
‘unattractive sound’ in it, as I noticed from the initial responses of many people that asked for
my graduation subject over the last 10 months. However, after only a short explanation most
people become very enthusiastic, as they realize this subject is in fact really exciting and of
great current interest. For me, this project was a great opportunity to learn a lot more about
wastewater treatment processes and to witness the regular day-to-day practice on a
wastewater treatment plant. And although it did not always smell like flowers, I enjoyed
every bit of it!
I would like to take the opportunity to thank all people that helped me and supported me
during this research.
First of all, I want to thank all people at Waternet for giving me the opportunity to perform
this research project at their company. Alex Veltman, for being my supervisor at Waternet,
for his help with setting up the experiments and for all nice discussions. Joost Kappelhof, for
all his sharp feedback and for speeding up the process whenever it seemed to stagnate. Hans
Nieuwenhuizen, Peter Wind, Michel Collin and Marcel van der Blom, for providing me with all
the required equipment to successfully complete the experiments and for all their effort in
making things work. And everybody at wastewater treatment plant West, for their hospitality,
their interest and their help with all kinds of small problems.
Also special thanks to Sepp Helbers and BASF Nederland, for borrowing me the equipment for
dewaterability measurements.
Furhtermore, I would like to thank several people at the TU Delft. Professor Luuk Rietveld, for
being an excellent and enthusiastic supervisor, not only during my graduation project, but
also during my additional thesis project and during my internship in Surinam. Professor Jules
van Lier and professor Mark van Loosdrecht for their feedback and all their suggestions to
improve my report. And Patrick Andeweg, for donating the laboratory column and for the
very useful information on rheology and crystallization.
I also want to thank my friends and housemates, for their unique personalities, for always
being there and for the necessary recreation after many study days.
Most of all, I would like to thank my family. My parents, Martin and Odette, for all their
patience, trust, help and support during my entire studies. My brother Paul, for being a lot of
fun and for all his good hugs. My sister Margot for all her crazy stories and for always keeping
in touch. And my sister Hanneke, for the unbelievable amount of trust. Special thanks to
Martin and Hanneke for all the critical comments on my report: it helped a lot!
Bart
February 2011
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Content
Introduction
1.1
Phosphorus and phosphate
1.2
Phosphate issues
1.3
Phosphate recovery
1.4
Wastewater treatment plant West
1.5
Thesis outline
2.
State of the art
2.1
Classification
2.2
Fully operational techniques
2.3
Developing techniques
2.4
General state of development
3.
Research outline
3.1
Research problem
3.2
Research approach
4.
Theoretical background
4.1
Sludge
4.2
Struvite crystallization
4.3
CO2 stripping
5.
Materials and methods
5.1
Experimental setup
5.2
Experimental procedure
5.3
Mathematic modeling
6.
Results and discussion
6.1
Digested sludge conditions at WWTP West
6.2
Sludge dewaterability
6.3
Phosphate removal
6.4
Struvite recovery
6.5
CO2 stripping
7.
Implementation at WWTP West
7.1
Objectives
7.2
Constraints
7.3
Potential profits
7.4
Alternatives
7.5
Next steps
8.
Conclusions and recommendations
8.1
Conclusions
8.2
Recommendations
List of references
Annexes
Annex 1 – Process scheme water line WWTP West
Annex 2 – Process scheme sludge line WWTP West
Annex 3 – Summary of the Waternet pilot
Annex 4 – Experimental data
Annex 5 – Matlab script for component availability and saturation
Annex 6 – Matlab script for pH increase due to CO2 stripping
1.
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List of illustrations
1.1
1.2
1.3
1.4
1.5
2.1
2.2
2.3
4.1
4.2
4.3
4.4
4.5
4.6
4.7
4.8
4.9
4.10
4.11
4.12
4.13
4.14
5.1
5.2
5.3
5.4
5.5
5.6
5.7
6.1
6.2
6.3
6.4
6.5
6.6
6.7
6.8
6.9
6.10
6.11
6.12
6.13
6.14
6.15
6.16
6.17
6.18
6.19
6.20
6.21
6.22
6.23
6.24
6.25
6.26
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Key processes of the natural phosphate cycle
Scaling problems in pipes caused by struvite deposits
Treatment scheme of WWTP West
Average daily phosphorus loads at WWTP West
Thesis outline
AirPrex technology
The Seaborne process
The Sephos process
Basic activated sludge process
The UCT-principle
Dewatering centrifuges at WWTP West
Dewatered sludge
Divalent Cation Bridging theory
Distribution of magnesium over different compounds
Distribution of ammonium over different compounds
Distribution of phosphate over different compounds
Relative supersaturation as a function of pH
The sensitivity of supersaturation for ionic strength variations
Concentration gradients adjacent to a crystal surface
Influence of supersaturation on nucleation and crystal growth rates
Determination of the solubility and supersolubility curves
Carbonic equilibrium
Short column (schematically)
Short column (impression)
Aeration system short column
Long column (schematically)
Long column (impression)
Dewaterability test (schematically)
Dewaterability test (impression)
Close up of sediment
Watery zone in the sludge
Seasonal orthophosphate variation
Pre dewaterability
Dewaterability at aeration to different pH values without MgCl2 dosing
Dewaterability at different MgCl2 dosages
Dewaterability at different pH values after MgCl2 dosage
Dewaterability at different polymer dosages after treatment
Relations between component concentrations
Error including ammonium
Error excluding ammonium
Determination of needed MgCl2 dosage
Removal ratios at different pH setpoints
Removal ratios at different dosing ratios
Concept of the model
Removal curves
pH drop at different dosing ratios
pH drop at different initial pH values
Cumulative H+ increase at different initial pH values
pH drop with stirring instead of aerating
Sieve curves from experiments with different initial pH values
Sieve curves from experiments with different dosing ratios
Exceeding of the metastable limit due to high dosing ratios or high pH values
Component concentrations during the growth experiments
Sieve curve from the growth experiment
pH increase at different airflow rates
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6.27
6.28
6.29
6.30
6.31
6.32
7.1
7.2
7.3
7.4
7.5
7.6
7.7
7.8
7.9
7.10
7.11
7.12
pH as a function of RQ value
pH increase in the pilot (without struvite formation)
pH increase in the experiments (without struvite formation)
pH increase at different filling heights
Lower overall pH course due to struvite formation
pH increase after struvite formation
Topview of WWTP West
Available space for struvite recovery
Airlift reactor with gravity settling
Airlift reactor with a hydrocyclone
Fluidized bed reactor/airlift reactor
Stirred reactor with tilted plate settling
Fitting of a struvite recovery installation at WWTP West
Horizontal cross section of the USB
Longitudinal section of the USB
Heaps of struvite sediments in the USB
Outlet pipes in the USB
Adjustments to convert the USB into a struvite reactor/separator
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List of tables
1.1
2.1
4.1
4.2
4.3
4.4
5.1
6.1
6.2
6.3
6.4
7.1
7.2
7.3
Global phosphate rock reserves
Phosphate recovery techniques
Solid content at WWTP West
Thermodynamic equilibria
List of precipitates included in the model by Gadekar et al.
Thermodynamic equilibria with CO2
Overview of the experiments
Digested sludge characteristics
Final magnesium and orthophosphate concentrations
Testing the model
DSC increase due to struvite formation
Digested sludge discharge and characteristics
Profits from struvite separation
Profits from improved dewaterability
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List of abbreviations
ATP
DNA
ADP
BNR
WWTP
UCT
CSI
AEB
TSS
MAP
USB
CSTR
EBPR
AOB
PAO
EPS
DCB
CSD
DSC
CST
XRD
10
adenosine triphosphate
deoxyribonucleic acid
adenosine diphosphate
biological nutrient removal
wastewater treatment plant
university of Capetown
central sludge intake
afval energie bedrijf
total suspended solids
magnesium ammonium phosphate (struvite)
digested sludge buffer tank
continuous stirred tank reactor
enhanced biological phosphorus removal
ammonium oxidizing bacteria
phosphorus accumulating organisms
exocellular polymeric substances
divalent cationic briding
crystal size distribution
dry solid content
capillary suction time
X-ray diffraction
1. Introduction
This research focuses on the recovery of phosphate-rich ‘struvite’ from the digested sludge at
wastewater treatment plant (WWTP) West. This introduction describes the phosphate-related
problems that form a basis for this research, and gives a quick overview of WWTP West.
1.1 Phosphorus and phosphate
‘Phosphate’ is one of the keywords of this report. As the following paragraphs outline,
phosphate is essential to life, and wastewater treatment reveals both threats and
opportunities that relate to it. In wastewater engineering, the terms ‘phosphate’ and
‘phosphorus’ are both commonly used and often have the same meaning. This paragraph
describes the definitions as used in this work.
Phosphorus (P) is a nonmetallic chemical element discovered by the German scientist Hennig
Brand in 1669 (Sanderson, 2010). As it accounts for 0.10% of the earths crust (McMurry &
Fay, 2001) and is present in all living organisms, it is highly abundant although seldom visible.
Due to its high reactivity, it cannot be found in elementary form in nature on earth, but only
combined with other elements, like oxygen or hydrogen, into compounds called phosphates.
‘Phosphate’ refers to numerous and very different forms of phosphoric compounds, but in this
report its definition is limited to the usual forms that can be found in aqueous solutions,
including orthophosphate, polyphosphate and organic phosphate (Metcalf & Eddy, 2004).
Orthophosphate is represented by orthophosphoric acid (H3PO4), its conjugates H2PO4-,
HPO42- and PO43-, and all possible inorganic salts that can be derived from these conjugates,
such as MgPO4- and MgH2PO4+. Polyphosphates include inorganic molecules with two or more
phosphorus atoms combined into more complex structures like ATP (adenosine triphosphate).
Organic phosphate refers to phosphorus that is incorporated into organic structures, such as
DNA (deoxyribonucleic acid).
1.2 Phosphate issues
Disturbance of the global phosphate balance
Phosphate is essential for the growth of all living organisms. It plays an important role in
intercellular energy transfer, with the conversion of ADP (adenosine diphosphate) into ATP.
Next to that, it is a vital building block for DNA and, in its mineral form, for bones and teeth
(Vergouwen, 2010).
Figure 1.1 illustrates the natural cycle of phosphate, consisting of micro-cycles and a macrocycle. It should be noted that this is a strong simplification, showing only the key processes.
In its natural micro-cycle, phosphate is taken up from soils and water bodies by plants and
algae, after which it is transferred into organisms that consume these plants and algae, and
subsequently (possibly) into higher groups of organisms within the food chain. Eventually, all
phosphate will return to the soil or water, either via animal excretements or by decay of
organisms. Parallel to that, a macro-cycle can be distinguished, in which phosphate is
continuously transported from land towards inland or coastal waters by mechanical and
chemical erosion of the earths crust. During this transport, the phosphate can take part in
numerous micro-cycles. Eventually, the phosphate ends up in sediments at the bottom of
water bodies. After tens of millions of years, mountain formation processes force these
sediments to again appear at the earth’s surface as phosphate rock, thereby closing the
geological cycle (Vergouwen, 2010).
In the last few centuries, human intervention has had an enormous impact on the natural
phosphate cycle, disturbing its balance to a serious extent. Urbanization and intensification of
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stock farming led to unbalanced phosphate transport towards urban areas, causing
phosphate accumulation in these areas and, at the same time, the gradual depletion of
phosphate elsewhere. Next to that, deforestation and intensification of agriculture
significantly accelerated the natural transport of phosphate to the sea, thereby even further
depleting the farming lands phosphate reserves.
Fig. 1.1 – Key processes of the natural phosphate cycle
To overcome this unbalance, well-developed countries are currently relying on a twofold
solution. On the one hand, surplus phosphate in the form of organic solid waste, human
excreta and part of animal manure is collected and processed to end up mostly as sludge
(ash) in landfills or in non-arable soils. On the other hand, the phosphate reserves in farming
soils are replenished with fertilizers, which are mainly produced from phosphate rock.
Although the latter may seem satisfactory for well-developed areas, it causes great
disturbances in the global phosphate balance, due to extensive phosphate transports from
the few countries that have natural phosphate rock reserves (table 1.1) towards countries
without any significant reserves. Most of the time, these transports move into the same
direction as global food transports (also containing phosphate): from developing countries
towards well-developed countries. On top of that, the complex system of phosphate mining,
transport and processing causes major phosphate losses. At present, 80% of phosphate
mined is lost in fertilizer production, field application and food processing, and therefore does
not reach the food we consume (Barnard, 2009).
Table 1.1 – Global phosphate rock reserves (von Horn & Sartorius, 2009)
Country
Morocco
Jordan
China
Israel
Russia
Senegal
S. Africa
USA
Togo
Other countries
World total
Reserves (1000 tonnes)
5,700,000
900,000
500,000
180,000
150,000
150,000
100,000
100,000
30,000
1,200,000
12,000,000
Depletion of phosphate rock reserves
While phosphate fertilizers have become essential for sustaining high crop yields, all modern
agricultural systems currently rely on constant input of mined phosphate rock. However,
phosphate rock, like oil, is a finite resource (Cordell et al, 2009). Predictions as to when the
reserves will run out vary, but it is generally accepted that in less than 50 years there will be
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fewer producers, and thus possibilities of severe competition, and that the known reserves
may be depleted within around 200 years if nothing is done to recover and recycle phosphate
(Barnard, 2009).
Although social awareness of this problem is still very limited, phosphate rock depletion is
increasingly receiving attention, not only from authorities and industries, but also in regular
media.
For example, in April 2010, ‘Der Spiegel’ published the following 1 :
“While the term “peak oil” - the point at which production capacity will peak before oil wells
gradually begin to run dry – is well known, fewer people know that phosphate reserves could
also be running out. Experts refer to this scenario as “peak phosphorus”…. A phosphate crisis
would be at least as serious as an oil crisis. While oil can be replaced as a source of energy –
by nuclear, wind or solar energy -, there is no alternative to phosphorus. It is a basic element
of all life, and without it human beings, animals and plants could not survive.”
And in August 2010, ‘Dagblad De Pers’ published 2 :
“Experts already warn for upcoming scarcity that could lead to an extreme increase in price,
and that would mainly affect developing countries… The upcoming phosphate scarcity could
have great geopolitical consequences and could possibly lead to violence, if it would threaten
the food supply.”
Phosphate issues in wastewater treatment plants
Phosphate pollution in surface waters can lead to problems with eutrophication in the
receiving water such as excessive fish mortality and odor nuisance (Jaffer et al., 2002). Since
the 1990’s, stricter European regulations to reduce these problems have led to the
development of new treatment processes for removing compounds containing nitrogen and
phosphorus from the wastewater.
Implementation of these enhanced processes, especially BNR (Biological Nutrient Removal),
resulted in higher concentrations of phosphorus, nitrogen and magnesium in the excess
sludge (Doyle & Parsons, 2002). This resulted in several phosphate-related problems in
WWTPs with BNR:
- The reject stream (centrate) of the sludge line, which is fed back to the inlet of the
WWTP, is containing high concentrations of phosphorus that can lead to instabilities
and the need for additional chemical dosages in the treatment. Phosphorus feedback
in BNR plants can be responsible for 20-50% of the total phosphorus entering the
WWTP (Jaffer et al., 2002).
- Under specific conditions, the presence of high concentrations of compounds
containing phosphorus, nitrogen, magnesium and calcium will lead to the formation
of different precipitates, such as struvite and apatite crystals (Wang et al., 2009).
These precipitates can cause scaling problems in pipes (figure 1.2) and treatment
installations (Doyle & Parsons, 2002).
These disadvantages are thus negatively influencing the stability and reliability of the
treatment process and can result in a significant cost increase for operation and maintenance.
As an example, annual costs for a mid-size treatment plant (about 95,000 m3/day) related to
struvite deposits can easily exceed 100,000 US dollar (Benisch et al., 2000).
1
Source: Der Spiegel online article: Essential element becoming scarce – Experts warn of
impending phosphorus crisis, by Hilmar Schmundt, 21-04-2010.
2
Source: Dagblad De Pers: Kunstmestcrisis is pas echt schrikken – Fosfaat raakt op, kans op
nieuwe voedselcrisis, by Jan-Hein Strop, 24-08-2010 (translated from Dutch).
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Fig. 1.2 – Scaling problems in pipes caused by struvite deposits (at WWTP West)
1.3 Phosphate recovery
The issues as described in the preceding paragraph show that the removal of phosphate
containing compounds in WWTPs is highly desirable. Controlled phosphate removal from the
sludge line can lead to significant savings on operational and maintenance costs, as scaling
problems could be prevented and less phosphate is recycled to the plant inlet (to be removed
again). The recovered phosphate can be re-used as a fertilizer, either directly or after
processing by fertilizer industries. Not only would this generate extra income for WWTPs.
Since phosphate in human emissions represents more than 10% of phosphate rock
production (Barnard, 2009; Benisch et al., 2000), it could also be a promising opportunity for
closing phosphate cycles and therefore being less dependent on global phosphate rock
reserves. Besides this advantage, it was observed that a significant improvement in sludge
dewaterability can be achieved when phosphate was recovered as struvite directly from the
digested sludge by the addition of magnesium chloride (MgCl2) (Veltman et al., 2010). This
leads to smaller final sludge volumes and hence to lower transportation and disposal costs
and is therefore an important economical incentive.
1.4 Wastewater treatment plant West
As the first full water cycle company of the Netherlands, Waternet is responsible for the
production and distribution of drinking water, the collection and treatment of wastewater and
the control and maintenance of all surface waters within the city of Amsterdam and in a large
area in the provinces of Utrecht and Noord-Holland. Waternet annually treats about 125
million m3 of wastewater in 12 WWTPs with a combined capacity of 2.3 million p.e.
(population equivalent).
The newest and largest of Waternet’s WWTPs, ‘WWTP West’, was started up in 2005 and has
a total capacity of 30,000 m3/h influent. To assure sufficient capacity in the future, an
expansion of 10% (to 33,000 m3/h) has already been accounted for during construction. The
connected sewage system is a combined system, thus collecting both wastewater and
stormwater. The water is biologically treated in an activated sludge system based upon the
UCT (University of Cape Town) principle, after which the effluent is discharged onto the
surface water of the Amsterdam harbour.
The sludge line of WWTP West is the largest of the Netherlands, in processed volume per day.
WWTP West not only processes the sludge from its own treatment, but also that of the
closeby WWTP Westpoort and that of some other external sources. The external sludge is
collected in the CSI (Central Sludge Intake). The primary sludge is thickened by gravity
thickeners and the secondary sludge by belt thickeners. After that, all sludge is treated in
anaerobic digesters, where the sludge is stabilized, the sludge volume is reduced and biogas
is produced. At last, the sludge is dewatered in bowl centrifuges, after which it is transported
for incineration. Section 4.1 further explains the treatment of sludge.
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An innovative and symbiotic cooperation exists between the WWTP West and the adjacent
AEB (Afval Energie Bedrijf). The WWTP delivers about 25,000 m3 of biogas and 230 tons of
dewatered sludge per day to the AEB. The AEB uses this biogas and sludge to produce
energy. In turn, all energy that is consumed by the WWTP is delivered by the AEB, and the
surplus warmth from the AEB energy production is used for stimulating the WWTPs anaerobic
digestion process (Baeten, 2005).
Figure 1.3 displays the key processes of both the water line and the sludge line of WWTP
West, as well as its connections to WWTP Westpoort, the CSI and the AEB. Annex 1 and
Annex 2 outline the water line and sludge line separately and in more detail.
Coarse
screens
Influent
Effluent
Primary
clarifier
Coarse
material
Secondary
clarifier
Activated
sludge
Primary
sludge
Sand
Return activated sludge
Primary sludge
WWTP
Westpoort
Surplus activated sludge
Surplus
Gravity
thickener
Thickened
sludge
Belt
thickener
Primary
sludge
Supernatant
water
Surplus
Activated
sludge
Filtrate
Thickened sludge
CSI
Thickened sludge
Digested sludge
Biogas
Anaerobic
digestion
Buffer
Centrifuge
Digested sludge
AEB
Dewatered
sludge
USB (buffer)
Centrate
Fig. 1.3 – Treatment scheme of WWTP West
In figure 1.3, the points in the sludge line where scaling problems with phosphate precipitates
occur are marked red. Figure 1.4 illustrates the average daily phosphorus loads in the water
line and the sludge line, based on raw measuring data from Waternet.
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Influent
1215 kg p/day
Effluent
110 kg P/day
Water line
1640 kg P/day
Sludge
1530 kg P/day
Reject water
425 kg P/day
Sludge line
2540 kg P/day
Dewatered sludge
2115 kg P/day
External sludge
1010 kg P/day
Fig. 1.4 – Average daily phosphorus loads at WWTP West
1.5 Thesis outline
Section 2 summarizes and discusses the current state of the art on phosphate recovery from
wastewater streams. In section 3 the research problem and research questions are
formulated. The research problem and research questions are derived from the knowledge
gaps in the current state of the art, in combination with the phosphate-related problems at
WWTP West. Section 4 provides the required theoretical background information to
understand the processes and parameters influencing phosphate recovery from digested
sludge, especially in the form of struvite. Section 5 describes the set-up and procedure of the
experiments that have been conducted as a part of this research. In section 6, the results of
the experiments are presented and discussed. In section 7, attention is paid to the
possibilities of implementing a struvite recovery technique at WWTP West. The conclusions
and recommendations of this research are presented in section 8.
1. Introduction
2. State of the art
phosphate
issues
knowledge
gaps
3. Research outline
4. Theoretical
background
5. Materials & methods
6. Results & discussion
7. Implementation at
WWTP West
8. Conclusions &
recommendations
Fig. 1.5 – Thesis outline
16
2. State of the art
2.1
Classification
The previous section explained why the recovery of phosphate from wastewater streams is
highly desirable. This section gives a quick overview of the current state of the art on
phosphate recovery from wastewater. Since many techniques are very similar, not all
different techniques are included. For a complete overview, the following recent reports can
be consulted:
- Kalogo, Y.; Monteith, H. (2008): State of science report: Energy and resource
-
recovery from sludge, written for the Global Water Research Coalition.
Hermann, L. (2009): Rückwinnung von Phosphor aus der Abwassereinigung – Eine
Bestandesaufnahme, written for the Bundesamt für Umwelt (BAFU).
Phosphate recovery techniques developed for municipal wastewater treatment works can be
applied at various points in the treatment process. The following classification can be made:
- Phosphate recovery from the water line.
- Phosphate recovery from the sludge line:
o From dewatering reject streams.
o From digested sludge.
- Phosphate recovery from sludge ash.
Phosphate recovery from the water line is mostly applied at the plant effluent. In this way,
maximum control of phosphate emissions on the surface water is gained. However, this ‘endof-pipe’ solution does not reduce phosphate related problems in the treatment process.
Besides that, the phosphate concentration in the effluent is relatively low (figure 1.4), which
complicates the recovery of phosphate.
In the sludge line, the concentrations of phosphate are much higher, making efficient
recovery easier. Currently, most techniques aim at recovering phosphate from dewatering
reject streams. The low TSS (Total Suspended Solids) concentration in the reject stream
makes it relatively easy to separate phosphate precipitates from the water. However, this
method does not prevent scaling problems in the sludge line. When phosphate is recovered
from the sludge directly after anaerobic digestion, the risk at scaling problems in the
remainder of the sludge line could be significantly reduced. On top of that, the removal of
phosphate by adding magnesium can improve the dewaterability of the sludge, thereby
reducing the sludge disposal costs.
In dewatering reject streams or in the digested sludge, dissolved phosphate from the liquid
phase is recovered. This is only a small part of the total phosphate. At WWTP West, only 1020% of the phosphorus is present in the liquid phase of the digested sludge. The remainder
(80-90%) is bound in the solid phase. Both parts can be recovered independently. Recovery
of phosphate from the solid phase is mostly applied on the sludge ash (after incineration).
Recovery of phosphate from the sludge ash generally takes place at an external and central
location, together with sludge ash from other WWTPs.
Table 2.1 shows the techniques that are included in this section, labeled according to the
above-mentioned classification. A distinction is made between techniques that are fully
operational and techniques that are still in a developing phase (laboratory or pilot research).
Focus is mainly set on techniques that are most relevant to this research.
17
Table 2.1 – Phosphate recovery techniques
Technique
Company/
institute
Applied on
Developing phase
Product
Treatment principle
AirPrex
PCS
Digested sludge
Fully operational
MAP
airlift reactor
followed by sedimentation
Waterschap
Digested sludge
Velt en Vecht
Fully operational
-
aeration in a basin,
no separation
Crystalactor
DHV
rej. water/
plant effluent
Fully operational
MAP/MP
CP/KMP
fluidized bed reactor
Phospaq
Paques
rej. water/
plant effluent
Fully operational
MAP
CSTR with separatation in
a special outlet construction
Pearl
Ostara
reject water
Fully operational
MAP
fluidized bed reactor
Seaborne
Seaborne
reject water
Fully operational
MAP
CSTR followed by centrifuge
ASH DEC
ASH DEC
sludge ash
Developing
P-rich granules
chemical/thermal treatment
of sludge ash
Developing
MAP
CSTR with separation
in a hydrocyclone
Developing
CaPO4
CSTR followed by sedimentation
SEPHOS
2.2
Ebara
Environmental Digested sludge
Engineering
Ruhrverband
sludge ash
Fully operational techniques
PCS AirPrex
The AirPrex technology has been developed by the ‘Berliner Wasserbetriebe’ after massive
incrustations were found in the centrifuges of some WWTPs that proved to consist mainly of
struvite with small portions of different calcium phosphate compounds (Heinzmann & Engel,
2006). In the AirPrex technology, the digested sludge is led through a so-called ‘airlift reactor’,
in which air is being used to create internal recycle flows (figure 2.1). Ammonium ions (NH4+)
and phosphate ions (PO43-) are present in sufficient concentrations in the digested sludge,
and magnesium ions (Mg2+) are added as magnesium chloride (MgCl2) to the reactor. Struvite
is formed according to the following equation:
Mg 2+ (aq ) + NH 4+ ( aq) + PO43− (aq) + 6 H 2O(l ) → MgNH 4 PO4 • 6 H 2O ( s )
(1)
Air is applied for two reasons. Firstly, it increases the pH by stripping CO2 from the digested
sludge. As pH increases, struvite solubility decreases, as will be explained in section 4.2.
Secondly, the internal recycle flows, resulting from the air supply, allow the struvite crystals
to grow, until they reach a size at which they can escape from the recycle flow and settle. In
a second tank, smaller struvite crystals are allowed to settle and are washed together with
the collected struvite from the reaction tank.
MgCl2
To dewatering
Digested sludge
Air
MAP
MAP
Fig. 2.1 – AirPrex technology
18
In 2006, the German company PCS (Pollution Control Service) obtained the licence to market
the AirPrex technology. At the moment, two full-scale plants are operational, one in Berlin
and one in Mönchengladbach. In these plants, 80-90% of the orthophosphate is removed
from the liquid phase of the digested sludge. The phosphate product, struvite, is sold to the
fertilizer industry at 40-60 euro’s/ton (2009) 3 .
Considering the resemblance in scaling problems that underlie the need for controlled
phosphate recovery, the German AirPrex technology is believed to be a promising alternative
for implementation at WWTP West. For this reason, the experiments in this research are
based on the AirPrex technology.
However, as will be explained in the rest of this report, several variations on the AirPrex
technology are imaginable, that might perform even better on struvite formation or sludge
handling.
Waterschap Velt en Vecht (waterboard)
In july 2010, Waterschap Velt en Vecht was the first Dutch waterboard to implement struvite
formation in a domestic WWTP at full-scale, to improve the dewaterability of the sludge. At
WWTP Emmen, the digested sludge is led through a basin in which CO2 is stripped by bubble
aeration and in which magnesium is added. The struvite is not removed from the sludge. 4
Since both treatment plants use different principles of phosphate removal and sludge
dewatering, the results of WWTP Emmen are of limited use for the situation at WWTP West.
Nevertheless, the practical performance of e.g. aeration and magnesium dosage at WWTP
Emmen provides information that can be of use in designing an installation at WWTP West.
DHV Crystalactor
The Crystalactor was originally developed in the early 1980’s by the Dutch consultancy and
engineering company DHV to remove calcium (hardness) from drinking water. Soon the
technology was used to remove several other components, such as phosphate and heavy
metals from process water, drinking water and wastewater streams. At present, the
Crystalactor is fully operational at numerous water treatment locations worldwide. The
Crystalactor is a fluidized-bed type crystallizer, in which the water flows in upward direction
through a cylindrical reactor that is partly filled with seed material. By maintaining the
appropriate conditions and adding reagents, the components that need to be removed
crystallize at the seed material’s surface to form ‘pellets’. As the pellets grow, they become
heavier and eventually settle at the reactor bottom, where they are retracted. As a result of
phosphate removal, the Crystalactor produces magnesium phosphate (MP), calcium
phosphate (CP), struvite or potassium magnesium phosphate (KMP, or ‘potassium struvite’),
depending on the added reagent and process settings. In domestic wastewater treatment,
the Crystallactor treats either the plant effluent or dewatering reject water streams, thereby
producing phosphate-rich pellets that are sold to the phosphate processing industry. 5
Although currently only applied on dewatering reject streams and the plant effluent, the
Crystalactor technology could possibly be adjusted to be made suitable for recovering
phosphate products directly from digested sludge.
Paques Phospaq
The Phospaq technology recovers struvite from dewatering reject streams in a CSTR
(Continuous Stirred Tank Reactor) by aeration and addition of a magnesium source, in this
3
Source: brochure ‘AirPrex-Verfahren’ by P.C.S.
Source: www.veltenvecht.nl
5
Source: brochure ‘Crystalactor’ by DHV
4
19
case magnesium oxide (MgO). A patented outlet construction prevents the struvite from
flushing away with the effluent. The struvite is harvested at the bottom of the reactor. 6
Since the outlet construction is designed specifically to separate water and struvite, this
technology is at the moment not suitable for application on more viscous streams such as
digested sludge. The use of MgO instead of (more commonly applied) MgCl2 could have a
positive effect on crystal growth, as will be made clarified section 7.4. For this reason,
studying the performance of the Phospaq technology on crystal growth could help Waternet
in selecting the most suitable reagent.
Ostara Pearl
The Pearl technology was developed by the American company Ostara and was first
successfully operated in a pilot plant in Edmonton in 2006. Currently, four full-scale plants
have been implemented in the USA and several pilot projects are running worldwide. The
Pearl technology is very similar to the DHV Crystalactor, recovering struvite from wastewater
reject streams in a fluidized-bed reactor. To further improve the performance of the process,
phosphate is stripped from the activated sludge (before digestion) in an anaerobic zone and
added to the reject water: the so-called WASSTRIP process. This partially prevents scaling
problems in the remainder of the sludge line. The produced struvite is sold as ‘Crystal Green’
to the fertilizer industry. Ostara claims that the profits from selling the struvite are fully
covering the operational costs of a Pearl installation. Next to that, a reduction in sludge line
maintenance should guarantee a payback period of 3 to 10 years. 7
Seaborne
The Seaborne process was developed by the Seaborne Environmental Research Laboratory in
Germany and is presented schematically in figure 2.2.
Fig. 2.2 – The Seaborne process (Muller et al., 2007)
In the Seaborne process, struvite is produced from centrifuge reject water by adding
magnesium hydroxide (Mg(OH)2) as a magnesium source in one CSTR, and by adding NaOH
to raise the pH value in a second CSTR. The struvite is subsequently separated from the
water by a centrifuge. After the separation of struvite, ammonium sulfate is recovered from
6
7
Source: www.fluidsprocessing.nl
Source: a presentation by Ostara
20
the (still ammonium-rich) reject water. Ammonium sulfate can also be re-used in agriculture.
Next to that, heavy metals are being removed and the biogas from the anaerobic digester is
cleaned by the precipitation and filtration of metal sulfides, just before struvite production
(Tisza, 2001). In 2005, the first large-scale installation was started up at the WWTP of
Gifhorn in the northern part of Germany. A number of mechanical as well as procedural
problems occurred. The separation of heavy metal sulfides before struvite production causes
the major problem. Being of colloidal size, the heavy metal sulfides are very difficult to
remove. Insufficient removal, however, leads to unacceptably high heavy metal
concentrations in the produced struvite for direct re-use (Tisza, 2001).
2.3
Developing techniques
ASH DEC
The Austrian company ASH DEC Umwelt AG developed the ASH DEC technology for
recovering phosphorus from incinerated sludge ash. This technology applies a thermochemical treatment to the dewatered sludge: A mix of ash and additives is heated to 1,000 ºC
in a reactor. The heavy metals are converted to the gaseous state, leave the reactor and are
captured in a gas cleaning system. Simultaneously, the phosphorus reacts with non-volatile
additives to form phosphorus-rich granules that are mechanically separated and sold to the
fertilizer industry. ASH DEC participated in the recently completed SUSAN project that
investigated possibilities to recover phosphorus from sludge ash. Currently, pilot installations
in Austria and Germany are running. The first full-scale installation is planned to be
operational in 2012. 8
Ebara Environmental Engineering
In 2008, Shimamura et al. published a paper that describes experiments in which phosphate
is recovered directly from digested sludge in a 50 m3/day facility (Shimamura et al., 2008). In
these experiments raw sludge was fed to a CSTR. Magnesium hydroxide was added as a
magnesium source and sulphuric acid (H2SO4) was dosed to adjust the pH. The
orthophosphate (PO4-P) concentration was reduced from 268 mg/L to 20 mg/L and the
crystal sizes were between 0.21 and 0.24 mm. They concluded that in a 21,000 m3/day
WWTP with Enhanced Biological Phosphorus Removal (EBPR) and anaerobic digestion 315
kg/day of struvite can be recovered. This technique differs from AirPrex in two ways. The
sludge is mixed by stirring rather than aeration, and another magnesium source (magnesium
hydroxide instead of magnesium chloride) is used.
Sephos
The Sephos process is being developed by the institute WAR of the TU Darmstadt and the
Ruhrverband in Germany. In the Sephos process, sludge ash is acidified to pH=1.5 to release
phosphorus and heavy metals. Residuals (mostly sand) are removed in a hydrocylone.
Subsequently, the pH is raised to 3.5 to precipitate aluminum phosphate (AlPO4), which is
removed in a second hydrocylone. The aluminum phosphate can be used as a raw material in
the phosphate processing industry, or can be processed further to calcium phosphate (CaPO4-)
in the ‘advanced Sephos process’ (figure 2.3). In the advanced Sephos process, the aluminum
phosphate is dissolved by raising the pH value to 12-14. The (insoluble) heavy metals are
removed and calcium is added to precipitate with the phosphate. The resulting product,
calcium phosphate, can be used directly in agriculture (Berg & Schaum, 2005).
8
Source: www.phosphorus-recovery.tu-darmstadt.de
21
Figure 2.3 – The Sephos process (Berg & Schaum, 2005)
2.4
General state of development
Over the past few years, phosphate recovery from wastewater streams has received more
attention and several different techniques have been developed. The previous paragraphs
showed that several companies are currently successfully exploiting their technologies to
recover phosphate from dewatering reject water, digested sludge or sludge ash. Since
phosphate recovery from the liquid phase (e.g. AirPrex, Pearl and Phospaq) can be applied
independently of phosphate recovery from the solid phase (e.g. ASH DEC and Sephos), both
methods should be further researched and developed in parallel tracks.
Remarkably, phosphate recovery from the liquid phase is mostly applied on dewatering reject
streams, while phosphate recovery directly from the digested sludge has the (very attractive)
advantages of reducing scaling problems in the sludge line and improving sludge
dewaterability. Further research on phosphate recovery from digested sludge and especially
the impact on the sludge dewaterability could perhaps change this. Besides that, it should be
made clear to researchers, waterboards and engineering companies that:
- It is very well possible to recover a phosphate product from digested sludge that
does not exceed the legal restrictions for hazardous components and that is valuable
as a fertilizer.
- The recovery of phosphate from the liquid phase of digested sludge does not
interfere with the recovery of phosphate for the solid phase (as sludge ash).
At present, Dutch legislation does not allow the application of any products from municipal
wastewater treatment works in agriculture 9 . This means that companies are forced to export
any phosphate product that is recovered from wastewater streams to countries where
application is allowed, such as Germany. This hinders the development and implementation of
phosphate recovery techniques in the Netherlands. Efforts should be made to convince the
government to change the law in such a way that the recovered phosphate can be used in
the Netherlands as soon as possible. Of course, careful restrictions have to be maintained.
9
Source: www.wetten.overheid.nl (Koninklijk Besluit gebruik meststoffen)
22
3. Research outline
3.1
Research problem
Over the past few years, phosphate precipitates have caused scaling problems in pumps,
pipes and dewatering facilities of the sludge line of the WWTP West. During an extensive
cleansing operation in 2009, it was discovered that approximately 150 tons of struvite had
formed in the USB (digested sludge buffer tank, figure 1.3) during four years of regular plant
operation. The controlled crystallization and separation of struvite, or Magnesium Ammonium
Phosphate (MAP), is one of the most widely recommended technologies for dealing with the
abovementioned scaling problems, especially in WWTPs with EBPR, such as WWTP West
(Martí et al., 2010).
PCS claims that the AirPrex process (section 2.2) can effectively prevent the scaling problems
at WWTP West, while significantly improving the dewaterability of the digested sludge and
producing a precipitation product (struvite) that can be sold as a fertilizer. Waternet is
interested in this technique and recently conducted a small-scale research using a pilot
(AirPrex) reactor. The results of the pilot research are found to be promising, in both
reducing the phosphate concentration and improving the sludge dewatering. Annex 3
summarizes the pilot results.
Although full-scale implementation of the AirPrex process at the WWTP West seems sensible
based on current knowledge and the pilot results, additional research is needed to optimize
the parameters that influence both struvite formation in digested sludge and the sludge
dewaterability. Since every WWTP has its own unique sludge composition and characteristics,
it will pay off determining optimal process settings for the specific sludge conditions at WWTP
West in laboratory experiments, before making a full-scale design. The majority of published
reports on struvite recovery focus on the controlling parameters for struvite recovery from
relatively watery streams such as centrifuge reject water (centrate), rather than directly from
the digested sludge. Next to that, no published reports were found that discuss struvite
recovery from digested sludge in relation to the dewaterability of the treated sludge. Since
improving sludge dewaterability is an important economical incentive, it should be seen as
one of the most important parameters in struvite recovery from digested sludge and it should
be investigated in more detail.
Concluding, the research problem for this thesis project is stated as follows:
“Literature and previous experiments do not provide a sufficient basis for making an optimal
design for a full-scale struvite recovery reactor at WWTP West, due to the unique character of
the digested sludge and the lack of available research on the relation between struvite
recovery and dewaterability of the digested sludge.”
3.2
Research approach
Based on the research problem as stated above, the following research questions are
formulated:
Main question:
“What are the optimal process settings for recovering struvite by the addition of MgCl2 from
digested sludge at WWTP West, while optimizing sludge dewatering?”
23
Sub-questions:
- What are the main mechanisms and what are the most important parameters in
struvite formation and sludge dewaterability?
- What are the digested sludge conditions at WWTP West?
- What are the process settings for reaching optimal sludge dewaterability?
- What are the process settings for reaching optimal phosphate removal?
- What are the process settings for reaching optimal struvite recovery?
In answering these questions, several research methods are used:
- Literature study to investigate and describe the working principles of struvite
formation and separation.
- Experiments to gain data on struvite formation and sludge dewaterability.
- Quantitative data analysis to analyze the data gained by the experiments, as well as
existing data, in finding optimal process settings.
- Modeling as a tool in finding optimal process settings.
24
4. Theoretical background
4.1
Sludge
‘Wastewater sludge’ is the semisolid, nutrient-rich by-product that occurs when wastewater is
treated to be returned into the environment. Although sludge represents only a few percent
of the volume of processed wastewater, processing it stands for up to 50% of total operating
costs (Turovskiy & Mathai, 2006; Spinosa, 2007; von Sperling, 2007). Therefore sludgeprocessing strategies are of great economical relevance in wastewater treatment. Several
types of wastewater sludge can be distinguished, dependent on the WWTPs treatment
principal and the location within the treatment process.
Primary sludge
Primary treatment is the first step in wastewater processing, in which readily settable solids
are removed from the wastewater by sedimentation. The sludge product of this treatment
step, ‘primary sludge’ is usually a gray and slimy substance that, in most cases, has an
extremely offensive odor. The composition of primary sludge is to a large extent dependent
on the catchment area characteristics and, in the case of a combined sewer system, may
show strong seasonal variations. Primary sludge consists for a large part of organic matter
such as feces, paper, leafs and wasted food, and typically contains around 6.5% of solids by
weight (Metcalf & Eddy, 2004).
Secondary sludge
Secondary sludge, also known as biological sludge, is the sludge byproduct of biological
treatment processes such as the activated sludge process, membrane bioreactors and
trickling filters (Turovskiy & Mathai, 2006). Biological treatment is mainly used to (Metcalf &
Eddy, 2004):
- Convert dissolved and particulate biodegradable constituents into acceptable end
products.
- Capture and incorporate suspended and non-settleable colloidal solids into a
biological floc or film.
- Transform or remove nutrients such as ammonium and phosphate.
Further explanation is limited to the activated sludge process, as this is the biological
treatment process implemented at WWTP West.
The activated sludge process was developed in the UK in 1913-1914. It was discovered that a
highly treated effluent can be obtained using a draw-and-fill reactor: adding raw wastewater
to previously settled sludge, aerating this mixture for several hours, letting the sludge settle,
carefully removing the treated supernatant, and leaving (part of) the sludge for treatment of
the next batch of raw wastewater (Henze et al., 2008; Weismann et al., 2007). This basic
activated sludge process can also be operated in continuous mode, as shown in figure 4.1.
The activated sludge process, as all biological treatment processes, relies on the natural
metabolism of microorganisms (mainly bacteria) that live inside the sludge(Metcalf & Eddy,
2004). Different species of microorganisms have different chemical processes for e.g. cell
growth and energy supply, thereby converting or taking up different constituents that are to
be removed from the wastewater. By engineering preferable environmental circumstances for
certain types of microorganisms, biological treatment processes can be designed to effectively
remove specific wastewater constituents.
25
Fig. 4.1 – Basic activated sludge process
Conventional (basic) activated sludge processes as shown in figure 4.1 principally aim to
remove organic carbonaceous material (Seviour et al., 2003). In these processes, aerobic 10
heterotrophic 11 microorganisms are used, which can readily take up some organic molecules
and which can take up more complex organic molecules after breaking them down with
enzymes (Metcalf & Eddy, 2004). Effective removal of ammonium and phosphate requires
combinations of other types of microorganisms, and thus combinations of other
environmental circumstances.
Ammonium is traditionally removed from wastewater in two steps: nitrification and
denitrification (Henze et al., 2008). In the nitrification step, ammonium is converted to nitrite
(NO2-) and subsequently to nitrate (NO3-) by Ammonia Oxidizing Bacteria (AOB) (Henze et al.,
2002). In the denitrification step, nitrate is converted into atmospheric nitrogen by (mainly)
facultative aerobic 12 bacteria. Since the facultative aerobic bacteria prefer oxygen as an
electron acceptor, it is important that none or very little oxygen is present (anoxic
circumstances) when denitrification is desired (Henze et al., 2002). Therefore, in biological
nitrogen removal, an aerobic and an anoxic zone can usually be distinguished that are either
incorporated in the activated sludge process or that are implemented elsewhere in the
treatment process. It should be noted that in the last decade several new innovative
techniques for biological nitrogen removal have been developed, that rely on different types
of bacteria and do not always need a separate aerobic and anoxic zone (Henze et al., 2008).
As a typical bacteria cell contains about 2% of phosphorus (Metcalf & Eddy, 2004), all
activated sludge processes remove phosphate to some extent by wasting of the surplus
sludge. However, larger amounts of phosphate can be removed when so-called Phosphorus
Accumulating Organisms (PAOs) are used. PAOs are able to take up phosphate under aerobic
conditions and store it in granules within their cells as energy-rich polyphosphates, often
referred to as ‘luxury-uptake’, along with magnesium, calcium (Ca2+) and potassium (K+)
cations. Under anaerobic conditions, phosphate and these cations are released (Metcalf &
Eddy, 2004). Since the discovery of PAOs in the 1970’s (Henze et al., 2008), a series of
activated sludge processes with biological phosphate removal has been developed, referred
to as Enhanced Biological Phosphorus Removal (EBPR). At WWTP West, the UCT-principle
(University of Cape Town) is applied. The basic UCT principle is presented schematically in
figure 4.2.
10
Aerobic microorganisms use oxygen as the electron acceptor in redox-reactions for their
energy supply (Metcalf & Eddy, 2004).
11
Heterotrophic microorganisms obtain their carbon for cell growth from organic matter
(Metcalf & Eddy, 2004).
12
Facultative aerobic microorganisms use either oxygen or nitrate/nitrite as the electron
acceptor in redox-reactions for their energy supply (Metcalf & Eddy, 2004).
26
Fig. 4.2 – The UCT-principle
In the UCT-principle, the passage of nitrate (NO3-) into the anaerobic zone is avoided by
controlling the recycle stream from the aerobic zone to the anoxic zone in such a way that
the available nitrate amount is always smaller than the denitrification capacity in the anoxic
zone (van Haandel en van der Lubbe, 2007). An anaerobic zone without the presence of
nitrate is required in front of the process, since denitrifying bacteria inhibit the uptake of
organic substrate 13 by PAO’s, needed for growth (Seviour et al., 2003).
The composition of the (surplus) activated sludge strongly depends on the type of process
that is applied. In WWTPs with EBPR, the sludge contains relatively large concentrations of
phosphate, magnesium, calcium and potassium. Activated sludge generally has a brown
flocculent appearance and an inoffensive “earthy” odor. It typically contains 0.8 to 1.2% of
dry solids by weight (Metcalf & Eddy, 2004).
Chemical sludge
Chemicals are widely used in wastewater treatment to precipitate hard-to-remove substances
and to improve the removal of suspended solids, thereby producing chemical sludge.
Chemical removal of phosphate is often applied, either as an alternative for, or as an
extension of EBPR. The chemical precipitation can take place in the primary clarifier, in a
separate reactor, or can be incorporated within the activated sludge process (Turovskiy &
Mathai, 2006). At WWTP West, ferric chloride (FeCl3) is dosed just before the activated
sludge process, to precipitate phosphate according to reaction equations (2-3) (Metcalf &
Eddy, 2004), as an extension of EBPR. The chemical sludge is discharged together with the
surplus activated sludge.
FeCl3 ( s) → Fe3+ (aq) + 3Cl − (aq )
3+
3−
4
(2)
+
Fe (aq ) + H n PO (aq ) → FePO4 ( s ) + nH (aq )
(3)
Of course, the characteristics of the chemical sludge strongly depend on the precipitation
agent that is used and on the wastewater constituent that is removed. Isolated sludge from
chemical precipitation with metal salts is usually dark in color, though its surface may be red
if it contains much iron (Metcalf & Eddy, 2004).
Sludge thickening
Sludge thickening is applied on primary, secondary and chemical sludge for increasing the
concentration of dry solids (hence reducing the volume) by removing part of the liquid phase.
This volume reduction is beneficial to subsequent treatment processes such as digestion and
dewatering (Metcalf & Eddy, 2004).
13
Substrate: the carbonaceous organic matter that is converted during biological treatment
(Metcalf & Eddy, 2004).
27
Four types of water can be distinguished regarding wastewater sludge: free water, absorbed
water, capillary water and cellular water. Free water is not significantly bound to the solid
phase and can simply be removed from the solid state by gravitational forces, as happens in
gravity thickeners and gravity-belt thickeners. Absorbed water is bound directly at the solids
surface and can only be removed by chemical or thermal processes, just like capillary water,
which is captured inside pores by capillary forces. Cellular water is part of the solid phase as
it is captured inside sludge cells. It can only be removed through thermal forces that lead to a
change in the state of aggregation of the water (von Sperling, 2007). Generally, secondary
sludge is harder to dewater than primary sludge and chemical sludge because of the
relatively large content of cellular water and the relatively low content of discrete particles.
Several thickening techniques are available. Gravity thickening is commonly applied on
untreated primary sludge and gravity-belt thickening is commonly used to thicken surplus
activated sludge, as is the case at WWTP West. Gravity thickening is usually performed in
circular tanks. The diluted sludge enters through a center feed well in the upper part of the
tank and the thickened sludge is collected at the bottom. The liquid phase flows over weirs at
the edges of the tank and is recycled to the plant inlet. Deep trusses or vertical pickets stir
the sludge gently, thereby opening up channels for water to escape and promoting
densification (Metcalf & Eddy, 2004). In gravity-belt thickeners, the diluted sludge is
distributed on a rolling semi-permeable belt, through which the liquid phase can escape. After
removing the thickened sludge, the belt is continuously washed to avoid clogging.
Digestion
The raw sludge that is retracted from the treatment process (primary sludge, secondary
sludge, chemical sludge) is rich in pathogens, easily putrescible and has an offensive odor. To
make the sludge suitable for disposal, sludge stabilization is needed, either by biological,
chemical or thermal processes. Survival of pathogens, release of odors and putrefraction
occur when microorganisms are allowed to flourish in the organic fraction of the sludge.
Therefore all stabilization processes aim to reduce the organic (volatile) content or to create
conditions in which microorganisms cannot survive (Metcalf & Eddy, 2004). Biological
stabilization is the most widely used approach (von Sperling, 2007) and is achieved in aerobic
or anaerobic digestion processes, converting the raw sludge into ‘digested sludge’.
Aerobic digestion can be seen as a prolonged phase of the activated sludge process, in which
the microorganisms begin to consume their own protoplasm 14 to obtain energy for cell
maintenance reactions, as the supply of available substrate is depleted (Metcalf & Eddy,
2004). This process, known as the endogenous phase, the biodegradable cell mass (75 to
80%) is aerobically oxidized to carbon dioxide, ammonia and water. The ammonia is
subsequently oxidized to nitrate (von Sperling, 2007). While aerobic digestion is generally
implemented at small WWTPs (less than 20,000 m3/day), larger plants, such as WWTP West,
mostly rely on anaerobic digestion for sludge stabilization Turovskiy & Mathai, 2006).
In anaerobic digestion, the sludge is oxidized in three basic steps: hydrolysis, fermentation
and methanogenesis. In the first step, particulate material is converted to simple soluble
compounds that can be used by bacteria that perform fermentation. During fermentation, the
compounds are further degraded to hydrogen, carbon dioxide and acetate 15 . In the final step
(methanogenesis) a group of organisms called methanogens produce methane (CH4) and
carbon dioxide from the fermentation products, either by splitting the acetate or by a reaction
between hydrogen and carbon dioxide (Metcalf & Eddy, 2004). These processes are usually
operated in the mesophilic temperature range (30 to 35ºC) hence warmth has to be added.
At the WWTP West, this warmth is supplied by the AEB, that in turn uses methane, collected
from the anaerobic digesters, to produce energy (section 1.4).
14
15
28
Protoplasm: the cytoplasm and nucleus of a cell (Metcalf & Eddy, 2004).
Acetate: either a salt or an ester of acetic acid (CH3COOH).
Digestion not only reduces sludge volumes, it also has a major impact on the sludge
characteristics. Within the scope of this report, three changes in anaerobically digested sludge
are of particular importance:
- Phosphate, magnesium, calcium and potassium are released by the PAOs into the
liquid phase of the sludge, due to the anaerobic conditions in the digester.
- The liquid phase of the sludge is saturated with dissolved CO2 that is produced during
fermentation and methanogenesis.
- Since many organic structures are broken down during digestion, suspended matter
in digested sludge has relatively small dimensions and is less settleable than the
suspended matter in raw sludge.
Anaerobically digested sludge is typically an oil-like dark-brown to black substance containing
an exceptional large quantity of gas. It has an inoffensive odor, when thoroughly digested
(Metcalf & Eddy, 2004). Section 6.1 discusses the characteristics of the digested sludge at
WWTP West in more detail.
Sludge dewatering
In the final step of the sludge treatment line, the (digested) sludge is dewatered to reduce its
volume and to make it suitable for sludge disposal purposes. Sludge dewatering can be seen
as an advanced from of sludge thickening, in which the liquid content is even further reduced.
Table 4.1 gives an overview of the solid content at different places in the sludge line of
WWTP West.
Table 4.1 – Solid content at WWTP West
Sludge type
Primary sludge
Thickened primary sludge
Surplus activated sludge
Thickened surplus activated sludge
Digested sludge
Dewatered sludge
Dry solid content
(% by weight)
1.5
5.0
0.8
6.5
3.5
22.0
At WWTP West, commonly applied solid-bowl centrifuges are used to dewater the digested
sludge. In these centrifuges, digested sludge is fed at a constant rate, where it separates into
a dense cake containing the solids and a dilute called ‘centrate’ or ‘reject water’ (Metcalf &
Eddy, 2004). Figure 4.3 and 4.4 show the centrifuges and the dewatered sludge at WWTP
West, respectively.
Fig. 4.3 – Dewatering centrifuges at WWTP West
Fig. 4.4 – Dewatered sludge
Dewaterability of the digested sludge is an important parameter, because the final sludge
volumes that leave a WWTP have a significant influence on operational costs. The easier the
sludge can be dewatered, the lower transportation and disposal costs will be. There is no
consensus among scientists yet regarding the specific mechanisms and parameters
29
determining sludge dewaterability. The next part discusses three aspects that are (possibly)
related to dewaterability:
- The concentration of monovalent and divalent cations.
- The addition of polymers.
- The concentration of orthophosphate.
The concentration of monovalent and divalent cations.
Microorganisms produce biopolymers that are released to the exocellular environment, known
as exocellular polymeric substances (EPS). The biopolymers form a matrix in which the
microorganisms are encapsulated, stimulating floc formation and hence positively influencing
the sludge dewaterability. As the bioflocs are generally negatively charged, cations (positively
charged ions) have proven to be important in sludge dewaterability (Sobeck & Higgins, 2005).
Literature provides three different theories to explain the mechanisms by which cations
influence bioflocculation. In 2005, Sobeck and Higgins examined these three theories and
concluded that the Divalent Cation Bridging (DCB) theory explains the role of cations best.
According to the DCB theory, divalent cations, such as Mg2+ and Ca2+, bridge the negatively
charged groups present on the EPS. This bridging helps to aggregate and destabilize the
matrix of EPS and microorganisms (mostly bacteria), thereby promoting floc formation. Next
to that, high concentrations of monovalent cations, such as Na+, have been shown to cause a
deterioration of sludge dewaterability. The DCB suggests that this deterioration is caused by a
loss in cation bridging, since the monovalent ions occupy the negatively charged groups on
the EPS (Peeters & Herman, 2007). The DCB theory is illustrated in figure 4.5. It should be
stressed that the DCB theory is not generally accepted, since other theories (the Alginate
theory and the Double Layer (DLVO) theory) have not (yet) been invalidated.
Weak flocculation due to a lack of divalent cations
Strong flocculation
Na+
-
Na+
-
+
-
Na
Mg2+
Na+
Na+
-
-
Mg2+
-
Na+
-
-
-
Mg
-
Na
Na+
-
+
Na+
-
Na+
Mg2+
Bacteria
-
2+
Mg2+
-
-
Negatively
charged EPS
Fig. 4.5 – Divalent Cation Bridging theory (Peeters & Herman, 2007)
The addition of polymers.
To improve sludge dewaterability, additional polymers (either synthetic or natural) are dosed
in front of dewatering installations. Depending on the type of polymer, different working
principles can be distinguished. Cationic polymers are used as coagulants to neutralize or
lower the charge of (mostly negatively charged) wastewater particles, by absorbing at the
particle surface. At a lower charge, particle growth will easier occur as a result of particle
collision, making the particles easier to remove (Metcalf & Eddy, 2004). Anionic polymers are
used to form bridges between separate particles, with multivalent cations acting as links
between polymer and particle (Henze et al., 2002). Subsequently, bridged particles merge
with other bridged particles to three-dimensional structures that can be easily removed
30
-
(Metcalf & Eddy, 2004). Also nonionic polymers can be used for bridge formation, since polar
groups are present in the polymer chain where a positive and negative charge is found
around certain atoms (Henze et al., 2002).
The concentration of orthophosphate
PCS, the company that exploids the AirPrex process, states that high orthophosphate
concentrations stabilize the water absorbing colloid system that is present in the sludge. It is
thus suggested that lowering the orthophosphate concentration contributes to a better
dewaterability by partly destabilitzing this colloid system. No other sources were found that
support this theory, neither in published literature nor on the Internet.
Sludge disposal
After processing, the final sludge product can either be incinerated, be used in landfills, or be
used as biosolids in agriculture. The choice for a suitable final destination, and thus a suitable
processing strategy, is strongly dependent on the local socio-economic context and on
environmental factors. Strict regulations and directives for sludge disposal exist in most
countries, due to the hazardous compounds (such as heavy metals and pathogens) that
retain in most sludge products. The major part of the processed sludge in the Netherlands is
incinerated. The final sludge product of the WWTP West is incinerated at the AEB (section
1.4).
4.2
Struvite crystallization
Struvite
Struvite (chemical formula: MgNH4PO4•6H2O) is a phosphate mineral that was first described
as a component of urinary stones and guano in the late 18th century. Its occurrence in
wastewater treatment works was first reported by Rawn et al. (1939), who identified it in
hard crystalline deposits in the supernatant lines of a multiple stage sludge digestion system
(Cervantes, 2009). Struvite has a distinctive orthorhombic structure (Cervantes, 2009), which
means that at micro-scale its crystal surfaces are positioned perpendicular to each other,
while differing in size (Mullin, 2001). Depending on the process conditions during formation,
its appearance may vary in size, shape and transparency. Other precipitates, like calcium
phosphates, can compete with struvite formation and can be incorporated as impurities within
stuvite crystal aggregates (Hao et al., 2009). The main process parameters determining
struvite formation include pH value, temperature, ionic strength, dissolved component
concentrations (magnesium, ammonium, phosphate) and the presence of other compounds
(Cervantes, 2009).
If properly recovered, struvite can be used as a slow release 16 fertilizer. Several researchers
that focused on the usability of struvite as a fertilizer have investigated the presence of
contaminants. In most cases it was concluded that the hazardous components were far below
legal restrictions, even when the struvite was recovered directly from digested sludge liquors
(Heinzmann & Engel, 2006). In 2009, Weinfurtner et al. found some struvite products that
cannot be used directly as a fertilizer in Germany, because they exceed restriction values for
copper and nickel.
Crystallization
Crystallization is a transformation process in which a solute is separated from a solution
(mother liquor) through the creation of a solid phase. It is widely used in e.g. chemical,
pharmaceutical and food industries. Crystallization differs from ‘regular’ precipitation as
relatively pure crystal structures are formed, instead of flock precipitates. The driving force
16
Slow release fertilizers are characterized by a gradual release of nutrients (Sonneveld &
Voogt, 2009).
31
for crystallization is supersaturation, which can be defined as the difference between the
chemical potential of the solute in a solution and the chemical potential under equilibrium
conditions (Letcher, 2004). Under supersaturated conditions, solid crystals are formed
(nucleation) and increase in size (growth) until all of the available supersaturation has been
depleted (Hartel, 2001). The degree of supersaturation strongly determines the relation
between nucleation and growth, and therefore the crystal size distribution (CSD) of the
crystallized product.
Solubility and saturation
In multi-component systems such as struvite crystallization (magnesium, ammonium and
phosphate), whether or not supersaturated conditions are met depends on the
thermodynamic state parameters temperature, pressure and availability of the components
for reaction (Tisza, 2001). All current struvite recovery techniques (section 2.1) aim at
changing the availability of the components, rather than changing the pressure or
temperature in the solution, since the latter is relatively complex and expensive. In this report
the supersaturation is described as a function of the free component concentrations and the
ionic strength in the solution, based on a series of articles by Ali et al. (Ali et al., 2005; Ali,
2007; Ali & Schneider, 2008).
Under supersaturated conditions, struvite is formed by a chemical reaction between free Mg2+,
NH4+ and PO43- ions, with the incorporation of 6 H2O molecules (as presented in equation (1)).
In solution, magnesium, ammonium and phosphate are present in different forms. For the
calculation of struvite solubility, in most cases the following dissolved ionic species are
considered (Cervantes, 2009): H3PO4, H2PO4-, HPO42-, PO43-, MgH2PO4+, MgHPO4, MgPO4-,
MgOH+, Mg2+, NH4+ and NH3. The total concentrations of dissolved magnesium (CT, Mg),
ammonium (CT, NH4) and phosphate (CT, PO4) can be expressed as follows:
CT , Mg = ⎡⎣ Mg 2+ ⎤⎦ + ⎡⎣ MgOH + ⎤⎦ + ⎡⎣ MgH 2 PO4+ ⎤⎦ + [ MgHPO4 ] + ⎡⎣ MgPO4− ⎤⎦
(4)
CT , NH 4 = [ NH 3 ] + [ NH 4 ]
(5)
CT , PO 4 = [ H 3 PO4 ] + ⎡⎣ H 2 PO4− ⎤⎦ + ⎡⎣ HPO42− ⎤⎦ + ⎡⎣ PO43− ⎤⎦
+ ⎡⎣ MgH 2 PO4+ ⎤⎦ + [ MgHPO4 ] + ⎡⎣ MgPO4− ⎤⎦
(6)
These species are present according to the thermodynamic equilibria as presented in table
4.2, where {i} represents the ion activity of the corresponding specie.
32
Table 4.2 – Thermodynamic equilibria
Equilibrium equation
{Mg } ⋅{OH } = K
{MgOH }
{Mg } ⋅ {PO } = K
{MgPO }
2+
Value Ki
−
+
2+
MgOH
10-2.56
MgPO 4
10-4.80
3−
4
−
4
{Mg } ⋅{HPO } = K
2+
{MgHPO4 }
2−
4
{Mg } ⋅ {H PO } = K
{MgH PO }
{H } ⋅{PO } = K
{HPO }
{H } ⋅{HPO } = K
{H PO }
2+
10-2.91
MgH 2 PO 4
10-1.51
−
4
2
+
4
2
MgHPO 4
+
3−
4
2−
4
+
HPO 4
10-12.35
H 2 PO 4
10-7.20
2−
4
−
4
2
{H } ⋅{H PO } = K
+
2
{H 3 PO4 }
−
4
{H } ⋅{ NH } = K
{NH }
H 3 PO 4
10-2.15
NH 4
10-9.25
+
3
+
4
{H } ⋅ {OH } = K
+
−
W
10-14.00
The ion activity {i} is related to the ion concentration [i] according to equations (7-9), where
I is the ionic strength of the solution, Zi the valence of the corresponding specie, and γi the
activity coefficient of the corresponding specie. The DeBye-Hückel constant, A, has a value of
0.493, 0.499, 0.509 and 0.519 at 5, 15, 25 and 35ºC, respectively (Ali & Schneider, 2008).
{i} = [i ] ⋅ γ i
⎛
I
− Log ( γ i ) = A ⋅ Z i2 ⋅ ⎜⎜
⎝1+ I
I = 0.5 ⋅ ∑ Ci Z i2
(7)
⎞
⎟⎟ − 0.3I
⎠
(8)
(9)
The availability of free Mg2+, NH+ and PO43- is strongly dependent on the amount of H+, as
shown in table 4.2 and equations (4-6). Since pH is the negative logarithm of H+, pH is
directly related to the availability of the components and therefore it is the predominant
33
parameter for controlling struvite formation. Figure 4.6-4.8 show the distribution of total
magnesium, ammonium and phosphate over the different species as a function of pH value
for the following (constant) conditions: I=0.02mol/L, pKso=13.26, CT,Mg=25.4mg/L,
CT,NH4=900mg/L and CT,PO4=318mg/L.
Fig. 4.6 – Distribution of magnesium over several
compounds
Fig. 4.7 – Distribution of ammonium over several
compounds
Fig. 4.8 – Distribution of phosphate over several compounds
The solubility status of struvite under certain circumstances can be investigated by comparing
the conditional solubility product (Pcs) and the product of the analytical molar concentration
(Pso). Pcs relates to the solution properties, including ionization fraction (αi), activity
coefficients (γi) and the minimum struvite solubility product Kso. Pso relates to the total
concentrations of reactive constituents (Ali & Schneider, 2008).
α Mg
⎡⎣ Mg 2+ ⎦⎤
=
CT , Mg
α PO 4
34
⎡⎣ PO43− ⎦⎤
=
CT , PO 4
(10)
(11)
⎡ NH 4 ⎤⎦
α NH 4 = ⎣
+
CT , NH 4
Pcs =
K so
α PO 4 ⋅ α NH 4 ⋅ α Mg ⋅ γ PO 4 ⋅ γ NH 4 ⋅ γ Mg
Pso = CT , PO 4 ⋅ CT , NH 4 ⋅ CT , Mg
(12)
(13)
(14)
For Pso>Pcs, the solution is supersaturated. The degree of supersaturation can also be
expressed as the dimensionless supersaturation ratio Sc or relative supersaturation Sr:
1/ 3
⎛P ⎞
Sc = ⎜ so ⎟
⎝ Pcs ⎠
(15)
S r = Sc − 1
(16)
The solution is supersaturated for Sc > 1, or Sr > 0. Figure 4.9 shows the relative
supersaturation curve under the same conditions as figure 4.6-4.8: I=0.02mol/L, pKso=13.26,
CT,Mg=25.4mg/L, CT,NH4=900mg/L and CT,PO4=318mg/L. Varying these conditions will change
the saturation state. Figure 4.10 shows the relative supersaturation curves for a narrow pH
range, varying the ionic strength while keeping the other conditions at the abovementioned
values.
Fig. 4.9 – Relative supersaturation
as a function of pH
Fig. 4.10 – The sensitivity of supersaturation to ionic
strength variations
In the range that was studied, higher ionic strength leads to lower ionic activity (equations
(8-9)) and therefore to lower supersaturation. The value used here (I=0.02mol/L) has been
adapted from example calculations in digested sludge by Metcalf & Eddy (Metcalf & Eddy,
2004). Larger total concentrations of the reaction constituents will lead to larger Pso values
and therefore to higher supersaturation. The values used here are based on average values
found in the experiments (table 6.1). On the other hand, larger pKso values (smaller Kso
values) will lead to smaller Pcs values and therefore to higher supersaturation. The pKso value
used in the figures above (pKso=13.26) has been adapted from literature (Ali et al., 2005).
35
Nucleation & growth
The relation between nucleation and growth strongly determines the CSD of the formed
crystals and therefore is of main importance in the design of any crystallizer. Different
mechanisms of nucleation that occur can be generally divided into ‘primary’ and ‘secondary’
nucleation.
In primary nucleation no existing crystals are involved. New crystals (nuclei) will either form
spontaneously from the supersaturated solution (homogeneous nucleation) or at the surface
of any foreign objects in the solution (heterogeneous nucleation). In any supersaturated
solution, both the formation of instable clusters from solute molecules and the re-dissolution
of these clusters will take place in a continuous process. Only after a cluster has reached a
critical size (rc), this cluster will stabilize (homogeneous nucleation) and will act as a nucleus
upon which further growth is allowed (de Haan & Bosch, 2007). Homogeneous nucleation is
rare and only plays a role in extremely pure solutions or with very high supersaturation (van
Rosmalen, 1994). In heterogeneous nucleation, crystals are formed around foreign objects
such as dust particles, impeller blades or vessel walls. Since a significantly lower
supersaturation is needed, heterogeneous nucleation will normally outweigh homogeneous
nucleation (de Haan & Bosch, 2007).
Secondary nucleation is the formation of new nuclei due to the presence of solute crystals.
There are different mechanisms of secondary nucleation that can be generally divided into
four categories (van Rosmalen, 1994):
- ‘Initial breeding’ in which small fragments adhering to the crystals surface are
washed off to form new nuclei. This usually occurs when dry crystals are seeded to
the solution.
- ‘Dentritic breeding’ in which protruding parts of branched-shape crystals are forced
off by hydrodynamic forces to form new nuclei.
- ‘Contact nucleation’ in which new nuclei are forced off as a result of collisions
between two crystals or between crystals and other surfaces.
- ‘Shear breeding’ in which hydrodynamic sheer forces cause physical wear of the
crystal surface, resulting in new nuclei.
For design purposes, the following formula can be used to approach heterogeneous and
secondary nucleation rates (de Haan & Bosch, 2007):
B o ≈ k N ⋅ ( c − c∗ )
n
(nuclei/s*cm3)
(17)
B is the nucleation rate (number of nuclei per time unit per volume unit), c is the solute
concentration, c* is the equilibrium solute concentration, and kN and n are constants that
must be retracted from experimental data. For heterogeneous nucleation the order (n) can
range from 2 to 9. For secondary nucleation the order is significantly lower and will be in the
range of 0 to 3 (de Haan & Bosch, 2007). If supersaturation is first reached, there will pass a
period of time until nucleation occurs. This time is referred to as ‘induction time’. Several
researchers have investigated the induction time of struvite as a function of supersaturation
(Bhuiyan et al., 2008; Ohlinger et al., 2000; Bouropoulos & Koutsoukos, 2000; Galbraith &
Schneider, 20009). In general, higher supersaturation leads to shorter induction times. It
should be noted that the relations found in literature are of limited value here, since they
were all based on experiments in clear solutions, and most of them under very high
supersaturation.
Once nucleation has taken place, the nuclei start to grow. Crystal growth is a complex subject
on which various theories have been developed (Mullin, 2001). One of the most widely used
theories, the ‘diffusion theory’, divides crystal growth into two subsequent steps: diffusion
and incorporation (van Rosmalen, 1994). In this theory, it is assumed that there is a thin
laminar film of liquid adjacent to the growing crystal face, through which molecules have to
diffuse before they can be incorporated at the crystal’s surface (Mullin, 2001). Within this film,
36
the solute concentration is lower than outside the layer, as figure 4.11 shows. The rates at
which diffusion and incorporation take place can be described separately and are presented in
equation (18) and (19), respectively (de Haan & Bosch, 2007).
crystal
solution
c
(c-ci)
ci
(ci-c*)
c*
fig. 4.11 – Concentration gradients adjacent to a crystal surface (Garside et al., 2002)
dm
= k f ⋅ A ⋅ ( c − ci )
dt
(18)
n
dm
= kr ⋅ A ⋅ ( ci − c∗ )
dt
(19)
In these equations dm/dt is the rate of mass that is diffused or incorporated, A is the crystal’s
surface area, c is the solute concentration outside the film, ci is the solute concentration
directly at the crystal’s surface, c* is the equilibrium solute concentration, and kf, kr and n are
constants. For practical purposes, it is often assumed that n=1, and kf and kr are combined
into one overall rate constant (kG) to obtain the following equation:
dm
= kG ⋅ A ⋅ ( c − c ∗ )
dt
where kG =
kr ⋅ k f
(20)
kr + k f
nucleation rate or
growth rate (dm/dt)
Comparing equation (17) and equation (20), it is clear that both nucleation and growth are
strongly dependent on the difference between actual concentration and equilibrium
concentration, and thus on supersaturation. For growth, a first order dependency is assumed,
while for nucleation the dependency is usually of a higher order. This implies that for low
supersaturation, growth will outweigh nucleation, while for high supersaturation nucleation
will be predominant (figure 4.12).
growth
metastable
zone
nucleation
supersaturation (Sr)
fig. 4.12 – Influence of supersaturation on nucleation and crystal growth rates
(Wiesmann et al., 2007)
37
Metastability
The period in which growth significantly outweighs nucleation is referred to as the
‘metastable zone’. This zone is of particular interest for processes where a certain minimum
crystal size has to be reached. Many industrial crystallizers have separate reactors for
nucleation and growth, where the latter is operated under metastable zone conditions, in
order to have maximum control on the final product’s crystal size. Ali and Schneider (Ali &
Schneider, 2006) and Bhuiyan et al. (Bhuiyan et al. 2008) conducted laboratory experiments
in ultra-pure water to identify the metastable zone for struvite crystallization. In both works it
was concluded that the solubility and supersolubility curves, which are the boundaries of the
metastable zone, are almost parallel lines in the concentration range that was studied. The
minimal constituent concentrations that cause rapid nucleation (supersolubility), as found by
Bhuiyan et al., are processed in a Matlab model (Annex 5) to obtain the corresponding free
PO43-, NH4+ and Mg2+ concentrations in thermodynamic equilibrium. Next to that the free
PO43-, NH4+ and Mg2+ concentrations for solubility (Sr=0) are calculated within the same
concentration range. The resulting curves are presented in figure 4.13.
11
-log([PO4]*[NH4])
10
9
Bhuiyan
Sr=0 (pKso=13,26)
8
Linear (Bhuiyan)
7
6
5
1,5
2
2,5
3
3,5
4
-log[Mg]
Fig. 4.13 – Determination of the solubility (dashed) and supersolubility (solid) curves
Presence of other compounds
Next to magnesium, ammonium and phosphate, numerous other compounds are present in
digested sludge. As explained in section 4.1 (‘Secondary sludge’), PAO’s in the activated
sludge system do not only take up phosphate, but also potassium (K+) and calcium (Ca2+).
These components are, as is phosphate, again released in the anaerobic digesters. Martí et al.
[16] found an average K+ concentration of 237 mg/L and an average Ca2+ concentration of
50 mg/L in the effluent of a WWTPs anaerobic digester. The presence of K+, Ca2+ and other
compounds makes the thermodynamic system much more complicated. These compounds
participate in the chemical equilibria of the (sludge) solution, thereby changing the availability
of free Mg2+, NH4+ and PO43- ions and possibly also changing the equilibrium constants (Ki,
see table 4.2). Besides that, other crystal precipitates form. Gadekar et al. evaluated the
formation of different crystal precipitates in wastewater using a mathematic model (Gadekar
et al., 2009). The crystal precipitates they included are listed in table 4.3. They concluded
that the fraction of struvite in the total crystal precipitate increased as magnesium became
limiting, as the ammonia to phosphate ratio increased, or as the magnesium to phosphate
ratio decreased. Struvite accounted for 92-98.5 % of the total crystal precipitates in model
test runs for digested sludge reject water.
38
Table 4.3 – List of precipitates included in the model by Gadekar et al. (Gadekar et al., 2009)
Number
Chemical name/ Commercial name
Chemical formula
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
magnesium ammonium phosphate, struvite
magnesium hydrogen phosphate, newberyte (MHP)
magnesium phosphate, bobierrite (MP8)
magnesium phosphate, cattiite (MP22)
hydroxyapatite (HAP)
tricalcium phosphate, whitlockite (TCP)
monenite (DCP)
octacalcium phosphate (OCP)
dicalcium phosphate dihydrate, brushite (DCPD)
calcium carbonate, calcite
magnesium carbonate, magnesite
nesquehonite
dolomite
huntite
magnesium hydroxide, brucite
MgNH4PO4•6H2O
MgHPO4
Mg3(PO4)2•8H2O
Mg3(PO4)2•22H2O
Ca5(PO4)3OH
Ca3(PO4)2
CaHPO4
Ca8(HPO4)2(PO4)4•5H2O
CaHPO4•2H2O
CaCO3
MgCO3
MgCO3•3H2O
CaMg(CO3)2
CaMg3(CO3)4
Mg(OH)2
To determine the struvite content, crystal precipitates can be re-dissolved by acid solution
(like HCl) followed by element analysis. Hao et al. (Hao et al., 2009) used this technique to
analyze crystal precipitates that formed at several pH values (pH= 7-12) in ultra-pure water
and in tap water. The tap water contained about 87 mg/L of calcium. In ultrapure water, they
found a sharp decrease in ammonium content in crystal precipitates that formed at pH values
above 10.5. They concluded that the volatilization of NH3 caused this sudden ammonium
decrease, as at high pH values the ammonium is predominantly present as NH3 (see figure
4.7). In tap water, calcium could not be detected in crystal precipitates formed at pH values
below 8.5. At higher pH values (>8.5) struvite formation was found to be seriously limited by
Ca2+ compounds like Ca3(PO4)2 and CaHP4.
Besides the presence of dissolved components, also the presence of suspended solids could
influence the formation of struvite. Several authors observed an interference with struvite
precipitation at TSS concentrations higher than 1000 mg/L (Alp, 2010). The TSS
concentration in digested sludge is typically 40 g/L (Metcalf & Eddy, 2004). Therefore,
interference with suspended solids can be expected when struvite is recovered from digested
sludge. Of course the sedimentation of struvite is also influenced by the TSS concentration,
as the presence of suspended solids can hinder discrete settling.
4.3
CO2 stripping
As previously mentioned, aerating the sludge leads to an increase in pH, due to the stripping
of carbon dioxide (CO2). Just as magnesium, ammonium and phosphate, carbon dioxide is in
equilibrium with other compounds while in solution. At the same time, an equilibrium exists
between dissolved carbon dioxide (CO2(aq)) and carbon dioxide gas (CO2(g)). Of course, the
transformation of CO2(aq) to CO2(g) and vice versa can only take place at the contact surface
between the liquid and the gaseous phase, such as air bubbles or the free sludge/air surface.
CO2
Gaseous phase
CO2
+/- H2O
H2CO3
+/- H+
HCO3-
+/- H+
Liquid phase
CO32-
Fig. 4.14 – Carbonic equilibrium
39
The different compounds in the liquid phase are present according to the thermodynamic
equilibria as presented in table 4.4, where {i} represents the ion activity of the corresponding
compound (equations (7-9)). The reactions between CO2 and H2CO3 are very slow, whereas
the reactions between H2CO3 and HCO3- are almost instantaneous. Therefore, a combination
of both reactions into one equilibrium equation (K1) is allowed (Noorman et al., 1992).
Table 4.4 – Thermodynamic equilibria with CO2 (T=30°C) (de Moel et al., 2005)
Equilibrium equation
Value Ki
{H } ⋅{HCO } = K
4.5*10-7
+
{CO2 }
−
3
1
{CO } ⋅{H } = K
{HCO }
2−
3
+
−
3
2
4.7*10-11
The equilibrium between CO2(aq) and CO2(g) can be described by a single distribution
coefficient kD (van Dijk, 2008):
kD =
cw
cg
(21)
cw = equilibrium concentration of CO2 in the solution (g/m3)
cg = concentration of CO2 in air (g/m3)
Section 4.1 (‘Digested sludge’) outlined that the sludge is saturated with CO2 during
anaerobic digestion. Hence, CO2 is converted from the liquid to the gaseous phase in order to
restore the equilibrium of equation 21. As the concentration of CO2(aq) decreases, HCO3- is
converted to CO2(aq) and CO32- is converted to HCO3- in order to restore the equilibria of
table 4.4. During these conversions, H+ is taken up from the solution. Therefore, the pH rises.
The rate of CO2 transfer from the liquid to the gaseous phase can be written as follows (van
Dijk, 2008):
dcs
= k L a ⋅ ( cw − cs )
dt
(22)
cs = concentration of CO2 in the solution (g/m3)
kLa = gas transfer constant (1/s)
The gas transfer constant kLa is in fact the product of the constants kL and a. The constant a
(m2/m3) represents the specific contact surface area between the solution and air, whereas kL
(1/m*s) is a measure for the renewal of this surface. The larger the contact surface and the
faster the renewal, the faster the transfer of CO2 takes place. The term (cw-cs) can be seen as
the driving force for CO2 transfer. In this case, the concentration of CO2 in solution is higher
than the equilibrium concentration. Therefore, CO2 is removed from the solution and hence
the change in concentration (dcs/dt) is negative. As the concentration of CO2 in solution
decreases, so will the rate of CO2 transfer. For an infinitive aeration time (tÆ∞), the CO2
concentration in solution equals the equilibrium concentration cw (van Dijk, 2008).
40
5. Materials and methods
5.1
Experimental setup
Digested sludge
The digested sludge for the experiments was collected at WWTP West from the overflow of
the anaerobic digesters. Two preliminary tests were performed on the sludge:
- A 13L batch of sludge was washed (by repeatedly flushing with tap-water and careful
decanting in a bucket) and the resulting sediment was analyzed on composition and
crystal size distribution (CSD).
- A 200cm column (diameter=5cm) was filled with sludge and was left for settling.
After 3 days it was determined if settlement had occurred, and to what extent.
The chemical composition, dewaterability and dry solid content (DSC) of the raw digested
sludge were determined during the experiments, for comparison with digested sludge after
struvite formation.
Columns and collection of sediments
Two colums were used during the experiment.
The ‘short column’ is presented in figures 5.1 and 5.2. It was used as a reaction vessel for
struvite formation in batch experiments. This column (diameter=20cm, height=100cm) was
filled wit raw digested sludge (max. 25 L) and aerated with a custom-made bubble aerator
(figure 5.3). The airflow was measured with a VA (Variable Area) flow meter (range: 200-800
L/h). During the experiments, chemicals were added at the sludge surface. The following
chemicals were used:
- MgCl2 (33% solution)
- Cationic Polymer (0,21% solution)
- Ammonium chloride salt (NH4Cl (s))
- Potassium dihydrogen phosphate salt (KH2PO4 (s))
Samples were taken from a sampling point between the sludge surface and the aeration
system. The pH value and temperature were measured with a Hach HQ40d multi meter. At
the end of the experiment, after 30 minutes of sedimentation, the sediment was tapped from
a conical outlet at the bottom of the column. The remainder of the sludge was also tapped
from the conical outlet and was carefully washed by repeatedly flushing with tap water and
careful decanting in a bucket.
The ‘long column’ as presented in figures 5.4 and 5.5 was used to wash the collected
sediment from the short column. The sediment was disposed in the column (diameter =
0.05m, height = 2.00 m) and flushed in upward direction using tap water. The water flow
was measured with a VA flow meter (range: 20–100 L/h). The washing water (with sludge
particles) was collected in buckets via an overflow construction. After washing, the sediment
was collected from the bottom of the column and was left for atmospheric drying for several
days. After that, the CSD was determined.
41
pH
Chemicals
Sample
Flow
meter
Air
Outlet
Fig. 5.1 – Short column
(schematically)
Fig. 5.2 – Short column
(impression)
Fig. 5.3 – Aeration system
short column
Bucket
tap
Flow
meter
Bucket
Fig. 5.4 – Long column (Schematically)
42
Fig. 5.5 – Long column (impression)
Sample analysis
The samples taken from the short column were analyzed on Dry Solids Content (DSC),
orthophosphate (PO4-P), ammonium (NH4-N) and magnesium concentrations. The DSC was
automatically measured in a Sartorius MA35 DSC measuring device. To measure the
orthophosphate, ammonium and magnesium concentrations, part of the sample was
centrifuged, filtered over a 45μm nylon Syringe filter, diluted if necessary, and analyzed using
a Dr. Lange lasa50 cuvette test apparatus and the following cuvette tests:
- PO4-P: Hach Lange LCK350 (range: 0.5-50 mg/L)
- NH4-N: Hach Lange LCK302 (range: 47-130 mg/L)
- Mg2+: Hach Lange LCK326 (range: 0.5-50 mg/L)
Dewaterability tests
To determine the dewaterability before struvite formation (pre dewaterability), a 200 g
sample was prepared from sludge and polymer solution (0.21%) in such a way that 10 g of
polymer was available per kg of dry solids. Sludge and polymer were intensively mixed for 3
seconds using a kitchen blender with a timer connected to its power supply. For example, if a
DSC of 3.5% was measured in the untreated sludge, 171.4 g of sludge was mixed with 28.6 g
of polymer solution. Subsequently, the sludge/polymer mixture was put at once on a woven
synthetic filter with minimum filter resistance (figure 5.6-5.7). The amount of percolated
water was measured with a balance (accuracy = 0.01 g) after 5, 10, 15, 20, 30, 40, 50, 60,
90 and 120 seconds.
Sludge/polymer
mixture
Balance
Fig. 5.6 – Dewaterability test (schematically)
Fig. 5.7 – Dewaterability test (impression)
The dewaterability after treatment (aeration and/or MgCl2 dosage), or post dewaterability,
was determined in the same way and with the same ratio of sludge/polymer solution as for
the pre dewaterability, regardless the DSC after treatment. This was done to assure that a
change in dewaterability was not caused by a change in polymer addition.
The dewaterability test was borrowed from the company BASF Nederland. BASF uses it as a
quick field-test to determine suitable polymer dosing rates for thickening and dewatering
installations of their customers.
43
5.2
Experimental procedure
Apart from the preliminary tests (section 5.1, ‘The digested sludge’), six different kinds of
experiments were conducted. In table 5.1, an overview is given of the experiments, including
some typical values.
1. Aeration
In the aeration experiments, the short column was filled with a certain amount of raw
digested sludge. MgCl2 solution (33%) was added in such a way that the ratio of total
magnesium to total phosphate (CT_Mg/CT_PO4, or Mg:PO4) correspond to the values in
table 5.1. At t=0, the aeration was switched on and was adjusted to obtain the desired
aeration rate. During the experiments, the pH and temperature of the sludge were monitored.
The sediment was not collected and no dewaterability tests were performed.
2. Struvite formation (with aeration)
During these experiments, 20 L of raw digested sludge was aerated until a certain pH value
was reached (pH setpoint, table 5.1). At the desired pH, MgCl2 was added at t=0 to obtain
the desired Mg:PO4 ratio. The sludge was aerated at 500 L/h for 10 minutes, monitoring the
pH value at intervals of 30 s. After that (at t=10 min) a sample was taken and analyzed on
DSC, component concentrations and dewaterability. This sample was compared to a sample
taken from the raw digested sludge.
In experiments 10,11,12,13,14,19 and 22, the formed struvite was allowed to settle during 1
hour and subsequently collected and washed according to the method described in section
5.1. In experiments 15-18, next to a sample of the raw sludge and a sample at t=10 min,
also a sample was taken just before MgCl2 addition. Another sample was taken after the
sludge had been aerated (500 L/h, after struvite formation) until the pH setpoint was again
reached.
3. Struvite formation without aeration
To investigate whether a better dewaterability could be obtained if MgCl2 was added without
aeration, a beaker experiment was done in duplicate (experiments 24 and 26). A beaker was
filled with 0.4 L of sludge and continuously (automatically) stirred. At t=0, MgCl2 was added
to obtain a Mg:PO4 ratio of 1.5. At t=10 minutes, a sludge sample was taken and compared
to a raw sludge sample. In experiment 24, the pH value was monitored during the 10
minutes of struvite formation.
4. Struvite crystal growth
In experiment 25, a batch of sludge was aerated to pH=7.2 and MgCl2 was added to obtain a
Mg:PO4 ratio of 1.00. After 5 minutes of reaction, the component concentrations were
measured. NH4Cl and KH2PO4 were added until the original (initial) ammonium and
orthophosphate concentrations were reached. Then, MgCl was again added at a dosing ratio
of Mg:PO4=1.00. This cycle was repeated 3 times.
5. Dewaterability versus polymer dosage
In experiment 20, a 20 L batch of raw sludge was aerated until a pH value of 7.6 was
reached. Then, MgCl2 was added to obtain a Mg:PO4 ratio of 1.5. After 10 minutes of
aeration at 500 L/h, a sludge sample was taken and analyzed on dewaterability at different
polymer dosages (section 5.1, ‘Dewaterability test’).
44
6. Dewaterability versus pH
In experiment 9, a 20 L batch of raw sludge was aerated at 500 L/h. At several pH values, a
sample was taken and the dewaterability was determined.
Table 5.1 – overview of the experiments
Exp. no.
description
Volume (L) pH setpoint (-)
1
aeration
25
2
aeration
25
3
aeration
25
4
aeration
25
5
aeration
25
6
aeration
15
7
aeration
25
8
aeration
10
9
dewaterability versus pH
20
10
struvite formation
20
7.2
11
struvite formation
20
7.6
12
struvite formation
20
7.4
13
struvite formation
20
7.8
14
struvite formation
20
8.0
15
struvite formation
20
7.6
16
struvite formation
20
7.8
17
struvite formation
20
8.0
18
struvite formation
20
7.4
19
struvite formation
20
7.6
20
dewaterability versus polymer dosage
20
7.6
21
struvite formation
20
7.6
22
struvite formation
20
7.6
23
struvite formation
20
7.6
24
struvite formation without aeration
0.4
25
struvite crystal growth
20
26
struvite formation without aeration
0.4
-
5.3
Mg:PO4 (-) Q air (L/h)
no dosage
500
1.5
500
no dosage
500
1.5
750
no dosage
750
no dosage
300
1.0
500
no dosage
200
no dosage
500
1.5
variable
1.5
variable
1.5
variable
1.5
variable
1.5
variable
1.5
variable
1.5
variable
1.5
variable
1.5
variable
2.0
variable
1.5
variable
1.0
variable
0.5
variable
1.2
variable
1.5
no aeration
variable
variable
1.5
no aeration
Mathematic modeling
The concentrations of phosphate (PO4-P), ammonium (NH4-N) and magnesium that were
measured during the experiments represent the total concentrations CT_PO4, CT_NH4 and
CT_Mg (see section 4.2, ‘solubility and saturation’). To calculate the free ionic concentrations
of PO43-, NH4+ and Mg2+, the thermodynamic equilibria as given in section 4.2 were solved by
modeling the equilibria in Matlab. The script is presented in Annex 5.
Several other calculations were performed in Matlab, such as the determination of phosphate
removal at different magnesium dosing rates (section 6.3). These calculations are all based
on the abovementioned thermodynamic equilibria. The used scripts differ only slightly from
the script presented in Annex 5 and are not included in this report.
45
46
6. Results and discussion
This section presents and discusses the results of the experiments. The results are grouped
according to the research questions (section 3.2):
- What are the digested sludge conditions at WWTP West?
- What are the process settings for reaching optimal sludge dewaterability?
- What are the process settings for reaching optimal phosphate removal?
- What are the process settings for reaching optimal struvite recovery?
The change of pH due to aeration (CO2 stripping), as observed in the experiments, is
discussed separately.
6.1
Digested sludge conditions at WWTP West
The collected digested sludge has the specific characteristics as outlined in section 4.1: it is
an oil-like, apparently homogeneous, almost black substance with an inoffensive odor. The
sediment that was found after washing a 13 L batch of raw digested sludge seemed to
consist mainly of crystal precipitates (probably mostly struvite) and sand. Also, small bits of
wood and plant seeds (<5mm) were found. Figure 6.1 shows a close-up picture of this
sediment. The CSD of the raw digested sludge will be discussed in section 6.4. From the
settling test, the settleability of the sludge proved to be very bad. No sludge/water-interface
could be identified in the upper part of the column. Instead, a watery but very turbid zone
with sharp limits had become visible in the lower part of the column (figure 6.2).
Fig 6.1 – Close-up of sediment
Fig. 6.2 – Watery zone in the sludge
In table 6.1, average values and value ranges are given for the most important parameters of
the raw digested sludge, extracted from both data collected by Waternet and data collected
during the experiments. In Annex 4, the complete experimental data for raw digested sludge
is included.
Phosphate concentrations varied over a wide value range throughout the year, probably
caused by temperature variations and resulting variations in bacterial performance (Narashiah
& Morasse, 1984). Figure 6.3 shows the seasonal orthophosphate variation. The experiments
were conducted in the period October–November 2010. The orthophosphate concentrations
found in the experiments were slightly higher than the concentrations collected by Waternet
in the reject water in the same period of 2009.
47
Table 6.1 – Digested sludge characteristics
Data Waternet
(july 2009 to june 2010)
Experiments
parameter
unit
range
average
range
average
-
7.00 – 7.21
7.17
6.3 – 7.4
7.1
55.44
pH
Alkalinity
mmol/L
-
-
43.89 – 68.33
DSC
%
3.47 – 3.58
3.52
-
-
DSC (after dewatering)
%
-
-
21.7 – 23.2
22.3
ºC
28.3 – 33.6
31.2
-
-
mg/L
290 - 346
318
140 - 415 (a)
246 (a)
Temperature
PO4-P
NH4-N
mg/L
858 - 928
900
-
-
Mg
mg/L
21.3 - 27,9
25.4
-
-
(a) concentration in centrate
PO4-P concentration (mg/L)
500
400
300
200
100
jul-10
jun-10
mei-10
apr-10
mrt-10
feb-10
jan-10
dec-09
nov-09
okt-09
sep-09
aug-09
jul-09
0
month
Fig. 6.3 – Seasonal orthophosphate variation
(centrifuge reject water, adapted from raw measuring data by Waternet)
6.2
Sludge dewaterability
Figure 6.4 shows the average initial (pre) dewaterability of experiments 15-20 with error bars
to mark the standard deviation. The complete data is included in Annex 4. As is illustrated, an
improvement in sludge dewaterability leads to a steeper curve, while deterioration leads to a
flatter curve.
To investigate the influence of the pH value on sludge dewaterability, a batch of sludge was
aerated without addition of MgCl2. At Several pH values, a sample was taken and the
dewaterability was determined. Figure 6.5 shows the results. Until pH=7.5, no change in
dewaterability was found. However, at higher pH values (7.75 and 8.00) the dewaterability
curves showed a significant deterioration.
A possible explanation for this phenomenon can be given on the basis of the DCB theory
(section 4.1, ‘sludge dewatering’). At high pH values, calcium reacts with carbonate (CO32-)
into calcium carbonate (CaCO3 (s)) (de Moel et al., 2005). The resulting decrease in (divalent)
Ca2+ ions could lead to decreased bridge formation, thereby weakening the sludge
flocculation and deteriorating the dewaterability.
48
100
90
improvement
80
filtration units (g)
70
60
50
40
30
deterioration
20
10
0
0
20
40
60
80
100
120
140
time (s)
Fig. 6.4 – Pre dewaterability
100
90
filtration units (g)
80
pre dewaterability
70
pH = 7.50
60
50
pH = 7.75
40
pH = 8.0
30
20
10
0
0
20
40
60
80
100
120
140
time (s)
Fig. 6.5 – Dewaterability at aeration to different pH values without MgCl2 dosing
In experiments 15, 19, 21 and 22, the sludge was aerated up to pH=7.6. At that moment,
MgCl2 was dosed to obtain a certain Mg:PO4 ratio (0.5, 1.0, 1.5 or 2.0). After 10 minutes of
reaction (while aerating), a sludge sample was taken and tested on dewaterability. The
results are presented in figure 6.6.
As can be seen, the dewaterability improved at all dosing ratios. Of course, the addition of
MgCl2 leads to an increase in Mg2+ ions and would hence lead to an improvement in
dewaterability according to the DCB theory. Table 6.2 presents the final magnesium
concentrations (after 10 minutes of reaction). From these values, it seems logical that a
Mg:PO4 ratio of 0.5 leads to only a slight improvement in dewaterability, since only a slight
increase in magnesium occurs as compared to the initial concentration.
49
100
90
filtration units (g)
80
70
pre dewaterability
60
Mg:PO4 = 0.5
50
Mg:PO4 = 1.0
40
Mg:PO4 = 1.5
30
Mg:PO4 = 2.0
20
10
0
0
20
40
60
80
100
120
140
time (s)
Fig. 6.6 – Dewaterability at different MgCl2 dosages (aerated to pH=7.6)
Table 6.2 – Final magnesium and orthophosphate concentrations
exp. no. Mg:PO4 (-)
22
0.5
21
1.0
15
1.5
19
2.0
average initial
concentration
final CT_Mg
Final CT_PO4
concentration (mg/L) concentration (mg/L)
27.9
137
79.2
62.4
175
27.2
332
25.7
25.4
318
Another explanation for the improved dewaterability could be the reduction in
orthophosphate concentration (table 6.2). As mentioned in section 4.1 (‘Sludge dewatering’),
PCS claims that a lower orthophosphate concentration leads to an improved dewaterability. In
most cases, however, the reduction in orthophosphate takes place parallel to the increase in
magnesium concentration. For that reason, it is very difficult to determine which proposed
mechanism is actually occurring.
Dosing ratios higher than Mg:PO4=1.0 did not further improve sludge dewaterability. It is
uncertain what caused this sudden stagnation. From the point of view of the DCB theory, it is
imaginable that the available (negatively charged) locations for bridge formation at the sludge
particles were all occupied above a certain magnesium concentration. This would mean that a
maximum dewaterability is reached at some point, which is not influenced by any further
dosing.
Next to that, experiments were conducted with a constant Mg:PO4 dosing ratio of 1.5. In
these experiments, a batch of raw sludge was aerated to reach a certain pH value (7.4, 7.6,
7.8 or 8.0). MgCl2 was added and after 10 minutes of reaction (while aerated) sludge samples
were tested on dewaterability. The results are shown in figure 6.7.
No significant differences in dewaterability were found. It seems that the addition of MgCl2
compensates the deterioration of dewaterability that is normally caused by pH rise. This
agrees with the DCB theory and the statements made above, since a possible loss in calcium
ions could be compensated by the addition of magnesium ions. On top of that, it corresponds
to the presumption that the dewaterability can reach a certain maximum when the negatively
charged sludge particles are all occupied by positively charged ions.
50
100
90
pre dewaterability
filtration units (g)
80
pH = 7.4
70
60
pH = 7.6
50
pH = 7.8
40
30
pH = 8.0
20
stirring (1)
10
stirring (2)
0
0
20
40
60
80
100
120
140
time (s)
Fig. 6.7 – Dewaterability at different pH values after MgCl2 dosage (Mg:PO4=1.5)
In figure 6.7, also the results from two identical ‘stirring experiments’ are included. In these
experiments, a batch of sludge was stirred instead of aerated, while MgCl2 was added to
obtain a Mg:PO4 ratio of 1.5. Again, after 10 minutes of reaction time the sludge was tested
on dewaterability.
The dewaterability obtained in these experiments exceeded the best results of all other
experiments.
Since these findings do not fit the earlier reasoning and the DCB theory, no obvious
explanation for this phenomenon was found. As the best dewatering result was found in
these experiments, more attention should be paid to struvite formation with stirring instead of
aeration in future research.
In experiment 20, a series of dewatering tests was performed on a batch of (treated) sludge
to investigate whether the addition of MgCl2 could lead to savings in polymer use. In this case,
the MgCl2 was added at pH=7.6 and at a Mg:PO4 ratio of 1.5. After 10 minutes of reaction
(while aerated), samples were taken and tested on dewaterability. The results are presented
in figure 6.8.
With the same sludge/polymer solution ratio as before treatment, a similar dewaterability was
found as in previous experiments that were conducted at pH=7.6 and Mg:PO4=1.5. A
decrease in polymer solution volume (relative to sludge volume) led to a deterioration in
dewaterability. As can be seen in figure 6.8, at 80% of the initial polymer dosage, a sludge
dewaterability comparable to the pre dewaterability was found. Apparently, the addition of
MgCl2 cannot substitute the application of a (cationic) polymer, since the dewaterability was
found to deteriorate as soon as the polymer dosing ratio was reduced. A possible explanation
for this is that bridge formation is not the only working mechanism of the polymer. Polymer
molecules with a large molecular weight can also be adsorbed at the sludge particle surface
by attractive body forces. On top of that, the elongated shape of the polymer molecules
allows multiple sludge particles to be connected to a single polymer molecule, thereby even
further reinforcing flocculation. It should be stressed that returning to the initial
dewaterability compensates the improvement in DSC of the dewatered sludge, along with its
economical advantage.
51
100
90
filtration units (g)
80
pre dewaterability
70
60
polymer = 100%
50
40
polymer = 90%
30
20
10
polymer = 80%
0
0
20
40
60
80
100
120
140
time (s)
Fig. 6.8 – Dewaterability at different polymer dosages after treatment
A disadvantage of the used method to determine dewaterability, is the impossibility to
convert the results to a single parameter with economical significance, such as the DSC value
of dewatered sludge. The method is, however, very well suitable for comparing the
dewaterability performances of differently treated sludge.
6.3
Phosphate removal
A complete overview of the initial and final component concentration is included in annex 4.
Using the Matlab script in annex 5, the free ionic concentrations (PO43-, NH4+, Mg2+) were
calculated. By plotting the product of the ionic phosphate and ammonium concentrations
against the ionic magnesium concentration, a graphical presentation of the equilibrium
relations between the components is obtained. The result is presented in figure 6.9.
11
-log([PO4]*[NH4])
10
Bhuiyan
9
Sr=0 (pKso=13.26)
initial concentrations
8
final concentrations
Sr=0 (pKso=12.99)
Linear (Bhuiyan)
7
6
5
1,5
2
2,5
3
3,5
4
-log[Mg]
Fig. 6.9 – Relations between component concentrations
52
In this figure, some randomly picked initial concentrations are plotted as well as the final
concentrations from experiments 10-19, 21 and 22. Also, solubility curves (based on the
thermodynamic equilibria) for pKso=13.26 and pKso=12.99 and the supersolubility curve as
determined by Bhuiyan (see section 4.2, ‘metastability’) are added. The Ionic Strength (I)
was assumed to be 0.02 mol/L, based on literature (Metcalf & Eddy, 2004).
The trend-line through the combination of initial and final concentrations is equal to the
theoretical solubility curve for pKso=12.99. This suggests that the relation between
components in equilibrium can indeed be described by the thermodynamical equations as
given in section 4.2. On top of that, it means that the components were in equilbrium after
10 minutes of reaction, suggesting that the reaction time for struvite is less than 10 minutes.
By analyzing these experimental results, a model can be derived that calculates the amount
of MgCl2 addition required to obtain the desired phosphate removal at given conditions.
A relation between phosphate and magnesium is searched for. Therefore, first it was
investigated whether the ammonium concentration can be taken as a constant. In figure 6.10
the trend-line through initial and final concentrations is presented in a same manner as in
figure 6.9. In figure 6.11, the ammonium concentration is neglected and the ionic phosphate
concentration is simply plotted against the ionic magnesium concentration. The trend-line
that is obtained while neglecting ammonium has only a slightly lower R2 value (and hence a
slightly higher error) than the trend-line including ammonium. For this reason, it is concluded
that ammonium can be taken as a constant for the concentration range that is studied.
11
9
8, 5
10
2
R = 0.9751
9,5
-log[PO4]
-log([PO4]*[NH4])
10,5
9
8,5
R2 = 0.9708
8
7, 5
8
7
7,5
7
6, 5
1,5
2
2,5
3
3,5
4
1, 5
2
2, 5
-log[Mg]
3
3, 5
4
-log[Mg]
Fig. 6.10 – Error including ammonium
Fig. 6.11 – Error excluding ammonium
Next, the relation between phosphate removal and magnesium removal was evaluated. As
illustrated in figure 6.12, this relation is the final question mark in calculating the needed
Mg2+ dosage from initial (start) and final (end) concentrations.
-log[PO4]
End
Solubility curve
Start
?
Mg2+ dosage
-log[Mg]
Fig. 6.12 – Determination of needed MgCl2 dosage
53
1,00
1,00
0,90
0,90
0,80
0,80
∆CT_Mg/∆CT_PO4
∆CT_Mg/∆CT_PO4
The removal ratios (∆CT_Mg/∆CT_PO4) that were found at different pH setpoints and at
different dosing ratios (Mg:PO4) are presented in figure 6.13 and figure 6.14, respectively.
The complete data is included in Annex 4. Duplicate conditions are present in both diagrams.
However, the removal ratios found under identical conditions varied significantly. The
observed variation is within the same range as the variation for different conditions, making it
impossible to identify any relation between pH and removal ratio and between dosing ratio
and removal ratio based on this data. Because no clear relation was found between removal
ratio, dosing ratio and pH, the average value (∆CT_Mg/∆CT_PO4=0.8) was used in the model.
In other words, the model assumed that 0.8 mole of magnesium is needed for the removal of
1 mole of phosphate.
0,70
0,60
0,50
0,40
0,30
0,20
0,10
0,70
0,60
0,50
0,40
0,30
0,20
0,10
0,00
7,20
7,22
7,40 7,40
7,60 7,60
7,80
7,80 8,00
0,00
8,00
0,5
pH (-)
1
1,2
1,5
1,5
2
Mg:PO4 (-)
Fig. 6.13 – Removal ratios at different pH setpoints
(Mg:PO4=1.5)
Fig. 6.14 – Removal ratios at different
dosing ratios (pH setpoint=7.6)
It should be noted that the removal ratio for a 100% pure struvite product would be 1.00,
since magnesium, ammonium and phosphate react in a 1:1:1 ratio according to equation (1).
The experimental results suggest that phosphate is also removed by other reactions, forming
different precipitates such hydroxyapatite (section 4.2, ‘presence of other compounds’).
Assuming a removal ratio of 0.8 and a constant CT_NH4 concentration, a Matlab model was
constructed. Figure 6.15 illustrates the concept of this model.
initial CT_PO4
initial CT_Mg
constant CT_NH4
constant IS
constant pKso
removal ratio
constant pH
desired final CT_PO4
MODEL
Based on
thermodynamic equilibriums
Mg:PO4 (dosing ratio)
Fig. 6.15 – Concept of the model
Repeatedly running this model, using different pH values and different desired final
phosphate concentrations resulted in a series of ‘removal curves’ as presented in figure 6.16
(initial CT_PO4=318 mg/L P, initial CT_Mg=25.4 mg/L, constant CT_NH4=870 mg/L N,
constant IS=0.02 mol/L, ∆CT_Mg/ ∆CT_PO4=0.8, pKso=12.99). The constant CT_NH4
concentration was chosen to be 870 mg/L N since this was the average value of the
concentrations before and after struvite formation of all experiments (Annex 4, table A4.6).
54
2
1,8
1,6
P-final = 5 mg/L
Mg:PO4 (-)
P-final = 10 mg/L
P-final = 15 mg/L
1,4
P-final = 20 mg/L
P-final = 30 mg/L
P-final = 40 mg/L
1,2
P-final = 50 mg/L
1
0,8
0,6
7,2
7,3
7,4
7,5
7,6
7,7
7,8
7,9
8
pH (-)
Fig. 6.16 – Removal curves
To test the model, it was applied on the data collected in experiments with varying dosing
ratios. The results are presented in table 6.3. Generally, quite accurate results were obtained
(max error = 15%). The data from experiments 11, 19, 21 and 22 was used to build the
model. Therefore it cannot be used to validate the model. Applying the model on the data of
experiment 23, a good result was obtained. However, more independent experiments are
needed to determine the validity of the model, preferably in other experimental setups.
This model was based on average values. For that reason, the dosing ratio is sometimes
overestimated and sometimes underestimated. In practice, a slight overestimation is
preferred over an underestimation, to be on the ‘safe side’. For a ‘safe-side-prediction’, a
higher value for the removal ratio can be selected, for example ∆CT_Mg/ ∆CT_PO4=0.9. In
this way, the model assumes that more magnesium is needed (0.9 mole instead of 0.8 mole)
to remove 1 mole of phosphate.
Table 6.3 – Testing the model
experiment no. dosing ratio (model) dosing ratio (experiment) error (%)
11
1.37
1.5
-8.7
19
2.22
2
11.0
21
0.85
1
-15.0
22
0.53
0.5
6.0
23
1.04
1.2
-13.3
55
The model can be used in two ways:
- As a decision-making tool in the design process: with the curves of figure 6.16,
consequence of the choice for a certain (constant) pH and dosing ratio on
phosphate removal can be predicted.
- As a steering tool during operation: if the initial phosphate concentration and
(constant) pH value in the reactor are monitored, the model can determine
needed MgCl2 dosage to obtain the desired phosphate concentration.
6.4
the
the
the
the
Struvite recovery
During struvite crystallization, the ions Mg2+, NH4+ and PO43- are removed from the solution
by reacting with 6 H2O molecules to MgNH4PO4•6H2O (see equation 1). As these ions are
removed, the thermodynamic equilibria (table 4.2) force other components to react in order
to restore the balance. For example, in the pH range that is studied (pH=7-8), MgHPO4 splits
into Mg2+ and HPO42- (figure 4.6.). As the removal of PO43- forces HPO42- and H2PO4- to split
(figure 4.8), protons are released during reaction. These protons are partly taken up by
buffering components. A certain part of the protons, however, will remain in solution in its
free form, thereby decreasing the pH value.
Since the pH drops as struvite crystallizes, the pH value is not only one of the most important
parameters influencing struvite crystallization (section 4.2, ‘solubility and saturation’), but
changes in pH can also be used for monitoring the speed of struvite reaction. The pH dip
after MgCl2 dosing (at t=0) for different dosing ratios and for different initial pH values is
presented in figure 6.17 and 6.18, respectively. At higher dosing ratios, more struvite will
crystallize and therefore more protons are released. This was confirmed by the experiments
(figure 6.17), as a higher dosing ratio led to a larger pH decrease. In all cases the minimum
pH value was reached after approximately 5 minutes, suggesting this was the minimum
reaction time needed for struvite formation. After reaching its minimum value, the pH again
gradually increased as a result of CO2 stripping (section 4.3.).
7,7
8,2
7,6
8
7,8
Mg:PO4 = 0.5
7,4
Mg:PO4 = 1.0
Mg:PO4 = 1.5
7,3
Mg:PO4 = 2.0
7,2
pH (-)
pH (-)
7,5
7,6
7,4
7,2
7,1
7
7
0
2
4
6
8
10
time (min)
Fig. 6.17 – pH drop at different dosing ratios
6,8
0
2
4
6
8
Fig. 6.18 – pH drop at different initial
pH values (Mg:PO4=1.5)
In figure 6.18, the pH drops for several initial pH values at the same dosing ratio are shown.
These curves also imply a total reaction time below 5 minutes. It is difficult to compare the
curves as the pH is in fact a logarithmic scale (-log[H+]). For this reason, a conversion is
made to curves that display the cumulative change in H+ concentration over time (figure
6.19).
Higher initial pH values resulted in a smaller proton increase. This was not expected, since at
higher pH values the struvite production should be larger than at lower pH values. A possible
explanation is that at higher pH values (around pH=8), more HPO42- ions and less H2PO4- ions
56
10
time (min)
are readily available compared to lower pH values (figure 4.8). This means that at higher pH
values fewer protons have to be released in order to ‘produce’ PO43-. This phenomenon could
outweigh the increase in proton production due to the increase in struvite production.
7,0E-08
6,0E-08
initial pH = 7.2
[H+] (mol/L)
5,0E-08
initial pH = 7.4
4,0E-08
initial pH = 7.6
initial pH = 7.8
3,0E-08
initial pH = 8.0
2,0E-08
1,0E-08
0,0E+00
0
1
2
3
4
5
time (min)
Fig. 6.19 – Cumulative H+ increase at different initial pH values (Mg:PO4=1.5)
The disadvantage of the curves as presented above, is that they are in fact combined results
of struvite production on one hand, and CO2 stripping on the other hand. For that reason,
wherever a minimum pH is reached would mean that struvite formation and CO2 stripping
cause an equal (but opposite) change in H+ concentration at that point, rather than the
completion of struvite formation. Figure 6.20 presents the pH drop for experiment 24. In
experiment 24, the sludge was stirred instead of aerated and therefore the pH curve should
only display the pH decrease due to struvite formation.
7,3
pH (-)
7,2
7,1
7
6,9
6,8
0
5
10
15
20
25
30
time (min)
Fig. 6.20 – pH drop with stirring instead of aerating (Mg:PO4=1.5)
This curve shows a similar course as the previous curves. The pH value reaches a minimum
within 5 minutes. After that, the pH increases very slightly over time (note the difference in
scale). This slight increase could be caused by some CO2 stripping due to stirring of the
sludge.
In the next step, the sediment that was collected in the struvite formation experiments was
sieved to obtain information about the crystal size distribution (CSD). The sieve curves of the
different experiments were compared to each other and to the sieve curve of the washed
sediments of raw digested sludge (blank). As the major part of the sediment was identified as
semi-transparent crystals, it was assumed that the sieve curves represent the CSD of struvite.
57
The sieve curves of struvite formed at different initial pH values and at different dosing ratios
are presented in figure 6.21 and 6.22, respectively.
100
90
80
sediment passing (%)
70
pH = 7.4
60
pH = 7.6
50
pH = 7.8
pH = 8.0
40
blank
30
20
10
0
0,01
0,1
1
10
mesh size (mm)
Fig. 6.21 – Sieve curves from experiments with different initial pH values (Mg:PO4=1.5)
100
90
80
sediment passing (%)
70
60
Mg:PO4 = 1.0
Mg:PO4 = 1.5
50
Mg:PO4 = 2.0
blank
40
30
20
10
0
0,01
0,1
1
10
mesh size (mm)
Fig. 6.22 – Sieve curves from experiments with different dosing ratios (initial pH=7.6)
58
As can be seen, the sieve curves show very little variation. The collected struvite appears to
have almost exactly the same CSD in each experiment. Moreover, the struvite from the
stuvite formation experiments did not show any significant difference from the struvite that
was collected from the raw digested sludge. This suggests that the formed struvite had either
an insignificant mass in relation to the mass of the sediment that was already present, or was
too small to recover with the washing method as described in section 5.1.
If struvite is formed but not recovered, this should contribute to the DSC of the treated
sludge. In table 6.4, the DSC values before and after the experiments are printed in columns
1 and 2. Column 3 shows the DSC increase. In each experiment an increase of DSC was
found, mostly of around 0.23 percent points. Experiment 22 showed a smaller increase,
which could be caused by the low dosing ratio used in that experiment (Mg:PO4) and hence
the lower struvite production.
Table 6.4 – DSC increase due to struvite formation
column
experiment no.
1
2
3
4
5
6
7
8
DSC pre DSC post delta DSC delta PO4-P struvite max delta DSC sediment
struvite
(%)
(%)
(measured)
(mg/L)
(mg/L)
(calculated) found (g/L) calculated(g/L)
10
3.51
3.84
0.33
291.95
2313.82
0.23
1.70
2.31
11
3.51
3.82
0.31
300.05
2378.02
0.24
1.24
2.38
12
3.48
3.78
0.30
282.90
2242.10
0.22
1.21
2.24
13
3.49
3.70
0.21
309.80
2455.29
0.25
1.24
2.46
14
3.58
3.85
0.27
293.90
2329.28
0.23
1.78
2.33
19
3.56
3.81
0.25
292.30
2316.60
0.23
1.58
2.32
22
3.53
3.69
0.16
183.00
1450.35
0.15
1.09
1.45
raw sludge
1.05
From the measured change in orthophosphate concentration (column 4) the maximum
struvite production was calculated (column 5), assuming a reaction ratio of M:A:P=1:1:1.
From the struvite production the theoretical DSC increase was calculated, assuming that no
struvite was recovered (column 6). Against expectations, the measured DSC increase was in
most cases even higher than the calculated maximum DSC increase due to struvite formation.
A possible explanation is that also precipitates without phosphate (such as calcium carbonate
CaCO3) are formed and are retained in the sludge. In columns 7 and 8 the collected sediment
mass and the calculated struvite mass are displayed, converted to 1 L of sludge. From these
numbers it is clear that not all struvite was recovered, since the sum of the sediments in raw
sludge (1.05 g/L) and the calculated struvite (mostly around 2.32 g/L) strongly exceeds the
sediments found after struvite formation (1.09-1.78 g/L).
The lack of significant variation in sieve curves, the rise in DSC value and the relatively low
mass of sediments that were recovered, suggest that the struvite formed during the
experiments was too small to recover and was lost during the washing procedure. The (larger
sized) struvite that was collected after treatment was probably already present in the raw
sludge, and might have been formed in the anaerobic digesters. This suggests that during
struvite crystallization, nucleation was the predominant mechanism, rather than growth
(section 4.2, ‘nucleation & growth). To investigate whether this complies with theory, some
(theoretical) concentrations just after MgCl2 dosing but just before struvite formation were
plotted in relation to the metastable zone (figure 6.23).
This figure suggests that the struvite formation experiments were operated around the
boundary of the metastable zone. As explained in section 4.2, in the metastable zone growth
should theoretically outweigh nucleation. According to the figure, experiments at low pH
values and low dosing ratios would have been operated within the metastable zone, while
experiments at high pH values and high dosing ratios would have been operated outside the
limits of the metastable zone.
59
11
-log([PO4]*[NH4])
10
Supersolubility curve
Solubility curve
9
pH = 7.2
pH = 7.6
8
pH = 8.0
2.0 1.5 1.0
7
(Mg:PO4)
6
5
1,5
2
2,5
3
3,5
4
-log[Mg]
Fig. 6.23 – Exceeding of the metastable limit due to high dosing ratios or high pH values
For the growth experiment (experiment 25), a dosing ratio of Mg:PO4=1.00 was chosen in
combination with a (maximum) pH value of 7.2. Based on figure 6.23, these conditions
should guarantee operation within the metastable zone. Figure 6.24 presents the
concentrations of the components at several points in the experiment:
- Start: in the raw digested sludge (measured).
- Directly after dosage: after MgCl2 dosage but before struvite formation (calculated).
- Post (1): after 5 minutes of reaction (measured).
- Post (2): after dosing NH4Cl, KH2PO4 and MgCl2, and 5 minutes of reaction
(measured).
- Post (3): after dosing NH4Cl, KH2PO4 and MgCl2, and 5 minutes of reaction
(measured).
From this figure it seems that the growth experiment was conducted completely within the
metastable zone.
11
-log([PO4]*[NH4])
10
Supersolubility curve
Solubility curve
9
Start
Directly after dosage
8
Post (1)
Post (2)
7
Post (3)
6
5
1,5
2
2,5
3
3,5
4
-log[Mg]
Fig. 6.24 – Component concentrations during the growth experiments
Despite the attempt to operate within the metastable zone, the struvite that was collected
from the growth experiments did not show any increase in crystal size, as can be seen in
60
figure 6.25. The average crystal size was even slightly smaller than the average crystal size of
the sediments collected from the raw sludge (blank). The latter could simply be caused by
variance in composition of the raw sludge. It seems, however, that even in the growth
experiment no crystal growth was achieved.
100
90
sediment passing (%)
80
70
60
blank
50
growth
40
30
20
10
0
0,01
0,1
1
10
mesh size (mm)
Fig. 6.25 – Sieve curve from the growth experiment
A very plausible reason for the lack of crystal growth could be the manner in which the
chemicals (NH4Cl, KH2PO4 and MgCl2) were dosed. As was described in section 5.2, the
chemicals were added at once at the sludge surface within the reactor. Locally this could
have caused a great over-dosage, and thus a substantial exceeding of the metastable limit, in
combination with possible insufficient mixing and possible eddies just below the sludge
surface. Besides this, the supersolubility curve that was used in this report, which is the upper
limit of the metastable zone, was adapted from research by Bhuiyan in ultra-pure water
(Bhuiyan et al., 2008). The supersolubility curve in digested sludge could very well differ from
the supersolubility curve in ultra-pure water. Therefore, the metastable zone is possibly
narrower than assumed in this report.
6.5
CO2 stripping
As was expected on the basis of theory, larger airflow rates resulted in a steeper pH increase,
as shown in figure 6.26. Converting the airflow rate to RQ values (L of applied air per L of
sludge) resulted in almost identical curves for different airflow rates (figure 6.27), suggesting
a fixed relation between RQ and pH for this experimental setup.
8,2
8,1
8
7,9
pH (-)
pH (-)
7,8
7,6
Qa = 500 L/h
7,4
7,7
Qa = 500 L/h
7,5
Qa = 750 L/h
Qa = 750 L/h
7,3
7,2
7,1
7
0
20
40
time (min)
60
80
Fig. 6.26 – pH increase at different airflow rates
(25L of sludge)
0
5
10
15
20
25
RQ (V air/V sludge)
Fig. 6.27 – pH as a function of RQ value
61
Using a Matlab model based on the equations in section 4.3, a curve was fitted through the
collected pH values in both the pilot and the experiments. The results are presented in figure
6.28 and 6.29. For comparability, pH was plotted against RQ using identical scales in both
figures. The Matlab model is included in Annex 6.
Fig. 6.28 – pH increase in the pilot
(without struvite formation)
Fig. 6.29 – pH increase in the experiments
(without struvite formation)
After iteratively adapting the KLa value (being a measure for the efficiency of the system), in
both the experiments and the pilot the model closely approached the observed pH value. As
the KLa value in the pilot was higher (0.001 s-1 versus 0.0006 s-1), the aeration efficiency in
this system was higher. This could be caused by a larger reactor height, or by a smaller
bubble size or higher turbulence in the pilot reactor. As the applied air takes up CO2 gas from
the sludge during its way up in the reactor, a higher reactor provides a longer contact time
between air and sludge and therefore a more efficient gas exchange (at the same RQ value).
Of course, at a certain point, the air is saturated with CO2, after which an enlargement in
reactor height will not give further improvement in efficiency. To investigate whether the
larger reactor height in the pilot as compared to the experiments (+/- 1.5 m versus +/- 0.8 m)
caused the better efficiency, a series of experiments was done with varying sludge volumes
(and therefore varying filling heights) while maintaining the same RQ value. The results are
shown in figure 6.30.
8
8,4
8,2
7,8
7,8
7,6
pH (-)
pH (-)
8
V=25L Qa=500L/h
7,4
V=15L Qa=300L/h
7,2
no dosage
7,6
Mg:PO4 = 1,5
7,4
7,2
V=10L Qa=200L/h
7
7
6,8
0
20
40
60
80
time (min)
Fig. 6.30 – pH increase at different filling heights
(at the same RQ value)
0
50
time (min)
100
Fig. 6.31 – Lower overall pH course due to
struvite formation (Qa = 750 L/h, V = 25L sludge)
Almost identical pH courses were found at different filling heights at the same RQ value. It
was concluded that the influence of reactor height on efficiency was negligibly small within
this range of height variation (about 0.3m to 0.8m). For this reason, it was concluded that
the large difference in efficiency between pilot and experiments was not caused by the
62
150
difference in reactor height, but rather by the difference in bubble size and turbulence. At the
same RQ value, smaller bubbles provide a larger total contact surface between air and sludge
and therefore a quicker gas exchange. If the bubbles are not completely saturated, a better
efficiency will be reached. Higher turbulence stimulates the renewal of the liquid/air interface
and therefore has a positive effect on the speed of reaction.
In figure 6.31 the pH increase for an experiment without any MgCl2 dosage is compared to
the pH increase after struvite formation (at Mg:PO4=1.5). Struvite was in this case formed
before t=0. The overall pH course was lower due to the lower initial value. The relative
increase, however, seemed to be independent of the initial pH value. In figure 6.32, the
collected pH values after struvite formation were plotted in Matlab and a curve was generated
using the same conditions as figure 6.29.
Fig. 6.32 – pH increase after struvite formation
Of course, the chemical composition of the sludge changes if MgCl2 is added and struvite and
other precipitates are formed. This was not taken into account in figure 6.32, since the same
chemical conditions (ionic strength, alkalinity) are used without MgCl2 addition. Despite of this,
the pH course generated with the Matlab model appears to provide a reasonable
approximation of the pH course that was measured. From this it was concluded that for
design purposes, the pH change due to aeration and due to struvite formation can be
considered independent.
63
64
7. Implementation at WWTP West
In both the experiments and the Waternet pilot it was found that struvite formation by MgCl2
addition is an effective process to improve sludge dewaterability and to reduce the
orthophosphate concentration. This section focuses on the specific situation at WWTP West.
Firstly, the objectives and constraints of implementing a phosphate recovery technique at
WWTP West are discussed. After that, the potential profits are estimated and some
alternatives are given. Finally, an overview is given of the steps that still need to be taken
before a phosphate recovery technique can be implemented at WWTP West.
7.1
Objectives
The previous sections outlined that struvite recovery from digested sludge can have several
advantages:
- An improvement in dewaterability of the digested sludge.
- The prevention of scaling problems in pipes, pumps and dewatering facilities.
- A reduction of phosphate recycling to the WWTPs inlet. This will lead to a lower
overall phosphate concentration and therefore to a lower load on the EBPR.
Eventually, this could make chemical dosages unnecessary, which are now often
applied to support the EBPR. Next to that, it could lower the phosphate concentration
of the plant effluent, thereby reducing the risk of eutrophication of the receiving
water.
- The production of struvite that could be sold as a slow-release fertilizer and that
could help in becoming less dependent on phosphorus rock globally.
As mentioned in section 3.1, scaling problems were the direct reason for Waternet to start a
project on struvite recovery. However, the other advantages cannot be neglected in the
eventual design process. At some point, other advantages might even outweigh the
prevention of scaling problems. Luckily, the advantages are not contradicting each other and
in many cases they are even strengthening each other. For example, the formation of more
struvite does not only lead to an increased struvite production, but also reduces scaling risks
and phosphate recycling to the WWTPs inlet.
Ultimately, the goal of implementing a phosphate recovery technique at WWTP West is twofold: 1) minimization of costs, and 2) maximization of social/environmental benefits.
At present, the treatment of municipal wastewater results in costs, which are paid for by the
producers of the wastewater (the citizens of a community) in the form of taxes. Naturally,
these costs are to be kept as low as possible. As all possible choices within the design process
could be expressed in terms of costs or profits, cost minimization is a very straightforward
and quantifiable design approach. It should be stressed, however, that gaining a complete
insight in the cost consequences of different alternatives is a complex task. An even more
complex situation occurs when social and environmental benefits are taken into account.
Section 1.2 explained that the depletion of phosphate rock reserves can lead to severe food
crises in the future. Therefore, the recovery of phosphate from alternative resources (such as
wastewater) is of public importance. To a lesser extent, further reduction of phosphate
emissions through the WWTPs effluent is desirable from an environmental point of view. The
impact of these social and environmental benefits is difficult to quantify and even more
difficult to express in terms of money.
In practice, the final design is a compromise between the minimization of cost and the
maximization of social/environmental benefits. Instead of finding an optimal solution, the final
design needs to meet a pre-defined set of objectives, against minimal costs.
65
Waternet formulated the objectives in an earlier internal report as follows:
- The Dry Solid Content (DSC) of the dewatered sludge must improve with at least 2%
(from 22% to 24%).
- The orthophosphate concentration in the centrifuge reject water must be lower than
50 mg/L PO4-P.
- The formed struvite must be usable as a fertilizer.
7.2
Constraints
The most important constraint for implementing a phosphate recovery technique at WWTP
West is the very limited available space. The treatment plant was built on an elongated
terrain that is enclosed by roads and a railway, as illustrated in figure 7.1.
Railway
USB
Future expansion of the
activated sludge process
Fig. 7.1 – Topview of WWTP West (google earth)
Considering the plant’s layout, and keeping in mind that long conduits must be avoided as
much as possible (for scaling reasons), the available space is limited to a small region
surrounding the USB (figure 7.2). The limited space especially influences the choice for a
struvite separation process. Gravity separation requires a low upward flow velocity and
therefore a large surface area. The next paragraph discusses this in more detail.
20.0 m
Space needed for
dewatered sludge
transportation
USB
Installations
Installations
20.0 m
terrain border
main road of the WWTP
Fig. 7.2 – Available space for struvite recovery
Other constraints are the sludge discharge and the sludge characteristics. These are
presented in table 7.1.
66
Table 7.1 – Digested sludge discharge and characteristics
parameter
value/range
unit
nominal discharge
2,000
m^3/day
maximum discharge
2,400
m^3/day
pH
7.2
temperature
30-32
ºC
PO4-P concentration 150-400
mg/L
NH4-N concentration 850-950
mg/L
Mg concentration
20-30
mg/L
DSC
3.5
%
7.3
Potential profits
In the experiments it was found that struvite had been formed, but had not been removed
due to small crystal sizes. When particles of considerable size are left in the sludge, they will
accumulate in the rest of the sludge line and/or will cause damage to the dewatering
centrifuges. However, below a certain size particles stay in suspension during the rest of the
sludge processing and can be centrifuged with the sludge without causing damage. The
maximum particle size that can be left in the sludge without problems is at the moment
unknown. From a cost perspective, it could be beneficial to leave struvite in the sludge. Small
average crystal sizes could be obtained by stimulating nucleation. A separation zone, washing
facilities and the transport of struvite could thereby become unnecessary. However, it was
found in the experiments that large sized struvite was already present in the ‘raw’ digested
sludge and had probably formed in the anaerobic digesters. Leaving this struvite in the sludge
will continue to lead to accumulation problems and problems in the centrifuges. Besides this,
the recovery of struvite can generate income when properly grown and separated. Also, the
removal of struvite from the sludge causes a reduction in dewatered sludge volume and
therefore a reduction in sludge disposal costs. This is illustrated in a rough cost calculation
(table 7.2).
Table 7.2 – Profits from struvite separation
profits
(Euro/year)
calculations
Income from selling struvite
Digested sludge: 2,000 m3/day Æ 730,000 m3/year
(a)
Struvite formation: 2.3 g/L Æ 1679 tons/year
(b)
Assumed struvite recovery: 75% Æ 1259 tons/year
Selling price struvite: 50 Euro/ton
63,000
(c)
Savings from a reduction in sludge volume
Costs of dewatered sludge disposal: 66 Euro/ton
(a)
Total
83,000
146,000
(a) adapted from an internal report by Waternet (‘Struvietverwijdering op RWZI West’)
(b) average struvite formation in the experiments at Mg:PO4=1.5 (table 6.4)
(c) selling price in Germany in 2009, adapted from a PCS brochure
In this calculation, possible reductions in maintenance costs are not included. This number
(146,000 Euro/year) can therefore be seen as a maximum budget for upgrading a process
that only focuses on better dewaterability and the prevention of scaling problems, to a
process that also recovers usable struvite. The profits from an improvement in dewaterability
and a reduction in maintenance costs exceed the possible profits from struvite separation. In
the Waternet pilot, it was found that the DSC of the dewatered sludge could be increased
from 22% to 25% (Annex 3). If this improvement in dewaterability is achieved while 75% of
the struvite is removed from the sludge, simple calculations demonstrate that approximately
645,000 Euro/year can be saved on sludge disposal costs (table 7.3).
67
The reduction in maintenance costs is more difficult to estimate. It was estimated by
Waternet that the potential savings on maintenance are 100,000 Euro/year.
Table 7.3 – Profits from improved dewaterability
profits
(Euro/year)
calculations
Current dewatered sludge production: 85,000 tons/year
(a)
Current DSC dewatered sludge: 22%
Potential DSC dewatered sludge: 25%
645,000
Decrease in dewatered sludge production due to increase in DSC: 10,200 tons/year
Increase in sludge production due to struvite (25% stays in the sludge): 420 tons/year
Potential reduction in dewatered sludge production: 9780 tons/year
Costs of dewatered sludge/struvite disposal: 66 Euro/ton
(a)
(a) adapted from an internal report by Waternet (‘Struvietverwijdering op RWZI West’)
Adding up the potential profits, a technique that is effectively improving dewaterability, while
preventing scaling problems and producing struvite, results in a benefit of 891,000
Euro/year (assuming a DSC of 25%). It should be stressed that this is a potential benefit: no
costs are taken into account. Building costs and operational costs of the implemented
technique have to be determined for each alternative separately. A complete cost calculation
is outside the scope of this report.
7.4
Alternatives
A variety of techniques exists to recover phosphate from wastewater streams (section 3).
However, currently only two techniques are known that recover phosphate directly from
digested sludge: AirPrex and a technique under development by Ebara Environmental
Engineering. Both techniques recover phosphate as struvite by adding a magnesium source
while controlling the pH of the digested sludge. The formed struvite is either separated by
gravity separation (AirPrex) or with a hydrocyclone (Ebara). Observing these techniques, it
seems that a struvite recovery process should at least contain the following components:
- A magnesium source.
- A reaction zone with sufficient mixing, to:
o Mix the applied magnesium with the sludge.
o Keep the (small) struvite crystals in suspension.
- A separation zone, where the formed struvite is separated from the sludge.
- Recycling of small crystals to allow growth.
- (pH adjustment)
‘pH adjustment’ is placed between brackets, since it was found in the experiments that even
at low pH values the majority of the available phosphate can be recovered as struvite (section
6.3). Therefore, pH adjustment is no inevitable requirement.
A magnesium source needs to be added to introduce free Mg2+ ions in the solution. There are
different magnesium sources available that have different advantages and disadvantages. In
the AirPrex technology, a MgCl2 solution (33%) was chosen for its ease in handling. The
MgCl2 solution is a homogenous, water-like fluid and can be easily stored, pumped and dosed.
At low temperatures (during winter) some crystallization within the MgCl2 storage could occur.
This can be prevented by placing the storage facility in a closed hall together with the struvite
reactor, which acts like a heater due to the high sludge temperature (+/- 30ºC). Another
alternative is the addition of MgO or Mg(OH)2. Both these chemicals have two advantages.
Firstly, they are binding H+ ions when they dissolute, thereby increasing the pH. This pH rise
could be sufficient, since it was found in the experiments that with only a slight pH increase
already good results could be obtained. Secondly, both MgO and Mg(OH)2 have a very low
solubility. For this reason, they are added as a slurry rather than as a solution. As the
dissolved Mg2+ ions are used up, more MgO or Mg(OH)2 will dissolve. Due to the slow
magnesium release, the use of MgO or Mg(OH)2 can prevent local over-dosages such as
68
occurred during the experiments (section 6.4). However, the slow reaction can also be seen
as a disadvantage, since a higher volume of the reaction zone could be needed. Next to these
‘pure’ magnesium sources, also a mixture of chemicals can be used. PCS developed the
‘MgPlus’ magnesium source, which is a blend of MgCl2 solution and some additives. PCS
claims that these additives have a positive influence on the formation of the crystal-structure
of struvite 17 . No quantitative research results on MgPlus are available at the moment. With
any magnesium source, the risk at local supersaturation could be reduced if extra attention is
paid to the manner in which the magnesium is added to the sludge. Very slow (continuous)
addition and high turbulence around the dosing point may prevent local over-dosages. High
turbulence could be reached if the magnesium is dosed in the sludge inlet pipe. The mixing
may be even further improved using static mixers; obstacles within the inlet pipe that force
the flow pattern to become turbulent.
Sufficient mixing in the reaction zone can be accomplished in different ways. Mixing by
aeration has proven to be feasible in the AirPrex technology. Besides that, mixing by aeration
has the advantage of pH increase due to CO2 stripping. Stirring, on the other hand, was
found to give the best dewatering result in the experiments (section 6.2). As sludge
dewaterability is the most important economical incentive (section 7.3), mixing by stirring
should be investigated in more detail. The size of the reaction zone and the needed mixing
energy are strongly dependent on the kind of magnesium source that is applied.
At the moment, only two different techniques for struvite/sludge separation have been
implemented: gravity settling (AirPrex) and separation with a hydrocyclone (Ebara). In a
hydrocyclone, centrifugal forces direct larger (heavier) particles to the wall, where they flow
down to the underflow exit. At the same time, smaller (lighter) particles leave the cyclone at
the overflow exit, together with the bulk of the fluid (Rietema, 1961). The great advantage of
using a hydrocyclone for separation is the very small footprint. On the other hand, a
hydrocyclone is much more sensitive to clogging and scaling problems than a simple gravity
settler. Also techniques that have not yet been implemented could be suitable for
struvite/sludge separation. For example, tilted plate (lamella) settlers need less space than
gravity settlers, and could be less sensitive to clogging and scaling than hydrocyclones.
If the recovery of struvite is aimed for, small crystals should be recycled as much as possible
to allow them to grow. In the current AirPrex process, all struvite from the sedimentation
zone is directly collected and washed (see figure 2.1). This seems strange, since recycling
these (small) crystals back to the reaction zone would stimulate crystal growth. The larger
the final crystal size, the easier the crystals can be separated and washed. Regardless the
applied separation technology, the separated struvite crystals could be led back to the
reaction zone. The final product could either be collected from the bottom of the reaction
zone, where higher turbulence only allows larger crystals to settle, or it could be collected
intermittently from the sedimentation zone. A special case of reaction/separation is the
application of a fluidized-bed reactor. If such a reactor increases in diameter over its height, it
is in fact a gravity settler placed directly on a reaction tank. The advantage of a fluidized-bed
reactor is the space that is saved by placing the separation zone on top of the reaction zone.
The height of the construction could be a disadvantage, as an extra pumping phase needs to
be avoided.
At last, the pH of the sludge could be increased in different ways. Aeration has the advantage
that no chemicals are needed and that the sludge is at the same time thoroughly mixed. The
dosing of MgO or Mg(OH)2 have the advantage that the pH is increased, while magnesium is
added. Of course, dosing another base, such as NaOH, can also raise the pH. This has the
advantage that pH, mixing and the addition of magnesium can be controlled separately. A
disadvantage is the risk of local supersaturation due to bad mixing of a base with the sludge.
Besides that, the addition of monovalent cations, such as Na+, could have a negative
influence on sludge dewaterability (section 4.1, ‘sludge dewatering’).
17
Source: www.pcs-consult.de
69
Combining the alternatives for magnesium addition, mixing, separation, recycling and pH
control leads to a variety of possible concepts for a struvite recovery installation. Some
concepts are presented in figure 7.3 – 7.6.
Digested sludge
to dewatering
Digested sludge
to dewatering
Reaction
Sedimentation
Digested sludge
Reaction
Digested sludge
Recycling of small
struvite crystals
MgCl2
MgCl2
Recycling of small
struvite crystals
Air
Air
Struvite
Struvite
Fig. 7.4 – Airlift reactor with a
hydrocyclone
Fig. 7.3 – Airlift reactor with gravity settling
Digested sludge
to dewatering
Sedimentation
Digested sludge
to dewatering
Reaction
Digested sludge
Sedimentation
Reaction
Digested sludge
Air
MgCl2
Recycling of small
struvite crystals
MgCl2
Struvite
Struvite
Fig. 7.5 – Fluidized bed reactor/airlift
reactor
Fig. 7.6 – Stirred reactor with tilted plate setttling
All these concepts could effectively recover struvite, while preventing scaling problems and
improving sludge dewaterability. Eventually, the choice for a certain concept will be
determined to a large extent by building costs, operational costs and robustness.
To determine if these concepts will fit the plant’s layout, some rough calculations can be
made. If a combination of aeration and MgCl2 dosage is chosen, the volume of the reaction
zone will be strongly determined by the time that is needed to reach the desired pH value. In
the experiments, the time needed for pH increase was about 2 hours at maximum. Assuming
a maximum discharge of 2,400 m3/day of digested sludge (table 7.1), the reactor zone should
have a volume of 200 m3. In that case, a cylindrical reactor with a height of 8 m and a
diameter of 6 m would suffice. The needed surface area for gravity settling is more complex
to determine. As estimation, the settling velocity of a spherical struvite particle (0.1 mm) is
considered. The settling velocity can be estimated using the Stokes’ equation [3]:
vp =
70
g ⋅ ( ρ p − ρ w ) ⋅ d p2
18 ⋅ μ
= 1.9075 ⋅10−4
m/s
(23)
ρp= density of (pure) struvite particle = 1700 kg/m3 (Heinzmann & Engel, 2007)
ρw= density of the sludge ≈ density of water = 1000 kg/m3
μ = dynamic viscosity of the sludge ≈ 0.02 Pa*s (at 30°C) (Shafei et al., 2005)
dp = diameter of the spherical particle = 0.0001 m
g = gravity acceleration = 9.81 m/s2
As the maximum up-flow velocity should be smaller than the settling velocity, the needed
surface area can be calculated from the maximum sludge discharge and the settling velocity.
Assuming a maximum discharge of 2,400 m3/day gives a needed surface area of 146 m2. In
that case, a cylindrical settling tank with a diameter of 6.8 m would suffice. Projecting
cylindrical tanks for reaction (diameter = 6 m) and settling (diameter = 6.8 m) on the plant
layout, it seems that a struvite recovery installation with aeration and gravity settling will
easily fit the available space at WWTP West (figure 7.7).
Airlift reactor
MgCl2
storage
Space needed for
dewatered sludge
transportation
USB
Installations
Installations
Gravity settler
20.0 m
terrain border
20.0 m
main road of the WWTP
Fig. 7.7 – Fitting of a struvite recovery installation at WWTP West
Another alternative is to adjust the existing USB (digested sludge buffer tank) in such a way
that it functions as a struvite reactor/separator. To investigate the possibilities of this
alternative, the current design of the USB is considered (figure 7.8-7.9).
In the current situation, digested sludge enters the USB just a few meters from the point
were it is eventually extracted, as can be seen in figures 7.8 and 7.9. Besides that, the
extraction of digested sludge takes place at the lowest point of the USB, to ensure that it can
be totally emptied. In section 6.4, it was explained that the majority of the struvite that is
currently causing problems at WTTP West has probably formed in the anaerobic digesters.
Since the sludge in the USB is just moderately stirred by two small mixers, the struvite is
likely to settle close to the point were it is introduced, hence close to the point where it is
extracted. This was confirmed when the USB was emptied during an extensive cleansing
operation in 2009. Large heaps of struvite sediments (figure 7.10) were found close to the
outlet pipes. Therefore, the current layout seems to encourage problems with struvite
accumulation such as the blockage of pipes and damage to pumps and dewatering
centrifuges.
71
External
digested sludge
Mixer
To centrifuges
A
A'
To centrifuges
From digesters
Mixer
Fig. 7.8 – horizontal cross section of the USB (at N.A.P.)
Max. sludge level +7850
Mixer
N.A.P.
From digesters
gradient = 1:25
To centrifuges
Fig. 7.9 – Longitudinal section of the USB (AA’)
72
Fig. 7.10 – Heaps of struvite sediments in the
USB 18
Fig. 7.11 – Outlet pipes in the USB18
As a quick improvement, the elbow-shaped outlet pipes (figure 7.11) could be rotated with
180º, thus pointing in upward direction. In this way the risk of blockage of the outlet pipe
could be significantly reduced. The quantity of struvite that is led to the dewatering
centrifuges would probably also reduce. However, this adjustment would also have some
disadvantages. The USB could not be completely emptied anymore by the regular pumps.
Besides that, the buffering capacity would reduce as a consequence of the higher positioning
of the outlet openings.
More extensive adjustments are needed if the USB is desired to function as a struvite reactor/
separator. In figure 7.12, a concept for adjusting the USB is presented. In this concept, a
‘struvite trap’ is constructed by moving the digested sludge inlet to the other side of the USB
and by building an overflow construction around the gutter of the outlet pipes. To increase
struvite crystallization, magnesium is dosed directly in the USB. In the experiments it was
found that aeration is not inevitably necessary for obtaining final orthophosphate
concentrations below 50 mg/L P (section 6.3). Therefore, repositioning the existing mixers (to
prevent them to interfere with the settled struvite) could suffice. In this concept, no facilities
are included to remove the formed struvite from the USB. Instead, the choice is made to
remove the formed struvite once a year during a cleansing operation. In section 7.3 it was
estimated that about 1260 tons of struvite would annually settle. Assuming the struvite
density to be 1,700 kg/m3 (Heinzmann & Engel, 2007) and considering the USB surface area,
this means that yearly about 3 m of struvite would settle at the USB bottom (marked blue in
figure 7.12). To take an unequal distribution of the settled struvite into account, the overflow
barrier is higher than 3 m (6 m in this concept). The reduction in buffering capacity due to
the presence of the overflow barrier could be (partly) compensated by increasing the height
of the reactor. In this concept, an increase in reactor height of 4 m is chosen.
Of course, this concept has some drawbacks. Most importantly, an expensive cleansing
operation is needed annually to remove the formed struvite. Furthermore, it is uncertain if no
struvite crystals will exit the struvite trap to cause scaling and accumulation problems in the
remainder of the sludge line. It is also uncertain if the existing mixers will provide enough
turbulence to effectively mix the sludge and the magnesium chloride. More research on
struvite crystal growth and separation would be needed to determine if adjusting the USB is a
realistic alternative.
18
Pictures were taken during the USB cleansing operation in 2005 by Waternet
73
Max. sludge level +11850
MgCl2
Mixer
N.A.P.
From digesters
To centrifuges
Fig. 7.12 – Adjustments to convert the USB into a struvite reactor/separator
As the struvite appears to crystallize spontaneously within the anaerobic digesters (section
6.4), the anaerobic digesters of a new WWTP could be designed in such a way that struvite
can be recovered directly from it. Because of the size and complexity of the digesters at
WWTP West, as well as the lack of available space around them, the adjustment of the
existing digesters is probably no realistic alternative.
Concluding, it should be underlined that a wide variety of alternatives exists for recovering
phosphate and for gaining preferable sludge dewatering conditions at WWTP West. For
example, phosphate could also be recovered by the addition of calcium or aluminium (Al3+).
Since these are also cations with multiple valences, it is likely that the same positive effect on
dewaterability could be reached as with the addition of magnesium. The addition of calcium
will result in crystalline phosphate precipitates that could be recovered in the same way as
struvite. The addition of aluminium will result in additional (chemical) sludge containing high
concentrations of phosphate. Eventually the phosphate can be recovered from the sludge ash
using methods as described in section 2 (ASH DEC and Sephos).
7.5
Next steps
The experiments in this report provide insight in the relationships between magnesium
dosage, pH, phosphate removal and sludge dewaterability. However, several questions about
struvite crystal growth and struvite separation remain unanswered. It is not yet clear what
mixing energy and retention time are needed for sufficient crystal growth. Also, the maximum
particle size that does not lead to accumulation in the remainder of the sludge line and that
does not damage the dewatering facilities is still unknown. After investigating these matters,
cost calculations can be made in order to decide whether struvite should be separated or
should be left within the sludge. To make this decision, the social benefits of struvite recovery
should be taken into account as well.
Different alternatives, such as the different reaction/separation concepts as presented in the
previous paragraphs, should be further elaborated and should be evaluated on performance
and costs. Also the alternative of adapting the USB should be considered. After comparing the
alternatives on performance and costs, the best alternative can be chosen. If the chosen
alternative fits the pre-defined design objectives, a design can be made and the technique
can be implemented at WWTP West.
74
8. Conclusions and recommendations
8.1
Conclusions
The main research question of this thesis was:
“What are the optimal process settings for recovering struvite by the addition of MgCl2 from
digested sludge at WWTP West, while optimizing sludge dewatering?”
Digested sludge conditions
About 1 g/L of sediments was found in the untreated digested sludge. The sediments seemed
to consist mainly of struvite. Also, some sand, small bits of wood and plant seeds (<5mm)
were found. The settling properties of the digested sludge proved to be very bad. After 3
days of settling, no sludge/water interface could be identified in the upper part of the column.
Dewaterability
Aeration without MgCl2 addition appeared to negatively influence the dewaterability of the
sludge. When MgCl2 was added in combination with sludge aeration, the dewaterability
improved without a significant influence of pH value. Dosing ratio’s above Mg:PO4 = 1.0 did
not further improve the dewaterability. After treatment, 80% of the original polymer dosage
led to the same dewaterability as before treatment. The best dewaterability result was found
at a slight magnesium over-dosage (Mg:PO4=1.2) while stirring the sludge instead of
aerating it.
Phosphate removal
The concentrations of orthophosphate, ammonium and magnesium were found to match the
thermodynamic equilibria as described in section 4.2, both before and after MgCl2 addition.
The struvite solubility product pKso that was derived from the experimental data (12.99) was
slightly lower than the pKso value found in literature (13.26 (Ali et al., 20005)). A strong
correlation was found between free PO43- and free Mg2+. It was concluded that assuming a
constant ammonium concentration is allowable within the concentration ranges that were
studied. With the removal ratios of magnesium and phosphate (∆CT_Mg/∆CT_PO4) found in
the experiments, a mathematic model was constructed to calculate the amount of MgCl2
needed to reach a certain final phosphate concentration at a specific pH.
Struvite recovery
From the pH drop after MgCl2 addition it was concluded that struvite reaction was completed
within 5 minutes in each experiment. No great differences in reaction speed were found for
different initial pH values or for different dosing ratios. After analyzing the sieve curves, the
mass of the precipitates and the solids content of the sludge, it was concluded that the
struvite formed in these experiments was too small in size to separate it from the sludge and
was lost during the washing procedure. The struvite that was collected after washing was for
the largest part probably already present in the raw digested sludge.
CO2 stripping
A relation between RQ value (Vair/Vsludge) and pH value was found for this experimental setup.
The aeration system used in the experiments was less efficient than the aeration system in
the Waternet pilot (Annex 3), probably due to smaller air bubbles in the pilot. The pH
increase can be very well approached with a Matlab model, based on the equations in section
4.3. Struvite formation decreased the overall pH course. Apparently, pH changes due to
aeration and due to struvite formation can be considered independent.
Concluding, the optimal process settings for struvite recovery were not found in this research,
since the formed struvite was too small to recover. Optimal sludge dewaterability in
combination with sufficient orthophosphate removal (final PO4-P<50mg/L) can be reached
using a slight magnesium over-dosage (Mg:PO4=1.2), while stirring the sludge instead of
aerating it.
75
8.2
Recommendations
The experiments in this report provide insight in the relationships between magnesium
dosage, pH, phosphate removal and sludge dewaterability. However, several questions about
struvite crystal growth and struvite separation remain unanswered. To make a balanced
choice for the implementation of a phosphate recovering technique at WWTP West, the
following questions should be addressed:
- What mixing energy and retention time are needed for sufficient crystal growth?
- How do different magnesium sources (MgCl2, MgO, Mg(OH)2, MgPlus) influence
crystal growth and which magnesium source is most suitable to use at WWTP West?
- How do different techniques of magnesium dosing influence crystal growth (e.g.
dosing in the inlet pipe, or directly in the reactor) and what is the most suitable
technique to use at WWTP West?
- Why does aeration negatively influence sludge dewaterability?
- What are the possibilities of stirring instead of aeration for the mixing of sludge and
chemicals? The results of the experiments suggest that this could have a positive
effect on sludge dewaterability.
- What is the maximum struvite particle size that does not lead to accumulation in the
remainder of the sludge line, or to damage to dewatering facilities?
- Which struvite/sludge separation technique is suitable to implement at WWTP West
and can separate struvite particles that exceed the maximum particle size from the
digested sludge?
- What are the possibilities of adapting the USB in such a way that struvite cannot
cause problems in pumps, pipes or dewatering facilities downstream?
- Could struvite be directly recovered from the USB?
- What are the performances and costs of different alternatives? Here, the social
benefit of recovering struvite should be taken into account.
Some general recommendations are made for future laboratory research on struvite recovery
from digested sludge:
- Use a CST (Capillary Suction Time) test instead of a gravity filtration test to
determine the dewaterability of the digested sludge. A CST test results in a single
value for sludge dewaterability and can be compared with values from literature.
Besides that, a CST test is less complex to perform and could therefore be more
accurate than a gravity filtration test.
- The Matlab model for component availability that was made during this research
(Annex 5) can be used to calculate the free ionic concentrations of magnesium,
ammonium and phosphate from the (measured) total concentrations.
- To investigate the purity of the formed struvite under different conditions, X-ray
diffraction (XRD) techniques can be used, or the precipitates can be pre-dissolved by
acid (like HCl) solution followed by element analysis (Hao et al., 2009).
Concluding, it is recommended to keep an open mind during the entire research/design
process. ‘Optimal’ solutions for both phosphate recovery and sludge dewaterability could be
found in other directions than readily available techniques such as AirPrex. For example, the
addition of other chemicals (like aluminium or calcium) could also improve the sludge
dewaterability, while producing useful phosphate precipitates.
76
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79
80
Annexes
Annex 1 – Process scheme water line WWTP West
Influent
(North/East cluster)
Influent
(South/West cluster)
Receiving
work
Coarse material
Coarse
screens
Terrain sewer (sludge line)
bypass
Dividing work 1
Primary
clarifier 1
Primary
clarifier 2
Primary
clarifier 3
Primary
clarifier 4
Primary sludge (sludge line)
FeCl3
Dividing work 2
(extension)
Anaerobic
tank 1
Denitrification
tank 1
Facultative
tank 1
...[Activated sludge installations 2-7]...
Nitrification
tank 1
Degassing
tank 1
return activated sludge
Surplus activated sludge (sludge line)
Secondary
clarifier 1
Secondary
clarifier 2
(extension)
Surface water
81
Annex 2 – Process scheme sludge line WWTP West
Primary sludge
WWTP Westpoort/CSI
Primary sludge WWTP West
Sand
capture 1
Sand
capture 2
Gravity
thickener 1
Gravity
thickener 2
Secondary sludge
WWTP West
Secondary sludge
WWTP Westpoort/CSI
Secondary sludge
buffertank
Belt
thickener 1
Belt
thickener 2
Belt
thickener 3
Belt
thickener 4
Belt
thickener 5
biogas
Biogas
buffer
Digestion
tank 1
Digestion
tank 2
Digestion
tank 3
Thickened sludge
CSI
Gas outlet
(flame)
Terrain sewer
(water line)
Digested
sludge
buffer
(USB)
AEB
Centrifuge
1
Centrifuge
2
Sludge
holding
tank 1
Dewatered sludge
82
Digested sludge
CSI
Centrifuge
3
Sludge
holding
tank 2
Centrifuge
4
Annex 3 – Summary of the Waternet pilot
Source: Veltman, A. (Waternet, 2011): Verwijdering van struviet op de rioolwaterzuivering
West – Fosfaatrecycling in de vorm van struviet.
Research motive and research goals
During a cleansing operation at WWTP West in 2009, 150 tons of crystal precipitate was
found at the bottom of the USB (digested sludge buffer tank). Analysis showed that this
precipitate consisted mainly of struvite. To investigate whether the AirPrex technology
(section 2.2) could prevent this excessive and unwanted precipitation in the future, a pilot
experiment was conducted in April 2010.
The pilot was typically set up as a “proof of principle” and aimed at the following results:
- An increase of the Dry Solid Content (DSC) of the dewatered sludge of 2 to 3
percentage points (from 22% to 24-25%).
- A concentration of orthophosphate in the dewatering reject stream below 50 mg/L
PO4-P.
Setup of the pilot
The pilot installation consisted of a 25 m3 reactor tank (figure A3.1) with a built-in aeration
system (figure A3.2), a MgCl2 dosing facility and a mobile dewatering centrifuge. In figure
A3.3, the pilot installation is presented schematically.
Fig. A3.1 – Reactor tank of the pilot installation
Fig. A3.2 – Aeration system inside the reactor tank
MgCl2
Polymer solution
Digested sludge
+/- 5 m3/h
1
Dewatered sludge
V = 25 m3
2
Centrifuge
3
4
Air: +/- 500 m3/h
Reject water
Fig. A3.3 – Schematical overview of the pilot installation with sampling points (1-4)
The digested sludge was diverted from the conduit between the anaerobic digesters and the
USB and was intermittently fed to the pilot reactor tank. When digested sludge was fed to the
83
reactor, MgCl2 was dosed simultaneously. Air was applied at a constant flow rate of
approximately 500 m3/h. The treated sludge was led to the mobile dewatering centrifuge at a
constant flow rate of 5 m3/h. The centrifuge was operated in the same way as the permanent
centrifuges at WWTP West: the polymer dosage was iteratively adjusted until an optimal
dewatering result was reached. No struvite was recovered, hence the formed struvite was
dewatered together with the sludge.
Within the pilot, four different experiments were performed:
1. A reference experiment, in which no air was supplied and no MgCl2 was dosed. The
aim of this experiment was to determine if the mobile dewatering centrifuge give
comparable dewatering results as the permanent centrifuges at WWTP West.
2. An experiment in which first air was applied without MgCl2 dosing until the pH
became stable. After that, the MgCl2 dosing was started. The aim of this experiment
was to determine what the separate effects of aeration and magnesium addition are
on phosphate removal and sludge dewaterability.
3. An experiment with constant aeration and MgCl2 dosing. The aim of this experiment
was to investigate how much time was needed for the process to become stable.
4. An experiment with a variable magnesium dosing ratio. The aim of this experiment
was to determine a suitable dosing ratio (Mg:PO4) for reaching satisfactory results.
As can be seen in figure A3.3, samples were taken from:
1. The digested sludge at the inlet of the reactor tank.
2. The treated digested sludge at the outlet of the reactor tank.
3. The dewatered sludge.
4. The reject water.
The sludge samples were analyzed on DSC, magnesium, phosphate, ammonium, chloride and
pH. The reject water was analyzed on Total Suspended Solids (TSS), magnesium, phosphate,
ammonium, chloride and pH.
Results
In table A3.1, the results of the reference experiment are compared to operational results
from WWTP West. From this data, it seems that the mobile dewatering centrifuge gives
comparable dewatering results as the permanent centrifuges at WWTP West.
Table A3.1 – Results of the reference experiment compared to operational results from
WWTP West
parameter
unit
pilot reactor
pH
temperature
DSC
TSS
orthophosphate
total P
ammonium
total N
magnesium
chloride
-
7.2
29.1
3.1
170
820
40.5
205
%
mg/L
mg/L P
mg/L P
mg/L N
mg/L N
mg/L
mg/L
WWTP West
pilot
digesters
dewatering installation
7.2
36
3.1
180
880
40.5
259
WWTP West
dewatering installation
7.7 (reject water)
7.6 (reject water)
28 (reject water)
30 (reject water)
22.2 (dewatered sludge) 22.4 (dewatered sludge)
240 (reject water)
300 (reject water)
150 (reject water)
160 (reject water)
170 (reject water)
185 (reject water)
690 (reject water)
680 (reject water)
23 (reject water)
240 (reject water)
-
The results from experiment 2 (first aeration until pH>, after that MgCl2 dosing) are given for
orthophosphate, pH and DSC in figure A3.4, A3.5 and A3.6, respectively. From these results,
it seems that aeration without magnesium addition slightly decreased the orthophosphate
concentration of the treated sludge. A further decrease was obtained after MgCl2 dosing was
started. The pH value increased gradually until a maximum is reached after approximately 2
hours of aeration without MgCl2 dosing. When MgCl2 dosing was started, the pH decreased
84
slightly. After that, the pH seemed to be stable at 8.0, apart from one different measured
value at t=13.20h (pH=7.8). The DSC of the dewatered sludge deteriorated when the sludge
was aerated without magnesium addition. After approximately 1 hour, the DSC seemed stable
at 20%. As soon as dosing of MgCl2 was started, the DSC improved. Within less than 1 hour,
the DSC reached values of above 25%. As soon as the MgCl2 dosing was stopped, the DSC
decreased to its original value (approximately 22%).
orthophosphate (mg/L P)
200
start MgCl2 dosing
10:40 h
180
160
start aeration
09:00 h
140
120
100
80
60
40
7:00
8:55
10:50
12:45
time
Fig. A3.4 – Orthophosphate concentration of the treated sludge during experiment 2
(sampling point 2)
8,2
8
start MgCl2 dosing
10:40 h
pH (-)
7,8
7,6
7,4
start aeration
09:00 h
7,2
7
7:00
8:55
10:50
time
12:45
Fig. A3.5 – pH of the treated sludge during experiment 2 (sampling point 2)
26
25
24
DSC (%)
23
stop MgCl2 dosing
13:20 h
start aeration
09:00 h
22
21
start MgCl2 dosing
10:40 h
20
19
18
17
7:00
8:55
10:50
12:45
time
Fig. A3.6 – DSC of the dewatered sludge during experiment 2 (sampling point 3)
85
The results of the stability experiment (experiment 3) are presented in figure A3.7, A3.8 and
A3.9. In this experiment, both the aeration and MgCl2 dosing were started at t=7.30h.
orthophosphate (mg/L P)
250
200
150
100
50
0
7:12
9:07
11:02
12:57
time
Fig. A3.7 – Orthophosphate concentration of the treated sludge during experiment 3
(sampling point 2)
8,4
8,2
pH (-)
8
7,8
7,6
7,4
7,2
7
7:12
8:24
9:36
10:48
12:00
13:12
14:24
15:36
time
Fig. A3.8 – pH of the treated sludge during experiment 3 (sampling point 2)
26,5
26
DSC (%)
25,5
25
24,5
24
23,5
7:12
8:24
9:36
10:48
12:00
13:12
14:24
15:36
time
Fig. A3.9 - DSC of the dewatered sludge during experiment 3 (sampling point 3)
From these results it was concluded that the both orthophosphate concentration and the pH
become stable after approximately 3 hours of operation, while the DSC of the dewatered
sludge becomes stable after approximately 5 hours of operation.
86
Figures A3.10 and A3.11 show the results of experiment 4, in which the magnesium dosing
ratio was varied. It seemed that the removal of orthophosphate was optimal at a slight
magnesium over-dosage (Mg:PO4=1.1). This is in contradiction with theory, since higher
dosing ratio’s should lead to higher supersaturation and therefore to a better phosphate
removal (section 4.2, ‘solubility and saturation’). No clear relation was found between
magnesium dosing ratio and the DSC of the dewatered sludge.
27
100
26
80
DSC (%)
orthophosphate (mg/L P)
120
60
40
25
24
23
20
22
0
0
0,5
1
1,5
2
2,5
Mg:PO4 (-)
Fig. A3.10 – Orthophosphate concentration of the
treated sludge at several Mg:PO4 dosing ratios
(sampling point 2)
0
0,5
1
1,5
2
2,5
Mg:PO4 (-)
Fig. A3.11 – DSC of the dewatered sludge at
several Mg:PO4 dosing ratios
(sampling point 3)
Conclusions and recommendations
From the pilot results it was concluded that the following goals could be reached using the
AirPrex technology at WWTP West:
- An increase of DSC of the dewatered sludge of 3 percentage points (from 22% to
25%).
- An orthophosphate concentration in the dewatering reject stream below 50 mg/L
PO4-P.
During the pilot, the struvite was not separated and the influence of pH and magnesium
dosage variations on the sludge dewaterability, the removal of phosphate and the struvite
crystal growth were not determined. It was recommended to investigate these matters in
more detail before implementing the AirPrex technology at WWTP West.
87
Annex 4 – Experimental data
Table A4.1 – Raw digested sludge characteristics
Exp. No.
pH
(-)
DSC
(%)
Temperature
(ºC)
PO4-P
(mg/L)
NH4-N
(mg/L)
Mg
(mg/L)
1
7.19
2
7.18
32
326
3
7.19
33.6
290
908
24
4
7.14
32.7
332
25.6
5
7.19
32.8
334
25
6
7.14
30.3
346
25.2
33.2
7
7.08
28.8
8
7.21
28.3
9
7
3.57
29.4
10
7.2
3.51
330
876
11
7.16
3.51
336
903
27
12
7.16
3.48
326
910
25.1
13
7.19
3.49
330
909
25.1
14
7.17
3.58
310
917
27.2
15
7.21
3.5
300
896
27.4
16
7.21
3.58
312
907
26.2
17
7.2
3.49
306
928
26.3
18
7.2
3.51
312
912
25.4
19
3.56
318
915
25.3
20
3.53
320
21
3.54
316
858
23.5
22
3.53
320
27.9
23
3.47
312
25.5
24
24.1
27.3
302
25
7.15
3.47
min
7.00
3.47
28.3
max
7.21
3.58
average
7.17
3.52
23.6
310
861
21.3
290
858
21.3
33.6
346
928
27.9
31.2
318
900
25.4
26
306
25.3
(Corresponds to table 5.1)
Table A4.2 – Pre dewaterability
exp no.
15
16
19
20
0
0
0
0
average
std dev
0
0
0
0.0
5
31
30
0.0
29
19
22
43
29.0
8.4
10
46
15
56
47
45
32
34
61
44.2
10.5
57
55
42
43
68
53.5
20
9.7
62
63
61
48
49
71
59.0
8.9
30
68
69
67
57
58
75
65.7
6.9
40
72
72
71
63
63
76
69.5
5.3
50
74
74
73
67
67
78
72.2
4.4
60
75
75
75
69
70
78
73.7
3.4
90
78
77
77
74
75
80
76.8
2.1
120
79
79
78
76
77
80
78.2
1.5
time (s)
17
18
percolated mass (g)
(Corresponds to figure 6.1)
88
Table A4.3 – Initial component concentrations
measured values
exp no.
10
pH (-) PO4-P (mg/L)
7.2
11
Mg (mg/L)
876
24.1
330
7.16
calculated with Matlab (see Annex 5)
NH4-N (mg/L)
336
903
PO4(3-) (mol/L) NH4(+) (mol/L) Mg(2+) (mol/L)
E
0.062
3.70 -04
E
0.064
4.20 -04
E
0.0645
3.97 -04
E
0.0644
3.88 -04
E
0.065
4.52 -04
1.07 -07
27
9.53 -08
12
7.16
326
910
25.1
9.28 -08
13
7.19
330
909
25.1
1.03 -07
14
7.17
310
917
27.2
9.03 -08
E
E
E
E
E
(Corresponds to figure 6.9)
Table A4.4 – Final component concentrations
measured values
calculated with Matlab (see Annex 5)
exp no.
pH (-)
PO4-P (mg/L)
NH4-N (mg/L)
Mg (mg/L)
10
6.96
38.05
811
189.5
PO4(3-) (mol/L) NH4(+) (mol/L) Mg(2+) (mol/L)
3.04E-09
0.0576
0.0072
11
7.18
35.95
793
192.5
5.27E-09
0.0562
0.0073
12
7.08
43.1
851
203
4.70E-09
0.0604
0.0076
13
7.28
20.2
841
190
3.82E-09
0.0595
0.0074
14
7.45
16.1
818
191
4.68E-09
0.0576
0.0075
15
7.6
12.8
853
142
6.42E-09
0.0597
0.0056
16
7.8
10.1
794
155
7.83E-09
0.055
0.0062
17
8
8.4
819
151
1.07E-08
0.0557
0.006
18
7.4
25.2
855
179
6.97E-09
0.0603
0.0069
19
7.19
25.7
885
332
2.70E-09
0.0627
0.0131
21
7.29
62.4
787
79.2
1.93E-08
0.0556
0.0026
22
7.46
137
880
27.9
9.07E-08
0.0619
6.43E-04
(Corresponds to figure 6.9)
Table A4.5 – Removal of magnesium compared to removal of phosphate
exp no.
initial pH (-) Mg:PO4 (-) initial PO4 (mol/L) initial Mg (mol/L) post PO4 (mol/L) post Mg (mol/L) deltaMg/deltaPO4 (-)
10
7.2
1.5
10.65547
15.98321
1.228608
7.79675
0.87
11
7.6
1.5
10.84921
16.27381
1.160801
7.920181
0.86
12
7.4
1.5
10.52632
15.78947
1.391669
8.352191
0.81
13
7.8
1.5
10.65547
15.98321
0.652244
7.817322
0.82
14
8
1.5
10.00969
15.01453
0.519858
7.858465
0.75
15
7.6
1.5
9.686794
14.53019
0.878269
7.200165
0.83
16
7.8
1.5
10.07427
15.1114
0.640943
8.475622
0.70
17
8
1.5
9.88053
14.82079
0.506942
7.69389
0.76
18
7.4
1.5
10.07427
15.1114
0.90733
8.311047
0.74
19
7.6
2
10.268
20.536
0.829835
13.65974
0.73
21
7.6
1
10.20342
10.20342
2.014853
3.258589
0.85
22
7.6
0.5
10.33258
5.16629
4.423636
1.147912
0.68
23
7.6
1.2
10.07427
12.08912
1.323862
4.155524
0.91
24
7.22
1.5
9.751372
14.62706
1.885696
7.488171
0.91
average
0.80
(Corresponds to tables 6.13-6.14)
Table A4.6 – Removal of ammonium
exp. no
10
11
12
13
14
15
16
17
18
19
20
21
22
NH4-N pre
mg/L N
876
903
910
909
917
896
907 928 912 915
858
NH4-N post
mg/L N
811
793
851
841
818
862
840 845 845 885
787 880
23
24 average
total average (pre+post)
903
838
870
89
Annex 5 – Matlab script for component availability and saturation
clear all
clc
% MODEL INPUT
dpH
pH(1)
pHfinal
K_w
K_MgOH
K_MgPO4
K_MgHPO4
K_MgH2PO4
K_NH4
K_HPO4
K_H2PO4
K_H3PO4
CT_PO4_mg
CT_NH4_mg
CT_Mg_mg
Mp
Mn
Mmg
A
(25°C)
IS
(constant)
Kso
product
=
=
=
=
=
=
=
=
=
=
=
=
=
=
=
=
=
=
=
0.1
7
9
10^-14
10^-2.56
10^-4.80
10^-2.91
10^-1.51
10^-9.25
10^-12.35
10^-7.20
10^-2.15
318
900
25.4
30.97
14.01
24.31
0.509
;
;
;
;
;
;
;
;
;
;
;
;
;
;
;
;
;
;
;
=
0.02
; % mol/L
=
10^-13.26
; % minimum struvite solubility
% CONVERSIONS
CT_PO4 = CT_PO4_mg/(1000*Mp)
concentration
CT_NH4 = CT_NH4_mg/(1000*Mn)
concentration
CT_Mg = CT_Mg_mg/(1000*Mmg)
concentration
%
%
%
%
%
%
%
%
%
%
%
%
%
%
%
%
%
%
%
pH step
initial pH
final pH
thermodynamic constant (25°C)
thermodynamic constant (25°C)
thermodynamic constant (25°C)
thermodynamic constant (25°C)
thermodynamic constant (25°C)
thermodynamic constant (25°C)
thermodynamic constant (25°C)
thermodynamic constant (25°C)
thermodynamic constant (25°C)
mg/L P total PO4 concentration
mg/L N total NH4 concentration
mg/L
total Mg concentration
g/mol
molar mass P
g/mol
molar mass N
g/mol
molar mass Mg
DeBye-Huckel constant
ionic strength
; % mol/L
initial orhtoP
; % mol/L
initial N
; % mol/L
initial Mg
% PRE CALCULATIONS
imax = ((pHfinal-pH(1))/dpH)+1;
% calculating number of pH steps
acNH4=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS); % activity coefficients
acNH3=1;
acPO4=10^-(A*3^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acHPO4=10^-(A*2^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acH2PO4=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acH3PO4=1;
acMg=10^-(A*2^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acMgOH=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acMgPO4=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acMgHPO4=1;
acMgH2PO4=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acH3O=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acOH=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
90
% LOOP
i=1;
% initializing loop
for i=1:imax
i
% H3O+
pH(i)=pH(1)+(i-1)*dpH;
H3O(i)=10^-(pH(i));
OH(i)=K_w/(H3O(i)*acH3O*acOH);
% NH4
NH4(i)=CT_NH4/(1+((K_NH4*acNH4)/(H3O(i)*acH3O*acNH3)));
NH3(i)=CT_NH4-NH4(i);
NH4frac(i)=NH4(i)/CT_NH4; % calculating fraction of CT_NH4 available
as NH4
NH3frac(i)=NH3(i)/CT_NH4;
% PO4 and Mg
res=solve('y=CT_PO4-((H3O(i)*y*acH3O*acPO4)/(K_HPO4*acHPO4))((H3O(i)^2*y*acH3O^2*acPO4*acHPO4)/(K_H2PO4*acH2PO4*K_HPO4*acHPO4))((H3O(i)^3*y*acH3O^3*acPO4*acHPO4*acH2PO4)/(K_H3PO4*acH3PO4*K_H2PO4*a
cH2PO4*K_HPO4*acHPO4))CT_Mg+(CT_Mg/(1+((OH(i)*acMg*acOH)/(K_MgOH*acMgOH))+((y*acMg*acPO4)/(
K_MgPO4*acMgPO4))+((H3O(i)*y*acH3O*acPO4*acMg*acHPO4)/(K_MgHPO4*acMgH
PO4*K_HPO4*acHPO4))+((H3O(i)^2*y*acH3O^2*acPO4*acHPO4*acMg*acH2PO4)/(
K_MgH2PO4*acMgH2PO4*K_H2PO4*acH2PO4*K_HPO4*acHPO4))))+((CT_Mg/(1+((OH
(i)*acMg*acOH)/(K_MgOH*acMgOH))+((y*acMg*acPO4)/(K_MgPO4*acMgPO4))+((
H3O(i)*y*acH3O*acPO4*acMg*acHPO4)/(K_MgHPO4*acMgHPO4*K_HPO4*acHPO4))+
((H3O(i)^2*y*acH3O^2*acPO4*acHPO4*acMg*acH2PO4)/(K_MgH2PO4*acMgH2PO4*
K_H2PO4*acH2PO4*K_HPO4*acHPO4))))*OH(i)*acMg*acOH)/(K_MgOH*acMgOH)');
sol=eval(res);
PO4(i)=double(sol(2));
% calculating PO4 concentration (implicit)
HPO4(i)=(H3O(i)*PO4(i)*acH3O*acPO4)/(K_HPO4*acHPO4);
H2PO4(i)=(H3O(i)^2*PO4(i)*acH3O^2*acPO4*acHPO4)/...
(K_HPO4*acHPO4*K_H2PO4*acH2PO4);
H3PO4(i)=(H3O(i)^3*PO4(i)*acH3O^3*acPO4*acHPO4*acH2PO4)/...
(K_HPO4*acHPO4*K_H2PO4*acH2PO4*K_H3PO4*acH3PO4);
Mg(i)=CT_Mg/(1+((OH(i)*acMg*acOH)/(K_MgOH*acMgOH))+...
((PO4(i)*acMg*acPO4)/(K_MgPO4*acMgPO4))+((H3O(i)*...
PO4(i)*acMg*acHPO4*acH3O*acPO4)/(K_MgHPO4*acMgHPO4*...
K_HPO4*acHPO4))+((H3O(i)^2*PO4(i)*acH3O^2*acPO4*...
acHPO4*acMg*acH2PO4)/(K_MgH2PO4*acMgH2PO4*K_HPO4*...
acHPO4*K_H2PO4*acH2PO4)));
MgOH(i)=(Mg(i)*OH(i)*acMg*acOH)/(K_MgOH*acMgOH);
MgPO4(i)=(Mg(i)*PO4(i)*acMg*acPO4)/(K_MgPO4*acMgPO4);
MgHPO4(i)=(Mg(i)*acMg*acHPO4*H3O(i)*PO4(i)*acH3O*acPO4)/...
(K_MgHPO4*acMgHPO4*K_HPO4*acHPO4);
MgH2PO4(i)=(Mg(i)*H3O(i)^2*PO4(i)*acH3O^2*acPO4*acHPO4*...
acMg*acH2PO4)/(K_MgH2PO4*acMgH2PO4*K_HPO4*acHPO4*...
K_H2PO4*acH2PO4);
PO4frac(i)=PO4(i)/CT_PO4;
HPO4frac(i)=HPO4(i)/CT_PO4;
H2PO4frac(i)=H2PO4(i)/CT_PO4;
H3PO4frac(i)=H3PO4(i)/CT_PO4;
Mgfrac(i)=Mg(i)/CT_Mg;
MgOHfrac(i)=MgOH(i)/CT_Mg;
MgPO4mgfrac(i)=MgPO4(i)/CT_Mg;
MgPO4pfrac(i)=MgPO4(i)/CT_PO4;
MgHPO4mgfrac(i)=MgHPO4(i)/CT_Mg;
MgHPO4pfrac(i)=MgHPO4(i)/CT_PO4;
MgH2PO4mgfrac(i)=MgH2PO4(i)/CT_Mg;
MgH2PO4pfrac(i)=MgH2PO4(i)/CT_PO4;
IAP(i)=PO4(i)*acPO4*Mg(i)*acMg*NH4(i)*acNH4;
91
Sc(i)=(IAP(i)/Kso)^(1/3);
Sr(i)=Sc(i)-1;
Z(i)=0;
if Sr(i)<=0;
S(i)=0;
else
S(i)=Sc(i)-1;
end
end
%
%
%
%
calculating supersaturation ratio
calculating relative supersaturation
zero line (for figure)
calculating thermodynamic driving force
% PLOTTING THE RESULTS
figure
plot(pH,PO4frac,'r-',pH,HPO4frac,'b-',pH,H2PO4frac,'g',pH,H3PO4frac,'m-',pH,MgPO4pfrac,'b-.',pH,MgHPO4pfrac,'r.',pH,MgH2PO4pfrac,'g-.')
xlabel(['pH'])
ylabel(['-'])
legend('PO4','HPO4','H2PO4','H3PO4','MgPO4','MgHPO4','MgH2PO4')
ylim([0 1])
title(['distribution of CT_P_O_4 over several PO4-complexes'])
figure
plot(pH,NH4frac,'r-',pH,NH3frac,'-')
xlabel(['pH'])
ylabel(['-'])
legend('NH4','NH3')
ylim([0 1])
title(['distribution of CT_N_H_4 over NH4 and NH3'])
figure
plot(pH,Mgfrac,'r-',pH,MgOHfrac,'b-',pH,MgPO4mgfrac,'r.',pH,MgHPO4mgfrac,'b-.',pH,MgH2PO4mgfrac,'g-.')
xlabel(['pH'])
ylabel(['-'])
legend('Mg','MgOH','MgPO4','MgHPO4','MgH2PO4')
ylim([0 1])
title(['distribution of CT_M_g over several Mg-complexes'])
figure
plot(pH,Sr,'r-',pH,Z,'k.')
xlabel(['pH'])
ylabel(['Sr'])
title(['relative supersaturation as a function of pH'])
figure
plot(pH,S)
xlabel(['pH'])
ylabel(['S'])
title(['thermodynamic driving force as a function of pH'])
92
Annex 6 – Matlab script for pH increase due to CO2 stripping
clear all
clc
% MODEL INPUT
dt
tend
pH(1)
m
CO2s
kLA
K_1
K_w
A
IS
Qa
V
=
=
=
=
=
=
=
=
=
=
=
=
600
7200
7.2
70e-3
1.09e-3
0.0010
4.5e-7
10^-14
0.509
0.02
500
25
;
;
;
;
;
;
;
;
;
;
;
;
%
%
%
%
%
%
%
%
%
%
%
%
s
size of timestep
s
duration of experiment
initial pH
mol/L
alkalinity
mol/L
equil CO2 concentration
s^-1
gas transfer coefficient
thermodynamic constant (25°C)
thermodynamic constant (25°C)
DeBye-Huckel constant (25°C)
mol/L
ionic strength
L/h
Aeration rate
L
reactor volume
% ACTIVITY COEFFICIENTS
acH3O=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acOH=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
acCO2=1;
acHCO3=10^-(A*1^2*(IS^0.5/(1+IS^0.5))-0.3*IS);
% PRE CALCULATIONS
H3O(1)=10^-(pH(1));
% calculating initial H3O+
concentration
OH(1)=K_w/(H3O(1)*acH3O*acOH); % calculating initial OHconcentration
HCO3(1)=m-OH(1)+H3O(1);
% calculating initial HCO3concentration
CO2(1)=(H3O(1)*acH3O*HCO3(1)*acHCO3)/(K_1*acCO2);
% calculating
initial CO2 concentration
CT_C(1)=CO2(1)+HCO3(1);
% calculating initial total anorganic
C concentration
imax = (tend/dt)+1;
% calculating number of timesteps
% LOOP
i=1;
% initializing loop
for i=1:imax
i
t(i)=(i*dt)-dt;
RQ(i)=((t/3600)*Qa)/V;
res=solve('m=(CT_C(i)/(1+((acHCO3*y*acH3O)/(acCO2*K_1))))+(K_w/(y*acH
3O*acOH))-y');
sol=eval(res);
H3O(i)=double(sol(1));
OH(i)=K_w/(H3O(i)*acH3O*acOH);
HCO3(i)=m-OH(i)+H3O(i);
CO2(i)=CT_C(i)-HCO3(i);
pH(i)=-log10(H3O(i));
% CO2 in next timestep
if i<imax;
CT_C(i+1)=CT_C(i)+kLA*(CO2s-CO2(i))*dt;
% calculating new total
anorganic C concentration
end
end
93
figure
plot(RQ,pH)
xlabel(['V_a_i_r/V_s_l_u_d_g_e'])
ylabel(['pH'])
title(['K_L_A=0.0010, IS=0.02mmol/L, Alkalinity=70mmol/L'])
hold on
% EXPERIMENT: 500 L/h of air at 25L of sludge, no struvite formation
%x=[0 1.67 3.33 5 6.67 8.33 10 13.33 16.67 20 26.67 33.33 40]; %RQ
%y=[7.19 7.28 7.37 7.44 7.5 7.57 7.63 7.72 7.8 7.87 7.98 8.07 8.06];
%pH
%scatter(x,y)
%PILOT: 500 m3/h or air at 25m3 of sludge, no struvite formation
%x=[0 10 20 35]; %RQ
%y=[7.2 7.9 8.03 8.07]; %pH
%scatter(x,y)
% EXPERIMENT: 750 L/h of air at 25L of sludge, struvite formation
completed before t=0
%x=[0 2.5 5 7.5 10 12.5 15 20 26.5 30 40 50 60]; %RQ
%y=[6.84 6.93 7.07 7.21 7.34 7.45 7.53 7.65 7.76 7.8 7.93 8.04 8.12];
%pH
%scatter(x,y)
94
!"
#$##