Recovery of Phosphorus from HTC Converted Municipal Sewage Sludge

Recovery of Phosphorus from HTC Converted Municipal Sewage Sludge

Matilda Sirén Ehrnström

Sustainable Process Engineering, masters level 2016

Luleå University of Technology

Department of Civil, Environmental and Natural Resources Engineering

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MASTER OF SCIENCE THESIS

R ECOVERY OF P HOSPHORUS FROM HTC

CONVERTED M UNICIPAL S EWAGE S LUDGE

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HTC-

Matilda Sirén Ehrnström 2016

Master of Science in Engineering Technology Sustainable Process Engineering

SUPERVISORS

Fredrik Öhman, C-Green Technology AB Fredrik Lundqvist, C-Green Technology AB

Lars Gunneriusson, Luleå University of Technology EXAMINER

Lars Gunneriusson, Luleå University of Technology

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I would like to start by thanking my main supervisor at C-Green Technology AB, Fredrik Öhman, for giving me the amazing opportunity to do my Master’s thesis project at the company.

I would also like to thank my second supervisor Fredrik Lundqvist for bringing new perspectives into the subjects and helping me sort them out. Thank you both for all your support, feedback and fruitful discussions throughout – sometimes leaving me with even more questions.

Thanks to my examiner Assoc. Prof. Lars Gunneriusson at Luleå University of Technology.

I am so very grateful to Lars-Erik Åkerlund, my practical supervisor, who have answered all my questions (relevant or not), and found solutions to every practical issue I have encountered.

Your support, knowledge and inspiring music have been invaluable during this project. I am also very thankful to Erik Odén at C-Green for your commitment that makes it all possible.

IVL Swedish Environmental Research Institute, and especially Mila Harding and Christian Baresel at Sjöstadsverket, are gratefully acknowledge for providing analytical instruments and support.

Special thanks to Helena Giers and Karin Lind at Stockholm Vatten AB who helped me get in contact with C-Green.

Slutligen vill jag tacka mina föräldrar som alltid backat upp mig, hjälpt och stöttat mig så att jag vågat anta nya utmaningar! Och min djupaste tacksamhet till Leo som gett mig så mycket kunskap om hur processer, och livet, verkligen fungerar.

Matilda Sirén Ehrnström Stockholm, August 2016

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With a growing population but scarce primary phosphorus sources, recycling of the vital element has become an important research area throughout the last decades. Several streams in society are potential resources for recirculation but municipal sewage is considered one of the most available materials. With current technologies in wastewater treatment, over 95 % of the influent phosphorus is captured in the sludge along with a variety of other nutrients. However, due to increasing fractions of pharmaceutical residues and heavy metals also following the sludge, direct use as fertiliser is being phased out in most European countries in favour of extraction methods. Extraction of nutrients from the sludge is problematic mainly because of dewaterability difficulties. Thus, pretreatment of the material is required to access the desired components at a reasonable cost and energy consumption. Hydrothermal carbonisation (HTC) is a technology showing high potential for treatment of wet carbonaceous material without necessity of prior drying. The resulting product is hygenised, essentially free from pharmaceuticals and easily dewatered.

In this Master’s thesis principal conditions for release of phosphorus from HTC converted digested sludge under acid leaching have been experimentally investigated. Dependence of time, temperature, dry solids (DS) content of HTC sludge and pH have been studied. Also, differences arising from acid type have been considered by comparing acidulation with sulphuric acid and hydrochloric acid. A short investigation of the recovery of the dissolved phosphorus from leachate by precipitation was also performed where calcium ions were added to both sulphuric and hydrochloric acid leachates.

Extraction of phosphorus from HTC converted sludge has shown to be easier than from pure metal phosphates under comparable leaching conditions and pH values. Also, the dissolved phosphorus concentrations obtained in the presence of HTC converted sludge was higher than for theoretical equilibrium concentrations where all phosphorus is in the form of iron(III) or aluminium(III) phosphate. A maximum leachate phosphorus concentration was around 2500 mg/L, recorded in leaching experiments performed at a dry HTC product concentration of 10

% (w/w) in an extraction solution of water acidified with sulphuric acid. Leaching procedures performed at pH values between 2 and 1 with 1 and 5 % DS HTC product resulted in dissolution of 90 % of ingoing phosphorus at an acid charge of 0.5 kg H2SO4/kg DS HTC product. At this chemical charge, release of phosphorus from converted sludge is fast. Similar amounts of dissolved phosphorus were recorded after 15 min as after 16 h retention time. Possibly, time

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dependence becomes relevant at lower charges. The dissolution of phosphorus is negatively affected by temperature increases at moderate acid loads, and by possibly by hydrochloric acid at pH values below 2.

Addition of calcium gave a dissolved phosphorus reduction of 99.9 % in both the sulphuric acid and hydrochloric acid leachates. Gypsum, CaSO4, also precipitates from the sulphuric acid leachate resulting in 67 % more dry mass. Due to high release of metals during acidulation, the precipitate was also contaminated with large fractions of metals in addition to calcium.

In summary, this investigation has demonstrated that up to 90 % of the phosphorus content of the HTC converted sludge can be released by acid leaching, and almost 100 % of the phosphorus can be recovered from the leachate by precipitation with calcium ions.

Key words: phosphorus leaching, acid leaching, phosphorus recovery, sewage sludge, clean sludge, hydrothermal carbonisation, HTC.

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Med en växande världspopulation och begränsade primära fosforresurser har forskningen kring återvinning av det livsnödvändiga grundämnet ökat under de senaste årtiondena. Potentialen för recirkulation från många olika källor har genom åren utvärderats men kommunala avloppsvatten anses vara en av de mest lättillgängliga resurserna. Med dagens teknologi inom vattenrening kan över 95 % av den till reningsverket inkommande fosforn fångas i avloppsslammet tillsammans med flera andra näringsämnen. Materialets direkta användning som gödsel fasas dock ut i många europeiska länder på grund av stigande halter av läkemedelsrester och tungmetaller som även följer med slammet. Istället satsas det på utveckling av metoder för extraktion av de betydelsefulla näringsämnena. Separationen är däremot inte helt oproblematisk till följd av slammets dåliga avvattningsegenskaper.

Förbehandling av materialet krävs därför för att göra de önskade ämnena tillgängliga till en rimlig kostnad och energiförbrukning. För behandling av vått kolrikt material har hydrotermisk karbonisering (HTC) visat stor potential utan krav på föreliggande torkningsprocess. Produkten som fås är hygieniserad, nästintill fri från läkemedelsrester och lättavvattnad.

I detta examensarbete har de grundläggande betingelserna för upplösning av fosfor från HTC behandlat rötat avloppsslam undersökts experimentellt under sura förhållanden. Betydelsen av ett antal parametrar så som tid, temperatur, torrsubstans (TS) av HTC-behandlat slam och pH har studerats. Eventuella skillnader mellan olika syror har också tagits med genom lakning med svavelsyra samt saltsyra. Utvinning av löst fosfor från surgjord lakvätska genom fällning undersöktes även i två experiment. Tillsats av kalciumjoner till lakvätskor surgjorda med svavelsyra respektive saltsyra möjliggjorde identifiering av eventuella skillnader även i detta steg.

Experimenten har visat att det är betydligt lättade att lösa upp fosfor från HTC-behandlat slam än från rena metallfosfat under liknande lakförhållanden och pH-värden. Koncentrationerna av fosfor som erhållits under lakning av HTC-behandlat slam ligger högt över de teoretiska jämviktskoncentrationerna där all fosfor föreligger som järn(III)- eller aluminium(III)fosfat. En maximal fosforkoncentration i lakvätskan tros dock ha nåtts runt 2500 mg/L vilket uppmätts i lakexperiment vid 10 % TS HTC-produkt i lösningsmedel av svavelsyra och vatten. För lakförsök utförda mellan pH 2 och 1 vid 1 och 5 % TS HTC-produkt har 90 % av ingående fosfor lakats ut med en svavelsyrasatsning på 0.5 kg H2SO4/kg TS HTC-produkt. Vid denna satsning har upplösningen av fosfor från HTC-behandlat slam visat sig vara snabb då lika

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mycket material har lösts upp under 15 min uppehållstid som under 16 h. Möjligen kan lakprocessen vara tidsberoende vid lägre syrasatsningar. Upplösningen av fosfor är något sämre vid högre temperatur och måttliga satsningar, liksom vid lakning med saltsyra vid pH-värden under 2.0.

Genom tillsats av kalciumjoner uppgick fosforreduktionen till 99.9 % i båda lakvätskorna surgjorda med svavelsyra respektive saltsyra. Mängden fällning från lakvätskan med svavelsyra var 67 % högre än i fallet med saltsyra vilket mest troligt beror på utfällning av gips, CaSO4. Under lakningen löstes utöver fosfor även tungmetaller ut, vilka i stor utsträckning följde med och förorenade fällningen från lakväskan utöver kalcium.

Sammanfattningsvis har den här studien visat att upp till 90 % av fosforinnehållet i det HTC- behandlade slammet gått att lösa upp genom lakning under sura förhållanden. Dessutom går nästan 100 % av fosforn att utvinna från lakvätskan genom fällning med kalciumjoner.

Nyckelord: lakning av fosfor, fosforutvinning, avloppsslam, rent slam, hydrotermisk karbonisering, HTC.

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1 INTRODUCTION ... 1

1.1 C-GREEN TECHNOLOGY AB ... 1

1.2 SCOPE ... 2

1.2.1 Objectives ... 2

1.2.2 Delimitations ... 2

2 LITERATURE REVIEW ... 3

2.1 PHOSPHORUS ... 3

2.2 WASTEWATER TREATMENT AND PHOSPHOROUS REMOVAL ... 5

2.2.1 Mechanical cleaning ... 6

2.2.2 Chemical cleaning ... 6

2.2.3 Biological cleaning ... 7

2.3 SLUDGE FROM WASTEWATER TREATMENT PLANTS... 8

2.4 SLUDGE MANAGEMENT AND REQUIREMENTS ... 10

2.5 TECHNOLOGIES FOR PHOSPHOROUS RECOVERY ... 11

2.5.1 Recovery of dissolved phosphorus in liquid ... 12

2.5.2 Phosphorous release and recovery from sewage sludge ... 12

2.6 HYDROTHERMAL CARBONISATION ... 14

2.7 LEACHING ... 15

2.7.1 Mass transfer in leaching and rate determining step ... 15

2.7.2 Parameters affecting a leaching operation... 16

2.7.3 Leaching of phosphorus from HTC converted sludge ... 18

2.8 RECOVERY OF PHOSPHORUS FROM LEACHATE ... 23

3 METHOD ... 25

3.1 MATERIALS ... 25

3.2 EXPERIMENTAL WORK ... 26

3.2.1 Preparation of HTC converted sludge ... 27

3.2.2 Leaching procedure ... 27

3.2.3 Precipitation procedure ... 30

3.3 ANALYSIS ... 30

3.3.1 Analysis of phosphorus ... 31

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4 RESULTS AND DISCUSSION ... 32

4.1 COMPOSITION OF RAW MATERIALS ... 32

4.2 RELEASE OF PHOSPHORUS FROM HTC CONVERTED SLUDGE ... 33

4.2.1 Retention time ... 33

4.2.2 Temperature ... 35

4.2.3 Dry solid content and sulphuric acid charge ... 37

4.2.4 Acid type ... 41

4.3 RELEASE OF PHOSPHORUS FROM PURE METAL PHOSPHATES ... 43

4.4 ELEMENTAL DISTRIBUTION ... 44

4.5 ASH CONTENT OF LEACHED MATERIALS ... 48

4.6 RECOVERY OF PHOSPHORUS FROM ACID LEACHATE ... 51

4.7 FURTHER DISCUSSION ... 55

4.7.1 Acid consumption ... 55

4.7.2 Material variations and changes ... 56

5 CONCLUSIONS ... 58

6 FUTURE WORK ... 60

7 REFERENCES ... 61

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Figure 1. The geological (long-term inorganic) and biological (short-term organic) cycles of phosphorus on earth including the human impact (Cornel & Schaum, 2009). ... 4 Figure 2. Phosphorous cycle diagram in water bodies (Correll, 1998). ... 5 Figure 3. Biological phosphorous removal under anaerobic and aerobic conditions

(Balmér, et al., 2007, modified). ... 8 Figure 4. Solubilities of metal phosphates at varying pH (Stumm & Morgan, 1996). ... 19 Figure 5. pH diagram over the phosphoric acid system. ... 19 Figure 6. Activity coefficients for ions in water solution according to the extended Debye-

Hückel equation (Snoeyink & Jenkins, 1980). ... 22 Figure 7. Overview of the process studied in this project; from WWTP to phosphorous

containing precipitate via HTC conversion and leaching. ... 26 Figure 8. Retention time dependence in phosphorous leaching. ... 34 Figure 9. Temperature dependence in phosphorous leaching. ... 35 Figure 10. Influence of DS content on phosphorus leaching at different pH values resulting

from different acid charges. ... 37 Figure 11. Concentration of dissolved phosphorus at different DS contents and pH values. .. 40 Figure 12. Comparison between phosphorous leaching with sulphuric acid and

hydrochloric acid. ... 42 Figure 13. Leaching of pure metal phosphates at different pH values. ... 43 Figure 14. Ash content of remaining solid fraction on dry basis for experimental series C

through F. ... 48 Figure 15. Ash contents of solid fraction after leaching with sulphuric acid at 1 % DS HTC

product. ... 51 Figure 16. Ash contents of solid fraction after leaching with hydrochloric acid at 5 % DS

HTC product. ... 51 Figure 17. Leaching efficiency for experimental series C to F. Each symbol corresponds to

an acid charge from 0.1 to 0.5 kg H2SO4/kg DS HTC product. ... 56

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Table 1. Salts, formulas and solubility product constants at 25 ^oC. ... 24

Table 2. Used chemicals, formula, grade and assay. ... 25

Table 3. Notation of the performed sets of experiments. ... 27

Table 4. Experimental conditions for investigation of retention time. ... 29

Table 5. Experimental conditions for investigation of chemical charge. ... 29

Table 6. DS and ash content of dewatered sludge and HTC product. ... 32

Table 7. Elemental composition of dewatered sludge and HTC product. ... 33

Table 8. Distribution of elements between solid and liquid phase for different DS HTC product concentrations, temperatures and acids. ... 46

Table 9. Properties of leachates used in the precipitation experiments. ... 51

Table 10. Composition of original leachates and the distribution of elements between filtrate and precipitate after precipitation of sulphuric acid and hydrochloric acid leachates. ... 53

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Phosphorus is a finite resource but a vital element in all plants, organisms and animals. The primary phosphorus sources are scarce and predicted to last a few decades. (Cornel & Schaum, 2009) At the same time, excessive concentrations of phosphorus cause eutrophication in water bodies worldwide (Correll, 1998). Phosphorous enrichment in fresh and coastal waters is a result of human activities; industry, agriculture, sewage disposal (Chislock, et al., 2013) and land-use changes (Smith & Schindler, 2008).

Sewage disposal has been considered a problem for a long time (Stark, n.d.) (Tideström, et al., 2007). Even in countries where wastewater treatment is regulated in law, sludge is problematic due to contamination with heavy metals, pathogens and pharmaceuticals. If these parameters were the only conditions to be considered, landfilling or incineration ought to be the only methods to discard the material. Controversially, the material also carries significant amounts of carbon, nitrogen and phosphorus which argues against landfilling. Thus, spreading of sludge on forest- and farmlands, as well as using it for soil improvement and restoration in open-pit mines are also possible ways to dispose the sewage sludge.

As a result of unsustainable sludge handling and scarce phosphorus resources worldwide, methods for recovering the element from the problematic material has been an important research area in almost twenty years. Numerous of technologies have been developed to recover as much of the resources as possible, while trying to meet future demands and regulations.

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C-Green Technology AB is a newly established company developing an environmentally and economically sustainable process for transformation of sewage sludge into a biofuel. The C- Green process is modularised and designed to fit into a standardised container unit which enables placement at any wastewater treatment plant despite space limitations. Each container has a capacity to treat sludge from up to 200 000 individuals, which corresponds to approximately 30 000 tonnes of wet sludge per year.

The technology is based on hydrothermal carbonisation (HTC) performed at about 200 ^oC followed by efficient heat recovery and mechanical separation of bio-coal from remaining process water. The patented and continuous reactor guarantees a certain residence time and enables separation of a particle-rich and particle-lean phase at reaction temperature. A pilot

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plant of the C-Green process is currently under construction and will be ready for start-up during autumn 2016.

While the resulting liquid phase is returned to the wastewater treatment plant, the solid fraction is a hygienised product with low moisture content that can be used as a low bulk biofuel at existing power plants. The resulting product streams also allow extraction of nutrients;

primarily phosphorus and nitrogen, as well as removal of heavy metals for special disposal.

Presently, extraction processes of such are under development.

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This master thesis of science aims to investigate the;

- principal conditions for extraction of phosphorus from HTC converted sewage sludge by acid leaching and filtration;

- recovery and separation of dissolved phosphorus from leachate via precipitation and filtration.

To fulfil the outlined objectives, the project was divided into two phases. Initially, a review of present technologies for phosphorus recovery from varying products and streams originating from wastewater treatment plants was made. Industrially relevant leaching parameters and ranges was also determined during this phase, as well as the interesting conditions for phosphorous precipitation. During the second phase the conditions were experimentally investigated and evaluated through several analyses.

1.2.2 DELIMITATIONS

In this project only one type of sludge is studied. Variations in sludge composition as a result of different purification methods and precipitation agents are not considered. Neither are different process conditions of the hydrothermal carbonisation step. Therefore, the HTC material used for leaching was produced from one single batch of dewatered sludge converted at the same specific conditions.

The downstream process is not to be studied in details. Due to limitations in the laboratory equipment, filtration is not possible to perform at industrially relevant conditions. Also, influences of displacement wash have been left out of this project and, as a result optimal wash ratios have not been investigated.

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2.1 P

All living organisms require phosphorus for growth and survival as it is a vital part of DNA, RNA and the energy carrier ATP (Cornel & Schaum, 2009). Moreover, bone is made up of phosphate mineral and the buffering system in blood involves phosphate ions. By being a part of cells, phosphorus becomes an irreplaceable element. (Rayner-Canham & Overton, 2010) As a result, the agricultural industry is the largest consumer of phosphorus by the use of fertilising products. Since the arable lands naturally lacks the element, phosphorus containing fertilisers are used to improve the harvests. Due to a growing world population and, thus, food demand, the demand for fertilisers is increasing every year. It is estimated that 4 million tonnes more phosphorus will be required annually corresponding to an expansion of 2 % of the market.

(Wiechmann, et al., 2013)

A few years ago, phosphorus was also an important ingredient in detergents, but due to ecological reasons is has been replaced by zeolites. However, toothpastes and baking powder still contain phosphorus compounds, and phosphoric acid is an important constituent in rust remover for industrial and domestic use. Phosphoric acid is also added to soft drinks to prevent bacterial growth and to canned food where phosphate ions react with leached metal ions to form an inert, harmless compound. Other applications of phosphorus compounds are as a selective solvent to separate uranium from plutonium compounds, and fire retardant. (Rayner-Canham

& Overton, 2010)

Figure 1 describes the interconnected cycles of phosphorus including the human activity and impact on the system. The slowest cycle time is millions of years and belongs to the inorganic cycle which is initiated by erosion of phosphorus containing minerals. The dissolved matter proceeds to the oceans where it ends up in the bottom sediments. By a tectonic uplift the material is brought to the atmosphere where it becomes available to plants. On land, phosphorus enters a second cycle as soon as the plants take up the element from soil. The cycle continues with the consumption of the plants by animals or humans and is enclosed when the organic waste is returned to the soil. The cycle time of the biological route ranges from some weeks up to a year. Without outside influence, these cycles are in balance and enclosed. Figure 1, however, describes how human activities remove phosphorus from the short-term, biological cycle, and introduce the material to the geological cycle of millions of years when disposing it in the oceans. Imbalance and lack of nutrients in the biological cycle arise from careless

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handling of phosphorus. The cycles also imply that the balance is restored after some millions of years if no action is taken. (Cornel & Schaum, 2009)

Figure 1. The geological (long-term inorganic) and biological (short-term organic) cycles of phosphorus on earth including the human impact (Cornel & Schaum, 2009).

Phosphorus is often the limiting factor for growth in coastal waters and freshwater ecosystems.

Thus, growth of microalgae and cyanobacteria is a sign of eutrophication resulting from elevated concentrations of phosphorus, and to some extent even nitrogen. In aquatic systems phosphorus is only found in pentavalent forms, such as orthophosphate, polyphosphate and organic phosphate esters. The element is transported to water bodies as mixtures of the mentioned molecules. In bottom sediments the organic particulates are deposited. While microorganisms make use of the organic matters, phosphates are released. Orthophosphate is also produced in enzymatic and chemical hydrolysis when the compounds are suspended in water. Phosphorus in the form of phosphate is the only compound algae, bacteria and plants can assimilate which result in an excessive productivity of the organisms. This in turn leads to anoxic waters due to high bacterial populations and respiratory rates. The biodiversity is jeopardised as fishes die at low concentrations of dissolved oxygen. (Correll, 1998) Figure 2 describes the phosphorous cycle in water bodies.

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Figure 2. Phosphorous cycle diagram in water bodies (Correll, 1998).

Phosphorus is most commonly produced from phosphorus-containing minerals such as apatite of various kind (Elding, n.d.). Apart from phosphorus, apatite usually contains cadmium, and uranium. Thus, mining of apatite rock results in exposure of heavy metals and radionuclides to the biosphere jeopardising the environment and health of many living species (Stark, n.d.) (Wiechmann, et al., 2013). The explored sources are estimated to be depleted in a few decades and the quality is gradually decreasing. Five countries only, control over 90 % of the explored phosphate reserves; Morocco, China, Algeria, Syria and Jordan, and some of them are politically unstable countries adding insecurities to the market. Consequently, recovery and recycling of phosphorus is becoming an important path to be able to meet the demand of the future. (Wiechmann, et al., 2013)

2.2 W

A wastewater treatment plant (WWTP) collects water from households and, in many cases, industries. The primary objective is to remove contaminants and reduce the concentrations of nutrients to obtain water clean enough to be discharged into the recipient without jeopardising the aquatic ecosystem. The phosphorus ending up in the sewage originates from human metabolic wastes, detergents and cleaning products as earlier described. Approximately 15 % of the phosphorus is organically bound, 50 % is present as inorganic orthophosphate and 35 % as complex inorganic phosphates, referred to as polyphosphates. In the sewage, however, polyphosphates are gradually hydrolysed into orthophosphate. In municipal sewage the ratio between BOD7 and phosphorus is 100 to 3, which is too high phosphorous concentration to remove all from the aqueous phase by biological treatment only where 1-2 g of phosphorus is

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assimilated for every 100 g of BOD7 consumed. Thus, additional process steps are required to achieve acceptable levels of nutrient concentrations. Most commonly the sewage water undergoes three different treatment methods; mechanical, chemical and biological cleaning.

2.2.1 MECHANICAL CLEANING

In the first step of the mechanical cleaning, objects with potential of damaging the following process equipment are mechanically removed from the stream by passing the incoming water through bar screens. In an often aerated sand chamber, grit and sand settle on the bottom and is removed while organic matter is kept suspended and transferred onwards. Before entering the last step of the mechanical treatment, air is bubbled through the water to accumulate fat and grease on the surface, which is either removed before or in the primary sedimentation tanks. In this treatment step, sedimentation of matter heavier than water occurs in large basins or tanks, which produces the so called primary sludge. Approximately 30 % of organic matter and 70 % of suspended particles are collected in the primary sludge.

2.2.2 CHEMICAL CLEANING

In many WWTPs the sedimentation in the pre-settling tanks is accelerated by addition or precipitation chemicals before entering the primary sedimentation units. Common precipitation agents are lime, and salts of iron; in ferrous (Fe²⁺) and ferric (Fe³⁺) forms, and aluminium. The idea is to reduce the phosphorous content in dissolved and colloidal form by up to 25 % by formation of chemical complexes. Precipitation of dissolved phosphorus by the use of ferric salts is described by Reaction 1. The iron(III) ion also reacts with water to form hydroxides according to Reaction 2.

𝐹𝑒³⁺+ 𝐻𝑃𝑂₄²⁻⇄ 𝐹𝑒𝑃𝑂₄(𝑠) + 𝐻⁺ (1)

𝐹𝑒³⁺+ 3 𝐻₂𝑂 ⇄ 𝐹𝑒(𝑂𝐻)₃(𝑠) + 3 𝐻⁺ (2)

The metal hydroxide is a gelatinous substance which captures precipitates, particulates and impurities while subsiding. Thus, the phosphate precipitates stick to the flocks which greatly improves the phosphorous reduction. Consequently, the primary sludge is phosphorous enriched. Flocculation also accelerate the sedimentation velocity and improves the reduction of organic particulates and other toxins that might harm the subsequent process step.

When aluminium salts are used as precipitation agent the reactions are comparable to Reaction 1 and 2. If a ferrous salt is used, dissolved phosphorus precipitates according to Reaction 3.

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3 𝐹𝑒²⁺+ 2 𝐻𝑃𝑂₄²⁻⇆ 𝐹𝑒₃(𝑃𝑂₄)₂(𝑠) + 2 𝐻⁺ (3)

However, as hydroxide precipitates do not form when iron(II) is used no significant phosphorous reduction is obtained unless the pH value in the sewage is above 8.5. Thus, oxidation of iron(II) into iron(III) is required to form hydroxide complexes. In a WWTP this is accomplished in the aerated sand chamber or in the aerated zone before the pre-settling tanks.

The reaction is described by Reaction 4. Once iron(III) has formed Reaction 1 and 2 are assumed to be valid for the proceeding reactions.

4 𝐹𝑒²⁺+ 𝑂₂+ 2 𝐻₂𝑂 ⇆ 4 𝐹𝑒³⁺+ 4 𝑂𝐻⁻ (4)

Chemical precipitation reduces the total amount of dissolved phosphorus in the sewage water by 80-95 %.

2.2.3 BIOLOGICAL CLEANING

The water leaving the pre-settling tanks is introduced to the biological purification step where the remaining organic carbon is removed, and the concentrations of nitrogen and phosphorous reduced. The biological basins are divided into aerobic and anoxic zones to maximise the reduction of undesired materials. Microorganisms, mainly bacteria, are the workhorses in the biological treatment. Bacteria of different kinds oxidise organic matter to obtain energy which is used for growth of new cells. Oxygen, in aerobe zones, and nitrate, in anoxic zones, are the primary oxidising agents in the biological treatment. Nitrate is formed when ammonia is oxidised while oxygen is provided through spargers in aerated zones in the basins.

The biological phosphorous reduction is referred to as bio-P and the principles of the process are described in Figure 3. During aerobic condition in aerated zones in the basins, specific bacteria, bio-P bacteria, store excess amounts of phosphorus as polyphosphates in the cell.

When an anoxic zone is reached, bacteria assimilate volatile fatty acids, VFAs, in organic polymers. The ability to oxides organic material to obtain enough energy needed for the storage process is limited when oxygen is lacking. Instead, the bacteria gain energy from hydrolysing the intracellular polyphosphates to phosphate. The dissolved phosphorus is then transferred out of the cell resulting in an increase of the phosphate concentration. When the bacteria again enter an aerated zone, the assimilated organic material is used for cell growth and energy for assimilation of phosphates. As previously described, phosphorous, but also nitrogen and a range of other elements, are essential building blocks in any living organism. Thus, the amount of nutrients in the sewage is readily decreased in the biological treatment step by the growth of biomass.

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Figure 3. Biological phosphorous removal under anaerobic and aerobic conditions. A: Release of phosphate to produce energy for assimilation of organic material under anaerobic conditions, and phosphate uptake and respiration during aerobic conditions. B: Concentration changes of VFA and phosphate in anoxic and aerobe zones (Balmér, et al., 2007, modified).

For every 100 g of BOD7 removed from the wastewater, 1-2 g of phosphorus is removed. The biological phosphorous reduction is 15 to 30 %. A secondary settling tank follows the aerated tanks where the bio-sludge accumulates on the bottom. The majority of the sludge is recirculated to the inlet of the biological treatment step to provide enough activated biomass.

Excess secondary sludge is removed from the process at the same rate as new biomass is generated.

If the phosphorous concentration in the water leaving the secondary settling tank is still high, or if precipitation agents have not been used in earlier process steps, chemical cleaning takes place at this point. Lastly, the water often passes through a sand filter to remove any remaining particulates, for example when precipitation of phosphorus occurs after the bioreactor.

2.3 S

In Sweden 1 million tonnes of dewatered sludge is annually produced, corresponding to 200 000 tonnes of dry solids (DS). Primary and bio sludge combined without any treatment is commonly referred to as mixed or raw sludge. The amount of primary sludge produced per person and year, 18 kg of suspended solids (SS), is larger than the bio sludge amount, 9-13 kg SS. The dry solids (DS) content of raw sludge is usually approximately 3-4 %, where the organic matter constitutes < 70 % of DS depending on precipitation agent and operating mode.

Macronutrients constitute a small part of the DS content of raw sludge; nitrogen 3.6 %, phosphorus 2.8 %, while calcium, magnesium, potassium and sulphur together constitutes

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approximately 4 % of DS. (Tideström, et al., 2007) In the sludge, phosphorus exists in both inorganic form as iron, calcium and aluminium salts, and organic form.

Sludge is a complex material where interactions between water and sludge particles define the characteristics and hinder water removal. In the voids between particles the most easily removed water is trapped and is simply separated by gravitational forces, giving so called thickened sludge. For water captured in capillaries between different particles, mechanical forces are required to increase the dry solids content (DS). This is achieved by centrifugal forces, vacuum or pressure resulting in dewatered sludge with a dry solids content of 20-35 % depending on sludge type and process. The remaining water is either adsorbed or bound in cells and can only be separated by evaporation. (Tideström, et al., 2007)

To minimise the risk of fermentation and unpleasant odour the sludge is stabilised biologically, thermally or chemically. The most economical chemical stabilising method is lime addition to sludge, most preferably to dewatered sludge. Heat is produced when quicklime (CaO) reacts with water and simultaneously increases the pH, aiming for a value above 11. At these conditions pathogens are killed. (UKWIR, 2015) (Tideström, et al., 2007) The disadvantages of lime stabilisation are the increased amount of sludge and a merely temporary stabilising effect as microbial processes starts again. Biological stabilisation methods are completely dominated by anaerobic digestion from which biogas, a valuable by-product, is obtained.

Primary sludge contains a larger fraction of easily accessible organic carbon than excess sludge which gives a high biogas production potential. Roughly 50 % of the organic matter is converted to biogas, which gives a significant reduction in sludge amount. The inorganic matter is intact after the digestion resulting in a larger inorganic fraction in the sludge after the digestion than before. (Tideström, et al., 2007) The energy content of mixed sludge is approximately 135 kWh which is equivalent to 30 kg DS per person and year. The annual energy recovery per person is about 75 kWh in the form of biogas from anaerobic digestion and 20 kg DS per person and year remains after the treatment. Digested sludge also contains approximately 0.65 kg phosphorus and almost 1 kg of nitrogen per person and year. (Svenskt Vatten AB, 2013) (Tideström, et al., 2007)

Before entering the digester, the sludge is thickened to a DS content of 5-8 % to reduce the required reactor volume and energy needed for heating the material. To reduce the volume of the digested sludge, dewatering is often applied giving a DS content of 20-35 %.

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2.4 S

Sludge has long been considered a disposal problem mostly because of the uncertainties and variation of the product composition. The large fraction of heavy metals and increasing concerns regarding pathogens and pharmaceutical residues are all reasons for the ongoing change in use of the sludge in many European countries. The amount of toxic materials accumulated in the sludge in WWTP is a result of the use of chemicals and consumption of foods originating from production of lower standards and qualities. (Naturvårdsverket, 2013a) In Europe the use of sludge in agriculture and landfill is being more and more restricted. Sweden is the only Scandinavian country allowing partly stabilised and hygienised sludge for agricultural use but spreading on farmlands is becoming rare (PURE, 2012). Statistics of the use of sludge from WWTPs in Stockholm demonstrate how 80-90 % was returned to arable lands during the 80s while only 20 % in 2013 despite lower concentrations of heavy metals and contaminants in the sludge. (Lücke-Johansson, 2014) (Tideström, et al., 2007) From the beginning of the 21^st century, the dominating sludge uses have been land applications including coverage of mining sites in the north of Sweden, construction of golf courts and sound barriers (Lücke-Johansson, 2014). Further treatment of sewage sludge is however necessary when used in agriculture and land applications. In Sweden companies working with recycling and waste management handle and hygienise sludge. Methods used are long-term storage (around six months) in aerated environment, or shorter storage time at a temperature of 55 ^oC. Composting of dewatered sludge together with dry bark and wood chips is also a possibility. Other ways for hygienisation are pasteurisation; heating the sludge to 70 ^oC for 30-60 min, before anaerobic digestion. (Olofsson, et al., 2007) Digestion, lime stabilisation, composting and pasteurisation are all common techniques of handling sludge in the countries around the Baltic sea. (PURE, 2012)

In 2005, Germany decided to only accept waste holding a maximum of 5 % organic matter in landfills (Stark, et al., 2005b) (PURE, 2012). While landfilling stopped completely, and composting and landscaping applications decreased, incineration of sludge started to grow rapidly. From 2007 and onwards mono- and co-incineration have been the dominating methods for handling sludge, and approximately 55 % of the material was treated in 2011. Mono- incineration are facilities where sludge is burned exclusively. This also includes gasification plants for production of syngas for heat and electricity recovery. In addition, combustion of sludge is frequently performed in coal fired plants, waste incineration plants and cement plants, referred to as co-incineration. In cement plants co-incineration is profitable as additives and

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fuel requirements decrease. According to statistics, the regulations have not affected the agricultural applications and the fraction of sludge used in farming has remained unchanged at 30 % (Wiechmann, et al., 2013). In Switzerland, however, spreading of stabilised sludge on arable lands as well as disposal have been prohibited since 2006. Instead, thermal disposal of sludge through incineration is used, either together with or without household wastes, or at cement plants. (Swiss Confederation: Federal Office for the Environment, 2015) Incineration is also how the Netherlands handle the sludge as agricultural use is banned due to phosphorous saturation of arable lands (Stark, et al., 2005b).

In 2012 the Swedish Environmental Protection Agency (Naturvårdsverket) was requested by the government to investigate the possibility of sustainable recycling of phosphorous to arable land. In the study, estimations of the phosphorous content of a vast number of resources and streams were made as well as an assessment of the potential as a source of phosphorus accounting for accessibility. (Naturvårdsverket, 2013a) (Naturvårdsverket, 2013b) The investigation showed that several waste streams in today’s society are enriched in phosphorus and bare high potential for recirculation. Approximately 5800 tonnes of phosphorus are accumulated in the sewage sludge from municipal WWTP of which only 25 % is returned to arable lands. However, to achieve a sustainable recycling of nutrients, a number of conflicting national environmental quality objectives must be taken into consideration. For this matter the objective regarding “a non-toxic environment” is of highest significance. The objective seeks to limit the dispersion of toxic pollutants and non-naturally occurring substances in the environment which in different ways have a negative impact on plants, animals and humans.

Despite the increasing consumptions of chemicals and products, the knowledge of the long and short term effects of the substances is lacking. Thus, restricting the accessibility and dispersion will possibly prevent predicted and unpredicted health concerns caused by toxins circulating in society. (Naturvårdsverket, 2016) (Naturvårdsverket, 2013a)

2.5 T

As a result of national objectives in many European countries as described above, several technologies for phosphorus recovery from WWTPs and products thereof have been developed in the last few years. Some methods aim to recover phosphorus from liquid streams, while many technologies use anaerobic sludge or ash from sludge incineration as starting point for further treatment. A few highly relevant and interesting technologies are described below.

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2.5.1 RECOVERY OF DISSOLVED PHOSPHORUS IN LIQUID

A few technologies developed aim to precipitate phosphate from substreams of the water purification process. Streams of interest are sludge from the biological treatment, both excess sludge and the recirculated fraction, as well as the reject water from dewatering of anaerobe sludge. For these streams, processes for precipitation of struvite (magnesium ammonium phosphate) for phosphorous recovery have shown potential. Other methods are crystallisation by addition of magnesium chloride, and sodium or magnesium hydroxide where again struvite crystals are formed. For small-sized and individual sewage disposal systems, focus has been upon adsorption of phosphates, where the adsorption modules are replaceable. Another approach to recover dissolved phosphorus has been by ion exchange and electrodialysis for production of phosphoric acid. (Tyréns AB, 2013)

2.5.2 PHOSPHOROUS RELEASE AND RECOVERY FROM SEWAGE SLUDGE

Sewage sludge is regarded as one of the most promising material streams that can be used for phosphorous recovery (Naturvårdsverket, 2013a). Many different strategies have been developed to recover the macronutrient from digested sludge. During digestion, calcium phosphate can be precipitated by addition of calcium silicate hydrate working as an adsorbent in the FIX-Phos process. In the AirPrex™ process struvite is precipitated from digested sludge in a separate tank.

However, the vast majority of developed processes pretreat the anaerobe sludge to make the fractionation of solid and liquid phase possible, before and/or after phosphorus release.

Different pretreatment methods have been investigated over the years. In Sweden supercritical water oxidation (SCWO) have been tested along with thermal hydrolysis. In Germany where incineration of sludge is very common, phosphorous release from incinerated ashes has gained much research. (Stark, et al., 2005b)

To release phosphorus from treated or untreated sludge separate leaching strategies are used in the different methods. Some processes are briefly presented below.

2.5.2.1 GIFHORN PROCESS, MODIFIED SEABORNE

The Seaborne process can be adapted to several organic materials including sewage sludge, manure and agricultural waste. Unprocessed biomass is treated with sulphuric acid and hydrogen peroxide to dissolve phosphorus and metals. The remaining organic material is separated from the leachate by centrifuges and sodium sulphide is added to the liquid stream to precipitate the heavy metals. After separation dissolved phosphorus is recovered by addition of

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magnesium- and sodium hydroxide to precipitate struvite. (Nieminen, 2010) (Tyréns AB, 2013) (P-REX, 2015)

2.5.2.2 BIOCON® PROCESS

Phosphorus and metals are leached from sewage sludge ashes by sulphuric acid in the BioCon®

process. The phosphate is recovered as phosphoric acid by passing the leachate over a series of ion exchangers. The ion exchangers are then regenerated by hydrochloric acid producing ferric chloride. (Levlin, et al., 2002) (Nieminen, 2010)

2.5.2.3 SEPHOS AND ADVANCED SEPHOS

In the SEPHOS process sewage sludge ashes are treated with sulphuric acid at a pH value below 1.5 to dissolve phosphorus and metals. The remaining solid phase is separated from the leachate, which in turn is treated with sodium hydroxide. Below pH 3.5 aluminium phosphates precipitate while heavy metals remain dissolved. Further treatment in the advanced SEPHOS recovers calcium phosphates. Analysis of the product presented a phosphorus content of 12 %, compared to 9.8 % in the ashes before treatment, and a significantly reduced heavy metal content. (Nieminen, 2010)

2.5.2.4 PASH PROCESS

Leaching of sewage sludge ashes by hydrochloric acid is performed in the PASH process.

Alamine 336 and tri-butyl-phosphate is added to the filtrate after separation to remove heavy metals. In a final step struvite or calcium phosphate is recovered to form a product of 16 % phosphorus. (Nieminen, 2010) (Tyréns AB, 2013)

2.5.2.5 AQUA RECI PROCESS

The Aqua Reci process is applied to sludge with phosphates strongly chemically bound to iron or aluminium ions. In this process supercritical water oxidation (SCWO) is used to disintegrate organic compounds and toxins at a temperature and pressure above 374 ^oC and 221 bar, respectively. From the remaining inorganic ashes, phosphates and coagulants are recovered.

(Stark, et al., 2005b) Experiments have indicated that leaching of phosphorous from residues from the SCWO process is easier than of ashes from incineration (Stark, 2005).

2.5.2.6 KREPRO

By thermal hydrolysis and addition of sulphuric acid to lower the pH phosphorous recovery is possible in the KREPRO process (Stark, 2002). Ferric phosphate, FePO4, was obtained in a pilot plant and the study claimed that the phosphate compound had considerable fertilising

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effects (Stark, et al., 2005b). These results have not been possible to verify in later experiments.

The chemical demand in this process is approximately 0.5 kg/kg DS (Stark, 2002).

2.5.2.7 AVA CLEANPHOS

AVA cleanphos is based on acid leaching of HTC converted sludge and is therefore the technology most related to this project. The process is developed by AVA-CO2 Schweiz AG and is made up of three steps; 1) acid leaching of grinded HTC product, 2) nanofiltration separating phosphoric acid from metal sulphates, and 3) concentration of phosphoric acid from 5 up to 75 %. The company claims that 80 % of the phosphorus is dissolved in the leaching and that the heavy metal fraction in the product is only 8-10 %.

2.6 H

Already in 1913 hydrothermal carbonisation (HTC) was demonstrated by Friedrich Bergius who simulated natural coalification and received the Nobel Prize in 1931 for the discovery.

HTC, sometimes referred to as wet pyrolysis or wet torrefaction, is a process for conversion of organic feedstock with high moisture content into a solid product denoted hydrochar or biochar.

(Libra, et al., 2011) (He, et al., 2013)

The HTC process is performed at elevated temperatures, in the range from 160 to 250 ^oC, and autogenous pressure. At the lower temperatures in the given range and at corresponding pressures the majority of the organics retains in solid state, resulting in only small volumes of gaseous materials. (Libra, et al., 2011) As the temperature and reaction time increases the amount of carbon remaining as biochar as well as the energy yield are reduced. HTC conversion at higher temperature and reaction time have been demonstrated to enhance the energy content of the biochar from 17 to 19 MJ/kg (Danso-Boateng, et al., 2015). The fuel ratio (fixed carbon to volatile matter) of the carbonaceous material is also improved, e.g. the fuel ratio of raw sludge has been reported to increase from 0.02 to 0.18 in the conversion process (He, et al., 2013).

Consequently, by the HTC conversion a more attractive fuel that is suitable for power plants is produced from a low-value material.

Despite large amounts of water, the HTC-process generates comparatively high yields without need for energy-intensive prior drying (Libra, et al., 2011). Compared to conventional drying methods of sludge, the HTC process has been demonstrated to save 60 % of thermal energy and 65 % of electric energy on laboratory scale. Moreover, the high carbon efficiency of the process minimises emissions of greenhouse gases. (vom Eyser, et al., 2015)

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The advantages of the HTC process regarding efficiency and yields still obtained in the presence of high moisture content open up for a wide range of potential carbon sources. Of particular interest are assorted waste streams, such as wet animal manure, aquaculture and algal residues, municipal solid waste and sewage sludge (Libra, et al., 2011).

By hydrothermal carbonisation the dewaterability of sewage sludge is readily improved at reaction times longer than 1 h when compared to untreated dewatered sludge. This is mainly due to a more hydrophobic character of the biochar resulting from the loss of oxygen containing functional groups during the process (He, et al., 2013). Enhanced dewaterability reduces drying cost (vom Eyser, et al., 2015) and is also one key step to enable an efficient fractionation of the HTC converted sludge.

Concerns regarding the pathogens in anaerobic sludge are tackled by HTC treatment. As a result of the relatively high temperatures (above normal temperatures used in autoclaves) and reaction times the sludge is hygienised since both bacteria and viruses are killed. Furthermore, studies of decomposition of some of the most common pharmaceuticals in the HTC process have reported reductions of above 95 % at a residence time of 4 h. (vom Eyser, et al., 2015)

2.7 L

Leaching, also referred to as solid-liquid extraction, is a common unit operation for separation of soluble components from insoluble ones by addition of a suitable solvent. There are two different ways in which leaching is accomplished; i) phase transition, A(s) → A(aq), and ii) chemical reaction followed by dissolution of the component, A(s) + X → AX(aq). (Theliander, 1996)

2.7.1 MASS TRANSFER IN LEACHING AND RATE DETERMINING STEP

Due to limited knowledge of the processes taking place and the many different phenomena encountered in a single leaching operation, one particular theory is practically impossible to apply. However, the mass transfer in a leaching process where a solvent is used to dissolve material from inside a particle is generally described by five steps (Theliander, 1996) (Geankoplis, 2013);

1. Transport of solvent molecule from liquid bulk to the surface of the solid particle.

2. The solvent molecule penetrates and/or diffuses into the solid and the dissolution (reaction) zone.

3. Dissolution (and reaction) of solute into solvent.

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4. Diffusion of solute and solvent from the dissolution (reaction) zone to the surface of the particle.

5. Transportation of solute and solvent from the reaction surface to the liquid bulk.

One or two steps from the above sequence is the rate controlling step of the leaching process.

Typically, the mass transfer occurring in step 1 and 5 is fast and the resistance can be neglected, leaving three possible controlling resistances. (Theliander, 1996) (Geankoplis, 2013)

The primary driving force in a leaching process is the concentration gradient of a component A, in this case a solute, between solvent and solid particle. The net flux of the solute is therefore from high to low concentration. In one direction, z, the mass transport in a fluid or solid for constant total concentration in the fluid is described by Fick’s law:

𝐽_𝐴𝑧^∗ = −𝐷_𝐴𝐵𝑑𝑐_𝐴

𝑑𝑧 (5)

where 𝐽_𝐴𝑧^∗ is the diffusion flux of A relative to a moving fluid in mole of A/(s ∙ m²), −𝐷_𝐴𝐵 is the diffusion coefficient of A in B in m²/s, and 𝑐_𝐴 is the concentration of A in mol/m³.

For diffusion in solids which is not dependant on the structure of the solid Fick’s law applies.

If 𝑁_𝐴 is the total flux of A relative to a stationary point, and any convective flux of A can be neglected, Equation 6 is valid.

𝑁_𝐴 = 𝐽_𝐴𝑧^∗ = −𝐷_𝐴𝐵𝑑𝑐_𝐴

𝑑𝑧 (6)

In leaching, this would be valid when the solids contain large amounts of water and the solute is diffusing through this relatively homogenous solution. If the solid instead is porous with interconnected voids, the tortuosity, τ, and the open void fraction, ε, have to be considered and form an effective diffusivity according to Equation 7. (Geankoplis, 2013)

𝐷_{𝐴 𝑒𝑓𝑓} = 𝜀

𝜏𝐷_𝐴𝐵 (7)

2.7.2 PARAMETERS AFFECTING A LEACHING OPERATION

Agitation prevents sedimentation and stagnant zones, which reduces the contact surface between solvent and solute resulting in an inefficient process requiring longer residence time for reaction/dissolution. Also the diffusion and mass transport is enhanced by stirring. (Ström, 1995)

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The particle size of the solid phase has a major impact on the efficiency. A narrow particle size distribution is desired in any leaching process as the residence time can be well optimised. To obtain a suitable, uniform particle size the solid material is often grinded. A small particle size gives a large contact area between solid particle and solvent. This shortens the distance the solvent and solution have to travel in the pores to and from the reaction zone. However, a small particle size prolongs the separation of solid matter from solution as the material is tightly packed. (Ström, 1995)

The solvent used in the leaching operation should be selective to obtain a pure overflow of desired components. A low viscosity of the solvent is preferable as this gives the most efficient agitation and the viscosity increases by the dissolution of solute. To enhance the driving force of the leaching process – the concentration gradient – a pure solvent without any dissolved matter is initially preferred. As the leaching continues the concentration gradient decreases and so does the rate of dissolution. (Ström, 1995) The pH value of a solvent is also significant as the solubility of a salt can be shifted to the right or left by a change in pH, i.e. hydronium ion concentration. In general, salts of weak acids are more soluble in acidic solutions than in pure water. This is due to the fact that the anion from the salt is the conjugate to a weak acid thus reacting with a hydronium ion from the strong acid in the solvent. When the anion is removed, more of the salt have to be dissolved to reach equilibrium according to Le Châtelier’s principle.

The temperature during a leaching operation is also to be considered. Most commonly, the solubility of a salt is enhanced by an increase in temperature. This is, however, not always the case. For a change in temperature, a new equilibrium constant can be estimated by van’t Hoff equation, equation 8

𝑑(𝑙𝑛𝐾) 𝑑𝑇 =∆𝐻^𝑂

𝑅𝑇² (8)

where K is the equilibrium constant at any temperature T, R is the ideal gas constant, and ∆H^o is the enthalpy of dissolution. Clearly, if the dissolution of a salt is exothermic (∆H^o < 0), a high temperature of the solution will have a negative effect on the release. But if the dissolution of the salt requires heat (∆H^o > 0), the process will benefit from a high temperature of the surrounding. Though, in leaching the diffusivity constant is also relevant and increases with higher temperature, which in turn will enhance the rate of dissolution of a compound (Ström, 1995). Consequently, a temperature increase might boost the mass transport but have the opposite effect on the dissolution process of a particular salt.

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2.7.3 LEACHING OF PHOSPHORUS FROM HTC CONVERTED SLUDGE

Leaching of phosphorus from HTC converted sewage sludge have not been studied in any large extent except for AVA cleanphos (AVA-CO2 Schweiz AG, 2015), thus data on the subject is limited. Extensive studies of phosphorus release by leaching have, however, been performed on sludge incineration ash and sludge treated by supercritical water oxidation (SCWO). The primary difference between the material from the HTC process and the residues from SCWO and ashes from incineration is the organic fraction which is non-existing in the two latter cases (Xu & Fang, 2014) but approximately 50-70 % in the HTC material (Libra, et al., 2011). A large fraction of metals is present as oxides in both ashes and SCWO residues and are formed during the treatment (Stark, 2005). In HTC converted sludge, metals exist in other kinds of complexes and as counterions to organic compound as well. The influence of the organic, and inorganic matter and form, in the leaching process is difficult to predict as the exact composition is unknown and will vary greatly with the original sludge.

Independent of the pretreatment method the dissolution of metal salts is controlled by the solubility which varies with temperature and pH. The solubility product constant, Ksp, is the equilibrium constant for the solubility equilibrium of a slightly soluble ionic compound.

According to reaction kinetics, it can also be expressed as the quotient between the rate of dissolution of the salt and the rate of precipitation. Thus, by altering the amounts of the components, the rate at which equilibrium is reached is changed.

As aforementioned, the dominating metal phosphates in sewage sludge depend on the agent used in the chemical precipitation step. Regardless, iron, aluminium and calcium are the most abundant phosphate salts in the sludge. The solubilities for some of the salts assumed present in the sludge are depicted against pH in Figure 4. Also, Figure 5 depicts the dominating orthophosphate species at varying pH values and the tendency of a strong acid to “donate”

hydrogen atoms to a weak acid, is demonstrated for the phosphoric acid system. Clearly, for efficient dissolution of FePO4 and AlPO4 rather extreme pH values are required for high dissolution.

At pH values above 7 the concentration of hydroxide ions in the liquid is sufficient to form metal hydroxides with Al³⁺ and Fe³⁺. As the cations are removed equilibrium is shifted and dissolution of these metal phosphates continues. It is also evident that some calcium complexes are not soluble at alkaline conditions. The solubility product constant of calcium hydroxide, Ca(OH)2, is 5.02 · 10^-6 (Dean, 1999) and lower than any calcium phosphate (compare Table 1).

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Thus, equilibrium is not shifted towards dissolution of calcium salts as no Ca²⁺ ions are removed by precipitation of calcium hydroxides. This phenomenon has also been demonstrated on SCWO treated sludge from Karlskoga and Stockholm containing 3 and 8 % calcium respectively. When leached with sodium hydroxide, 90 % of the phosphorus was dissolved from the residues originating from Karlskoga, while only 65 % was released from the Stockholm sludge. (Stendahl & Jäfverström, 2003a) If large fractions of phosphates are bound to calcium in the sludge, alkaline leaching is disadvantageous compared to acid leaching.

Figure 4. Solubilities of metal phosphates at varying pH (Stumm & Morgan, 1996).

Figure 5. pH diagram over the phosphoric acid system.

Total concentration 0.1 mol/L (Gunneriusson, 2012).

Studies conducted on acid and alkaline leaching of SCWO residues and incinerated sludge have reported complete phosphorous release at low acid concentrations (≈ 0.1 M HCl) while high release of phosphorous has been difficult to achieve even at concentrations around 5 M NaOH (Stark, 2002) (Stark, et al., 2006) (Biswas, et al., 2009). The chemical demand for acid leaching

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is thus less than for alkaline leaching. On the other hand, under acidic conditions metals have been shown to leach out to greater extent than under alkaline conditions. More advanced separation technologies will be required to extract a pure product suitable for the fertilising industry after acid leaching. (Stark, 2002) (Petzet, et al., 2012)

By assuming the phosphate precipitates present in the HTC converted sludge are the same as those formed during chemical precipitation in WWTP the theoretical acid demand for dissolution of the metal phosphates can be estimated as follows (Petzet, et al., 2012):

𝐴𝑙𝑃𝑂₄(𝑠) + 3 𝐻⁺ ⇄ 𝐴𝑙³⁺+ 𝐻₃𝑃𝑂₄(𝑎𝑞) (9)

𝐹𝑒𝑃𝑂₄(𝑠) + 3 𝐻⁺ ⇄ 𝐹𝑒³⁺+ 𝐻₃𝑃𝑂₄(𝑎𝑞) (10)

𝐹𝑒₃(𝑃𝑂₄)₂(𝑠) + 6 𝐻⁺ ⇄ 3 𝐹𝑒²⁺+ 2 𝐻₃𝑃𝑂₄(𝑎𝑞) (11) 𝐶𝑎₉𝐴𝑙(𝑃𝑂₄)₇(𝑠) + 21 𝐻⁺ ⇄ 9 𝐶𝑎²⁺+ 𝐴𝑙³⁺+ 7 𝐻₃𝑃𝑂₄(𝑎𝑞) (12) According to Reactions 9 through 12 three moles of hydrogen ions are required to dissolve one mole of phosphorus. Most likely, there are more acid consuming compounds of both organic and inorganic origin present in the HTC converted sludge (He, et al., 2013). Thus more hydrogen is needed for dissolution of each phosphate ion than theoretically (Petzet, et al., 2012).

Studies conducted on sewage sludge ashes (SSA), have used approximately 0.4 to 0.7 kg HCl/kg SSA. In leaching experiments where sulphuric acid has been used, charges has commonly ranged from 0.3 to 0.5 kg H2SO4/kg SSA. The fraction of dissolved phosphorus has ranged from 85 to 98 % and 84 to 99 % for hydrochloric acid and sulphuric acid respectively.

(Petzet, et al., 2012) Due to the minor differences reported on the leaching experiments with hydrochloric and sulphuric acid, both acids are suitable.

For a commercial process, however, the overall economy is of importance, cost and consumption of chemicals as well as wear on process equipment have to be accounted for.

Hydrochloric acid is more expensive than sulphuric acid and requires corrosion resistant materials that increase the costs even more. Sulphuric acid, on the other hand, is one of the cheapest and most frequently used acids on the market, and less caution must be exercised when used compared to hydrochloric acid.

Another important concept to consider in leaching of HTC converted sludge is the ionic strength of the solution. The ionic strength is a measure of the concentration of ions and their respective charge as described by equation 13 (Stumm & Morgan, 1996)

21 𝐼 =1

2∑[𝑖]

𝑖

∙ 𝑍_𝑖² (13)

where I is the ionic strength, [i] is the concentration of any ion i in the solution and Zi is the charge number of that particular ion. Clearly, the higher the ionic charge the higher the resulting ionic strength. The importance of ionic strength is revealed when the effective concentration, i.e. activity, of a solute is calculated (Ebbing & Gammon, 2009). For diluted systems, up to 0.005 mol/L, the activity is determined by the Debye-Hückel theory which uses electrostatic attraction and repulsion forces as a basis according to equation 14 (Stumm & Morgan, 1996)

lg 𝛾_𝑖= −𝐴𝑍_𝑖^2√𝐼 (14)

where γi is the activity coefficient and A is a solvent depending constant. At ionic strengths up to 0.1 mol/L, an extended version of the Debye-Hückel equation has to be used which includes a second solvent depending constant and the effective hydrated radius of the ion. High ionic strength shields the actual concentration of a solution, making the concentration of any salt appear lower. Thus, the higher the electrical charge of the present ions the lower the effective concentration. In Figure 6 activity coefficients for a number of ions in water solution are presented estimated by the extended Debye-Hückel equation.

Apart from the high content of salts and ions in the converted sludge, ions are also provided by addition of acids. According to the stated assumption, acidulation with a polyprotic acid would result in lower effective concentration than a monoprotic acid. Consequently, comparison of resulting salt concentration will partly depend on the charge of the anion accompanying the acid, i.e. a higher shielding and thereby release would be achieved if sulphuric acid was used compared to hydrochloric acid. For example, for an ionic strength of 0.01, the activity coefficient is 0.65 for sulphate and 0.84 for chloride according to Figure 6. The lower the activity coefficient the lower the effective concentration. Furthermore, metal ions of different charge also reduce the activity drastically. For the same ionic strength, iron(II) and calcium(II) have an activity coefficient of 0.68, while iron(III) and aluminium(III) have activity coefficients of 0.44. Even though the metal ions of a +3 charge result in lower activity, acid concentrations are high so the effect of the metal ions on the activity is probably comparably small.