w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

Available online at www.sciencedirect.com

ScienceDirect journal homepage: www.elsevier.com/locate/watres

Dewatering in biological wastewater treatment: A review Morten Lykkegaard Christensen*, Kristian Keiding, Per Halkjær Nielsen, Mads Koustrup Jørgensen Department of Chemistry and Bioscience, Aalborg University, Frederiks Bajers Vej 7H, DK-9220 Aalborg East, Denmark

article info


Article history:

Biological wastewater treatment removes organic materials, nitrogen, and phosphorus

Received 2 February 2015

from wastewater using microbial biomass (activated sludge, biofilm, granules) which is

Received in revised form

separated from the liquid in a clarifier or by a membrane. Part of this biomass (excess

15 April 2015

sludge) is transported to digesters for bioenergy production and then dewatered, it is

Accepted 17 April 2015

dewatered directly, often by using belt filters or decanter centrifuges before further

Available online xxx

handling, or it is dewatered by sludge mineralization beds. Sludge is generally difficult to dewater, but great variations in dewaterability are observed for sludges from different


wastewater treatment plants as a consequence of differences in plant design and physical-


chemical factors. This review gives an overview of key parameters affecting sludge dew-


atering, i.e. filtration and consolidation. The best dewaterability is observed for activated


sludge that contains strong, compact flocs without single cells and dissolved extracellular

Activated sludge

polymeric substances. Polyvalent ions such as calcium ions improve floc strength and


dewaterability, whereas sodium ions (e.g. from road salt, sea water intrusion, and industry) reduce dewaterability because flocs disintegrate at high conductivity. Dewaterability dramatically decreases at high pH due to floc disintegration. Storage under anaerobic conditions lowers dewaterability. High shear levels destroy the flocs and reduce dewaterability. Thus, pumping and mixing should be gentle and in pipes without sharp bends. © 2015 Elsevier Ltd. All rights reserved.



Municipal and industrial wastewater contain high amounts of COD, nitrogen, and phosphorus, which are usually degraded or removed by biological wastewater treatment (Lindrea and Seviour, 2002). The activated sludge process is by far the most common process, but alternative processes such as biofilm systems or granules systems also exist (de Bruin et al.,

2004). An integrated part of the biological wastewater treatment is the solideliquid separation, where the treated water is separated from the activated sludge. In the conventional activated sludge process, this is done by clarifiers, but there is an alternative: membrane bioreactors, where a membrane is used instead of the clarifier (Brindle and Stephenson, 1996; Lindrea and Seviour, 2002). The outcome of the process is treated wastewater (effluent), return sludge, and excess sludge.

* Corresponding author. Tel.: þ45 9940 8464. E-mail address: [email protected] (M.L. Christensen). http://dx.doi.org/10.1016/j.watres.2015.04.019 0043-1354/© 2015 Elsevier Ltd. All rights reserved.

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019


w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

Fig. 1 e Overview of parameters that directly or indirectly influence sludge properties.

In some cases, excess sludge is transported to digesters for sludge reduction and bioenergy production. However, in many cases, other types of sludge handling takes place, e.g. transportation to agricultural fields or drying and incineration. Since the water content of excess sludge is high, it must be dew before further handling, typically by belt filters, filter press, decanter centrifuges, and sludge mineralization beds (Sørensen and Sørensen, 1997). Thus, several solideliquid separation processes are involved in wastewater treatment for separating sludge from the treated wastewater as well as for sludge dewatering. The dewatering process is costly, and the composition and properties of the sludge are important for the separation process (Bruus et al., 1992; Sørensen and Sørensen, 1997; Chu et al., 2005). This paper reviews the existing literature on sludge dewaterability, i.e. sludge filtration and consolidation. Fig. 1 summarizes the key parameters that affect various sludge properties such ad dewaterability. Sludge contains flocs, and sludge properties are mainly determined by the size, shape, density and strength of the sludge flocs. Thus, an understanding of the sludge flocs is crucial for a more general understanding of sludge dewatering. Flocs, on the other hand, consist of microorganisms, extracellular polymeric substances (EPS), organic debris and inorganic particles. Some of the components are produced during the biological process and some of the components come from the influent. Further, floc density and strength are influenced the content of e.g. catons and inorganic particle and also by shear forces and thereby indirectly by the design and operation of the plant. The floc properties not only influence sludge filtration and consolidation but also other processes such as flocculation, settling and membrane fouling, i.e, literature data show that sludge components that cause problems in filtration and consolidation also cause problems in other types of separation processes (e.g. sedimentation, centrifugation, sludge mineralization bed, and membrane bioreactors). Thus, many of the conclusions from this paper are of generic value for all solideliquid separation processes for biological sludges.


adsorption of particles from the influent. The flocs consist of microorganisms, either as single cells, filamentous bacteria or microcolonies, organic fibers, inorganic particles (salt and sand), and extracellular polymeric substances (EPS). The typical size of the flocs is 129 ± 109 mm (Mikkelsen and Keiding, 2002) e see sketch of a typical sludge floc in Fig. 2. Sludge flocs have a fractal-like structure and are kept together by DLVO forces (van der Waals and electrostatic forces), non-DLVO-forces (bridging, hydrophobic forces), and physical entanglement (Namer and Ganczarczyk, 1994; Cousin and Ganczarczyk 1999; Nielsen, 2002). EPS components are particularly important for the floc properties. The EPS components are a mixture of different macromolecules, e.g. proteins, humic-like substances, polysaccharides, nucleic acids and lipids and contribute with 40e60% of the total dry matter of the flocs (Nielsen, 2002). They are negatively charged, and the charge density has been measured to be 0.2e1 meq/g EPS (Keiding et al., 2001; Mikkelsen and Keiding, 2002; Reynaud et al., 2012). Different methods exist for EPS extraction and analyses, and it is often difficult to compare literature data. Nevertheless, it is generally accepted that EPS can be classified as tightly bound EPS (TBEPS), loosely bound EPS (LBEPS), and suspended EPS. Further, a dynamic equilibrium has often been found between loosely bound and suspended EPS components (Nielsen and Jahn, 1999; Comte et al., 2006; Dominguez et al., 2010). The electrostatic interaction

Sludge composition

Biological activated sludge consists primarily of biological flocs that are formed by growth of microorganisms and by

Fig. 2 e Schematic picture of activated sludge flocs (the ideal floc) from Nielsen et al. (2012).

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019

w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

plays an important role for the equilibrium, and for this reason, the concentration and valence of ions play a major role for the floc structure, which will be discussed later in more detail. Sludge flocs contain large amounts of water; the reported water content varies from 63% to 99% (Andreadakis, 1993; Chung and Lee, 2003; Vaxelaire and Cezac, 2004). The water in the flocs and at the surface of the flocs is often denoted bound water as opposed to the free water, which is not affected by the solid particles (Vesilind, 1994). Further, bound water has been divided into three types of water pools: I) water trapped inside crevices and the interstitial space of the flocs (interstitial water), II) water physically bound to surfaces (vicinal water), and III) water chemically bound to solid materials (water of hydration). Alternatively, the high water content in flocs has been explained as a consequence of the colligative properties, i.e. the reduced water activity in the floc interior due to counter ions (osmotic water) (Keiding et al., 2001). Mikkelsen and Keiding (2002) use the term “water-holding” for the surface bound water, the osmostic water, and the trapped water. However, in both cases, the flocs contain water, some of which is removed during compression, i.e. according to Deng et al. (2011), interstitial water accounts for more than 50% of the water in the flocs and is at least partly removed by mechanical dewatering (Novak, 2006). The dewatering process therefore depends on the strength of the floc structure. The floc structure can vary from large compact flocs (the ideal floc), flocs with high abundance of filamentous bacteria (filamentous bulking), or small, light flocs without filamentous bacteria (pinpoint floc). In rare cases, no or few flocs are formed with many single bacteria (dispersed growth). Generally, the best separation properties are obtained if the sludge contains large compact flocs, few filamentous bacteria and few single cells (Bruus et al., 1992; Rasmussen et al., 1994). This gives the best settling in the clarifier, the highest permeate flux in MBR systems (lowest fouling), the highest filterability (belt filters and sludge mineralization bed), and the best effluent quality (decanter centrifuges) and lowers the amount of chemicals required for sludge conditioning (Lee et al., 2003; Masse et al., 2006; Dominiak et al., 2011a; Bugge et al., 2013). However, there are some important differences between the separation processes. The filamentous bacteria, for example, are important for settling and sludge mineralization (drainage) but not filtration and consolidation where higher pressures is applied for compression (Dominiak et al., 2011a; Bugge et al., 2013). The species composition of the activated sludge influences the floc properties to a certain extent and thus the solideliquid separation processes (Nielsen et al., 2002, 2004; Klausen et al., 2004; Larsen et al., 2006, 2008; Bugge et al., 2013). Some species form filaments, some strong microcolonies, and some weak flocs. They also produce different amounts and type of EPS with different water-binding properties. The variation observed in solideliquid separation processes in different treatment plants (see later) is therefore caused by variations in both microbial composition and water/floc chemistry. Recent studies by molecular DNA-based methods have furthermore revealed that, despite presence of numerous bacterial species in the wastewater treatment plants, the dominant and abundant ones can be found among


only approx. 150 species that are present in most plants (called core species) (Nielsen et al., 2010, 2012). These are now studied in great detail to understand their identity, physiology, ecology, impact on floc properties, and their possibilities of manipulation of the community composition and design of good solideliquid separation processes (McIlroy et al., 2015; see for example the open resource http:// midasfieldguide.org/).


Specific filtration flow rate

Several methods exist for comparing dewaterability of different types of sludge such as capillary suction time, sludge volume index, average specific resistance of the cake, and the specific filtration flow rate. The specific filtration flow rate (SFF) is a useful term especially for filtration and consolidation processes. Dewatering often involves both filtration (cake formation) and consolidation (cake compression), and for biological sludge it is difficult to distinguish between the two processes (Stickland et al., 2005). Thus, the term dewaterability is here used to describe the rate of both the filtration and consolidation. When SFF is used, the dewaterability is determined by the liquid flow through a cake consisting of the solid materials from the suspension. The liquid flow through a cake structure can be calculated by using Eq. (1) if the filter medium resistance is low: q¼

p maav uc


where q [m3/(m2s)] is the filtrate flux, m (Pa s) is the filtrate viscosity, p (Pa) is the filtration pressure, uc (kg/m2) is the mass of solid materials per unit filter media area, and aav (m/kg) is the average specific resistance of the cake. The average specific resistance is independent of filtration pressure for incompressible cakes. However, most cakes are compressible, i.e. cake porosity decreases with increasing pressure, whereby the average specific resistance increases as well. There exist several constitutive equations describing this relationship between average specific resistance and pressure, one such equation has been suggested by Tiller and Yeh (1987):  n p aav ¼ a0 1 þ ps


where a0 (m/kg) is the average specific resistance at zero pressure, and ps (Pa) and n () are empirical constants. The constitutive equation was originally developed for the local specific cake resistance, but is also applicable as an empirical equation for calculating the average specific cake resistance. For sludge, the average specific resistance usually increases almost linearly with pressure (Sørensen and Sørensen, 1997), i.e. n ¼ 1 and ps ≪ p. Thus, due to the high compressibility, it is necessary to know the filtration pressure in order to compare measured literature values of average specific resistances. However, there exists an alternative and more useful way to characterize the liquid flow through a cake with high compressibility, the specific filtrate flow rate (SFF). This has been defined in Sørensen et al. (1996) and Sørensen and Sørensen (1997):

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019


w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

SFF ¼ quc ¼

p maav


The permeate flux is inversely proportional to cake thickness, so the product of flux and cake thickness (SFF) does not change during constant pressure filtrations. For noncompressible cakes, SFF increases proportionally with applied pressure, whereas the SFF value is constant and independent of the pressure for highly compressible cakes, as an increase in pressure will be equalized by the subsequent elevation of specific resistance. Thus, for biological sludge that forms highly compressible cakes, SFF is the better measure of the dewaterability than e.g. the often used average specific resistance. In Fig. 3, it is seen that SFF increases with pressure for kaolin, whereas it is almost constant for biological sludge at pressures above 2 kPa and equals 2.7$105 kg/(ms) (Fig. 3). The data confirm that SFF for compressible cakes is independent of pressure except at low pressure and can therefore be used to compare data from different types of sludges. The SFF has been measured for excess sludge from seven different wastewater treatment plants with nutrient removal in Denmark, and the data show that SFF varies by a factor of 10 for the different types of activated sludges (Fig. 4). Table 1 shows the structure of the flocs for the filtered sludges. The data confirm that large compact flocs give highest SFF and thus the highest dewaterability. The main conclusion is that sludge composition has a high impact on sludge dewaterability. Several other studies confirm that the dewaterability varies for different sludges (Karr and Keinath, 1978; Katsiris and Kouzeli-Katsiri, 1987; Novak et al., 1988; Cho et al., 2005; C ¸ ic¸ek et al., 1999; How et al., 2005). Thus, in order to understand how sludge properties affect specific filtration flow rate, cake structure and compression will be discussed in more details.


Sludge cake compressibility and blinding

When the sludge cake is compressible, it means that the cake porosity, ε, decreases with increasing pressure, resulting in

Fig. 3 e Specific filtrate flow rate for kaolin and excess sludge, recalculated from Sørensen and Sørensen (1997).

Fig. 4 e Specific filtrate flow rate for activated sludge from 7 different full-scale wastewater treatment plants in Denmark, recalculated from Dominiak et al. (2011a).

higher resistance. This can be modeled by the following constitutive equations (Tiller and Yeh, 1987):  b p ð1  εÞ ¼ ð1  ε0 Þ 1 þ ps


where b is an empirical parameter and ε0 is the porosity of an uncompressed cake. The equation is similar to the one used for the average specific cake resistance, and combining the two equations gives aav ¼ k(1ε)m, where k is the ratio between a0 and (1ε0), and m is the ratio between n and b. Eq. (4) gives the cake porosity and thereby the final dry matter content of the cake, which for compressible cakes increases with applied pressure. Several mechanisms have been suggested to explain the porosity reduction. These mechanisms are summarized in Table 2 and Fig. 5. Many studies have focused on the filtration of inorganic particles, where the compressibility is generally well described and understood. Large inorganic particles (>10 mm) usually form incompressible cakes, and the porosity of the cake is mainly dependent on particle structure (Tiller and Yeh, 1987). Cakes consisting of colloidal particles may be compressible, depending on the degree of flocculation, i.e. highly flocculated colloidal particles form cakes with high porosity and high compressibility (Tiller and Yeh, 1987). The degree of flocculation depends on particle surface properties (mainly charge density) and physico-chemical properties of the suspension, e.g. ionic strength. During consolidation (compression) of inorganic cakes, two consolidation stages have been observed, of which one has been ascribed to the collapse of the global cake structure (Point 1, Table 2), and one ascribed to particle migration into a more stable configuration (Point 2, Table 2) (Shirato et al., 1986; Chu and Lee, 1999; Xu et al., 2004). The global cake collapse is controlled by the hydraulic resistance of the cake (Shirato et al., 1986). Particle migration is controlled by the highly viscous surface-absorbed

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019


w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

Table 1 e Sludge and floc properties (Dominiak et al., 2011a). Microscope analysis

Relative SFFa

Large compact, round, dark flocs Large, regular compact flocs Medium-sized flocs, both round, regular and open irregular Open irregular, medium-sized flocs Medium-sized flocs, both compact and open Very small, irregular, disintegrated flocs, many branched filamentous bacteria Small, irregular flocs of open structure

100 38 24 21 16 12

Plant Bramming South Esbjerg West Hjørring Esbjerg East Aalborg East Bramming North Aalborg West a

water between the particles and is a slow process, compared to the overall collapse of cake structure (Chu and Lee, 1999). Compression of cakes consisting of colloidal particles is also partly due to the reduction of the distance between the particles (Point 3, Table 2) (Koenders and Wakeman, 1997; Keiding and Rasmussen, 2003). The distance between particles is a function of the electrostatic repulsion, van der Waal attraction, and the external pressure (Koenders and Wakeman, 1997). For biological sludges, the compressibility is less well described and understood, but it is believed that the same mechanisms are relevant as for inorganic sludge (Chu and Lee, 1999). Furthermore, organic sludge consists of soft water-swollen materials; thus, deformation and compression of individual particles are important as well (Point 4, Table 2). The effect of compression of water-swollen particles has been investigated by synthesizing and filtering model particles. A study of synthetic polystyrene-co-poly(acrylic acid) shows that the soft polyacrylic acid shell deforms and compresses during filtration, which lowers the specific flow rate by a factor of 10e100 because the soft materials fill out the void between the particles (Lorenzen et al., 2014). The specific flow rate is low, compared to inorganic particles of the same particle size, but the SFF values measured are comparable with values found for biological sludge (Lorenzen et al., 2014). Furthermore, a relatively large reduction of the cake water content is observed during consolidation of both sludge and the synthetic polystyrene-co-poly(acrylic acid) particles (Christensen and Keiding, 2007; Christensen and Hinge, 2008). Hence, compression and deformation of individual particles explain the high compressibility for sludge (Hwang and Hsueh, 2003). For suspensions containing particles with large particle size distribution, small particles may be trapped within the

Table 2 e Cake structure phenomena in filtrations. 1

2 3 4 5


Relative SFF setting Bramming South to 100.

Collapse of global structure and dissipation of excess pore water. This includes bending and slipping of fibers. Particle migration into a more stable configuration. Reduction of inter-particular distance. Deformation and compression of individual particles. Cake blinding from small particles from the suspension (a) or from disintegration of flocs or individual particles in the cake (b).

cake pores, often denoted cake blinding (Point 5, Table 2) (Christensen and Dick, 1985; Sørensen et al., 1995). Sometimes filtration cannot be described using the traditional filtration theory. Such data are observed for sludge cakes and has been explained as an effect of cake blinding, i.e. small particles seem to be particularly important for the dewaterability of sludge. Cake blinding can also be observed due to floc disrupture and erosion (Sørensen et al., 1995). Thus, not only particle size, degree of aggregation and structure are important for dewaterability of sludge, but also the presence of small particles, water content of the flocs, and floc strength. Several factors influence floc properties, concentration of single cells, and suspended EPS. The physico-chemical properties of the suspension are important for the sludge properties and especially the ions in the solution.


Conductivity and water hardness

The composition of inlet wastewater varies from plant to plant, e.g. due to different industries, rainfall, etc. This affects both the biological wastewater treatment and the dewaterability of the biological sludge produced. Several studies have shown that the wastewater conductivity, water hardness, and pH vary; i.e. ionic composition and concentrations vary. This strongly affects the dewaterability of the biological sludge. Both floc structure and strength strongly depend on ionic composition and concentration. High concentrations of multivalent cations, such as calcium and magnesium, give strong and compact flocs (Biggs et al., 2001; Higgins et al., 2004a; Larsen et al., 2008). Data show that the porosity of the flocs decreases with increasing calcium ion concentration (Cousin and Ganczarczyk, 1999). Conversely, monovalent cations such as sodium and potassium lower floc strength (Higgins and Novak, 1997; Biggs et al., 2001). Different theories have been suggested to explain the role of cations in sludge flocculation, e.g. the alginate egg-box model, the DLVO theory, and divalent cation bridging, of which the divalent cation bridging model seems to describe the role of ions in sludge best (Sobeck and Higgins, 2002; Higgins et al., 2004a). According to the divalent bridging model, calcium and other divalent ions bridge the negatively charged sites on EPS and thereby form a matrix of EPS and single cells. Several studies have confirmed the positive effect that divalent ions have on floc structure and dewaterability. Activated sludge samples from different membrane bioreactor plants show that an increased

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019


w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

ion exchange between mono- and divalent ions, the ratio between monovalent (Mþ) and divalent cations (Dþþ) should, according to Higgins and Novak (1997), be lower than 2 on a meq/L basis to ensure good dewaterability. Others suggest that good sludge dewaterability is observed as long as Mþ/ Dþþ < 4 (Peeters et al., 2011). The literature therefore shows that water hardness (concentration of multivalent ions) is important for sludge dewaterability. Furthermore, addition of calcium ions will usually improve sludge dewaterability, and due to the beneficial effect of divalent ions, the literature has suggested to use alternatives to sodium-based chemicals, i.e. chemicals containing divalent cations instead of sodium (Higgins et al., 2004b). High conductivity due to monovalent ions reduces dewaterability, and this phenomenon can be observed in the northern part of Europe due to road salting during winter and intrusion of sea water. Moreover, wastewater from some types of industries has high conductivity. The typical conductivity of Danish activated sludge is 750 mS/cm, but values up to 4400 mS/cm have been observed (PH Nielsen, unpublished).


Fig. 5 e Illustration of mechanisms of cake compression and blinding described in Table 2.

ratio between divalent ions (calcium/iron) and EPS correlates with lower amounts of single cells in the bulk solution, increases the flocs sizes as well as the dewaterability (Bugge et al., 2013). Bruus et al. (1992) added EGTA to sludge in order to remove calcium ions from sludge flocs, resulting in destabilization of the flocs, increase of the concentration of single cells and soluble EPS, and thereby a reduction of the dewaterability. Peeters et al. (2011) showed that the exchangeable calcium fraction is about 0.7 meq/g MLVSS, whereas calcium is precipitated as calcium salts (e.g. calcium carbonate) within the floc structure at higher calcium concentrations. At high concentrations of monovalent ions, the divalent ions in the floc matrix are ion exchanged by monovalent ions, which weakens the floc structure. Bruus et al. (1992) showed that the concentration of calcium in the bulk increases after addition of monovalent salts, whereby the concentration of single cells increases, and the dewaterability drops. Due to the

Sludge pH

Activated sludge flocs contain a lot of EPS which contain titratable groups and are negatively charged at neutral pH. The EPS components are almost non-charged at pH around 2.6e3.6 (Liao et al., 2002), whereas the charge increases with pH (Raynaud et al., 2012). As EPS components and electrostatic forces play a central role in floc structure, sludge pH indirectly affects the floc structure and sludge dewaterability. At low pH, the bulk suspension only contains few colloidal particles, and the dewaterability of sludge is generally high (Karr and Keinath, 1978). At high pH, the number of colloidal particles and suspended EPS increases (floc disintegration), and the dewaterability drops (Karr and Keinath, 1978; Raynaud et al., 2012). By using data from Raynaud et al. (2012), it can be shown that the SFF values decrease from 4.4$105 kg/(ms) at pH 7 to 0.9$105 kg/(ms) at pH 9, which is a reduction of approximately 80%. The effect of adding acid (lowering pH) is not as pronounced as adding base (increasing pH). Raynaud et al. (2012) observed a small reduction in dewaterability by reducing pH to 3, where SFF was reduced to 4.2$105 kg/(ms), whereas Karr and Keinath (1978) observed higher dewaterability after addition of acid (pH ¼ 3). Liao et al. (2002) did not observe any change in dewaterability at low pH. However, the water content was reduced in the formed cake if the pH was lowered before filtration. Thus, pH affects dewaterability, and high pH value should be avoided.


Biological process

The solideliquid characteristics of the sludge is influenced by the wastewater composition and the way the sludge is produced, e.g. by the conventional activated sludge process, membrane filtration in MBR, biofilms, or by mesophilic and thermophilic digestion. Table 3 summarizes sludge characteristics and filtration properties from two surveys of sludge filtration properties in

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019


w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

terms of total EPS content, mean floc size, shear sensitivity for the release of particles during shear treatment, kSS, and SFF (Mikkelsen and Keiding, 2002; Bugge et al., 2013). The SFF of anaerobically digested sludge is low, compared with other types of sludges. Anaerobically digested sludge has low concentrations of EPS, compared with activated sludge, which seems to correlate with smaller flocs. Furthermore, the ionic strength and the number of single cells increase during anaerobic storage (Rasmussen et al., 1994). Sludge floc strength significantly decreases for anaerobically digested sludge (Table 3). Thus, floc deformation, compression, and disintegration are more pronounced in dewatering of anaerobically stored or digested sludge. The main difference between MBR and CAS system is the use of a membrane instead of a clarifier. Generally, MBR sludge has lower SFF, compared with CAS sludge (C ¸ ic¸ek et al., 1999; How et al., 2005). The low SFF of MBR sludge can be ascribed to the low degree of flocculation, differences in  et al., 2006, microbiology, and principle of separation (Masse Geng and Hall, 2007, Lee et al., 2003, Van den Broeck et al., 2010; Van den Broeck et al., 2012). In CAS treatment, there is a selection for flocculated bacteria, while dispersed bacteria are removed with the effluent. For MBR sludge, dispersed bacteria are too large to penetrate the pores of the membrane. Therefore, MBR sludge shows a higher content of single cells and soluble EPS, compared to CAS sludge (Wagner et al., 2000;  et al., 2006; Merlo et al., 2004; Witzig et al., 2002; Masse  randio et al., 2005), whereas CAS sludge has a higher Spe content of EPS (Table 3), which is important for bioflocculation  et al., 2006). Furthermore, the higher (Merlo et al., 2004; Masse level of shear in an MBR tank induced by e.g. air scouring of the membranes does not allow large flocs to form (C ¸ ic¸ek et al., 1999). Thus, the floc size is usually lower in MBR sludge than in other types of sludges. Lower floc sizes result in higher specific resistance to filtration due to cake blinding, smaller flocs, and less permeable and more compressible cakes (Lee et al., 2003). Thus, SFF can be a factor of three lower for MBR sludge than CAS sludge (C¸ic¸ek et al., 1999).


Sludge storage

Sludge is often stored before dewatering. However, the biological processes do not stop during storage; thus, floc

structure and composition change, which in turn affects and often reduces dewaterability (Bruus et al., 1993). Several factors are involved in these changes such as hydrolysis of EPS components, reduction of Fe(III) to Fe(II), which is a poorer flocculant, and production of sulfide by microbial sulfate reduction that subsequently precipitates and removes Fe (III) and Fe(II) (Nielsen and Keiding, 1998; Wilen et al., 2000a,b). A study of 10 days' anaerobic storage of sludge shows a significant increase in the number of single cells, conductivity, and bulk calcium concentration (Rasmussen et al., 1994). As discussed in the previous section, this reduces dewaterability, and in the cited study, anaerobic storage reduces the specific flow rate by 80%. Mixing of anaerobically stored sludge further lowers the SFF value (Parker et al., 1972; Larsen et al., 2006). The negative effect of storage can be limited by ensuring aerobic or anoxic conditions during storage, e.g. by aeration or addition of nitrate (Dominiak et al., 2011a). The reduced SFF after anaerobic storage can to some extent be improved again by aeration, i.e. ensuring aerobic storage (Parker et al., 1972; n et al., 2000b). The negative consequence of anaerobic Wile storage may also be important for the activated sludge process; there may exist anaerobic zones in the plants, which impairs the sludge, causes deflocculation and thereby reduces dewaterability of the biological sludge produced.


Pumping and stirring of sludge

Sludge flocs can be destroyed due to high shear levels which reduce sludge dewaterability. Particles and sludge flocs aggregate under low shear rates and break up at elevated shear rates (Mikkelsen and Keiding, 1999, 2002). Break-up of sludge flocs (fragmentation) lower the mean size of the flocs (Jarvis et al., 2005). At higher shear rates, smaller particles (e.g. single cells) are desorbed from the floc surface due to erosion (Mikkelsen and Keiding, 2002; Biggs et al., 2003). Both floc size and especially the number of single cells affect the dewaterability. The specific flow rate is reduced after vigorous stirring of the sludge (Dominiak et al., 2011b), as the increased number of single cells and lower particle sizes lead to cake blinding. The negative effect of high shear depends on the floc strength, i.e. the floc resistance to stirring. It has been shown that calcium ions reduce the effect of shear. Conversely, anaerobic storage results in weak flocs that easily break up during high shear (Rasmussen

Table 3 e Physical-chemical characteristics of primary, activated, digester (mesophilic, thermophilic) feed with surplus activated sludge and MBR sludge.

Total protein (mg/gSS) Total humics (mg/gSS) Total polysaccharides (mg/gSS) EPS (mg/gSS) Mean floc size (mm) Shear sensitivity, kSS SFF (kg/(m∙s)) a b

Activated sludgea

MBR sludgeb

Mesophilic sludgea

Thermophilic sludgea

346 ± 111 58 ± 35 101 ± 35 130 ± 65 125 ± 109 0.062 ± 0.049 83.3$107

185 ± 45 22 ± 8 111 ± 13 89 ± 11 65 ± 23 0.102 ± 0.066 72.9$107

248 ± 12 112 ± 108 70 ± 5 78 ± 49 51 ± 21 0.244 ± 0.016 9.7$107

155 ± 62 188 ± 92 78 ± 10 41 ± 9 57 ± 11 0.418 ± 0.337 0.78$107

Data from Mikkelsen and Keiding (2002). Data from Bugge et al. (2013).

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019


w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

et al., 1994; Larsen et al., 2006, 2008). In general, viscous shear should be avoided in order to ensure a high specific filtrate flow, e.g. by ensuring gentle pumping, gentle mixing, no storage in tank or pipes, and sharp pipe bends should also be avoided.

10. Summery of factors that influence sludge quality Thus, several parameters affect the dewaterability of sludge: the physico-chemical properties of the feed, the biological treatment, and the handling of the sludge before and during dewatering. Table 4 summarize the conclusions from the text. The composition of the incoming wastewater affects the properties of the sludge produced, especially the organic compounds, pH, and the ion composition. The biological process and the plant design as well as the further sludge handling (pumping, mixing, and storage) are important for the sludge flocs and the dewaterability of sludge.

11. Improvement of sludge filterability by flocculation Sludge dewatering is an expensive operation in wastewater treatment plants. It is not possible to improve the dewatering process by applying higher pressure in filtration processes due to the high compressibility of sludge cakes. Instead, sludge can be pre-treated by adding coagulants e.g. polyaluminium chloride (PAC) or ferric salts (FeSO4Cl), followed by addition of flocculants or by adding flocculants alone. This improves sludge dewaterability significantly and reduces the costs of the separation process. It should be mentioned that addition of inorganic salts may have negative effects on further sludge handling e.g. if the sludge is incinerated or if phosphorus from the sludge has to be reused as a fertilizer. The degree of sludge flocculation is enhanced by addition of coagulants and flocculants. Addition of e.g. ferric chloride, and thereby positively charged multivalent metal ions (Fe3þ), strengthens the floc structure and removes EPS and single n et al., 2008; Niu cells from the bulk (Poon and Chu, 1999; Wile

et al., 2013). Multivalent cations adsorb to surfaces, which reduces the electrostatic repulsion between negatively charged particles, e.g. flocs and single cells, whereby they aggregate (Niu et al., 2013). Addition of polyvalent cationic polymers enhances flocculation by charge neutralization and polymer bridging (Bolton and Gregory, 2007). Data show that addition of flocculant with efficient mixing increases floc size, increases SFF, and lowers cake compressibility (Chu et al., 2003; Chen et al., n et al., 2008). Low dosages do not provide suffi2005; Wile cient charge neutralization/polymer bridging for flocculation to be efficient, whereas very high concentrations lead to deflocculation due to charge inversion and/or steric hindrance, which demonstrates that an optimum dosage of flocculants exists (Abo-Orf and Dentel, 1997; Poon and Chu, 1999; Lee and Liu, 2000; Yen et al., 2002; Chu et al., 2003; Chen et al., 2005). The optimum dosage of polyelectrolytes increases with concentration of suspended materials and the concentration of single cells and EPS (Tiravanti et al., 1985; Mikkelsen and Keiding, 2001). The optimum dosage of cationic polymer for flocculation of municipal CAS sludge has been reported to be in the range of 0.01e0.06 mg/g SS (Yen et al., 2002; Chen et al., 2005). When colloidal material is released from the flocs due to factors such as shear or anaerobic conditions, higher dosages of polyelectrolytes are required (Mikkelsen et al., 1996; Abu-Orf and Dentel, 1997). In general, sludge with low dewaterability also requires a higher dosage of polymers.



Great variation in sludge dewaterability is observed among wastewater treatment plants; hence the floc and sludge properties have a high impact on the specific filtrate flow rate. The best dewaterability is observed for sludge that contains strong compact flocs and low concentrations of single cells as well as dissolved EPS. This gives the best sedimentation in the clarifier, the highest permeate flux in MBR systems, the highest filterability (belt filters and sludge mineralization bed), the best effluent quality (decanter centrifuges), and lowers the

Table 4 e Link between sludge treatment and dewaterability. Parameter



Changes in conductivity (high conductivity or dilution) lower specific flow rate This can be a problem due to road salting, intrusion of sea water and some industries High water hardness improves specific flow rate Calcium carbonate can be added to improved dewaterability High pH leads to floc disintegration, which lowers the specific flow rate The water content in the formed filter cake may be lower if the pH value is lowered. Anaerobic storage lowers specific flow rate Tanks/pipes with anaerobic pockets are problematic Addition of nitrate during storage or aeration can improve the filterability Vigorous pumping lowers specific flow rate Gentle pumping and mixing is recommended. Avoid sharp bends on pipes Conventional plant usually gives better sludge than membrane bioreactors (MBR). Sludge from digesters is difficult to dewater.

Water hardness pH


Pumping Treatment system

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019

w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

amount of chemicals required for sludge conditioning. High water hardness improves dewaterability because calcium ions improve floc strength and reduce the concentration of single cells and EPS. Variations in conductivity, and particularly high conductivity and pH, reduce dewaterability as flocs disintegrate. Thus road salt in the winter season, intrusion of sea water and some types of industries can result in lower dewaterability. Anaerobic storage lowers dewaterability as flocs are disintegrated and the conductivity increases. Anaerobic storage or tanks with anaerobic pockets are more problematic than aerobic or anoxic storage. High shear in pumps and pipes destroys the flocs and reduces dewaterability and should be avoided. The physico-chemical properties of biological sludge cakes govern the sludge dewaterability; hence, the filtration processes of biological sludge should be improved by improving sludge physico-chemical characteristics.


Abu-Orf, M.M., Dentel, S.K., 1997. Effect of mixing on the rheological characteristics of conditioned sludge: full-scale studies. Water Sci. Tech. 36, 51e60. Andreadakis, A.D., 1993. Physical and chemical properties of activated sludge floc. Water Res. 27, 1707e1714. Biggs, C.A., Ford, A.M., Lant, P.A., 2001. Activated sludge flocculation: direct determination of the effect of calcium ions. Water Res. 43, 75e80. Biggs, C., Lant, P., Hounslow, M., 2003. Modelling the effect of shear history on activated sludge flocculation. Water Sci. Tech. 47, 251e257. Bolton, B., Gregory, J., 2007. Organic polyelectrolytes in water treatment. Water Res. 41, 2301e2324. Brindle, K., Stephenson, T., 1996. The application of membrane biological reactors for the treatment of wastewaters. Biotechnol. Bioeng. 49, 601e610. Bruus, J.H., Nielsen, P.H., Keiding, K., 1992. On the stability of activated sludge flocs with implications to dewatering. Water Res. 26, 1597e1604. Bruus, J.H., Christensen, J.R., Rasmussen, H., 1993. Anaerobic storage of activated sludge: effects on conditioning and dewatering performance. Water Sci. Tech. 28, 109e116. Bugge, T.V., Larsen, P., Saunders, A.M., Kragelund, C., Wybrandt, L., Keiding, K., Christensen, M.L., Nielsen, P.H., 2013. Filtration properties of activated sludge in municipal MBR wastewater treatment plants are related to microbial community structure. Water Res. 47, 6719e6730. Chen, B.-H., Lee, S.-J., Lee, D.-J., 2005. Rheological characteristics of the cationic polyelectrolyte flocculated wastewater sludge. Water Res. 39, 4429e4435. Cho, J., Song, K.-G., Ahn, K.-H., 2005. The activated sludge and microbial substances influences on membrane fouling in submerged membrane bioreactor: unstirred batch cell test. Desalination 183, 425e429. Christensen, G.L., Dick, R.I., 1985. Specific resistance measurements: nonparabolic data. J. Environ. Eng. ASCE 111, 243e257. Christensen, M.L., Hinge, M., 2008. The influence of creep on cake solid volume fraction during filtration of coreeshell particles. Colloid Surf. A: Physicochem. Eng. Asp. 320, 227e232. Christensen, M.L., Keiding, K., 2007. Creep effects in activated sludge filter cakes. Powder Technol. 177, 23e33. Chu, C.P., Lee, D.J., 1999. Three stages of consolidation dewatering of sludges. J. Environ. Eng. ASCE 125, 959e965.


Chu, C.P., Lee, D.J., Tay, J.H., 2003. Gravitational sedimentation of flocculated waste activated sludge. Water Res. 37, 155e163. Chu, C.P., Lee, D.J., Chang, C.Y., 2005. Energy demand in sludge dewatering. Water Res. 59, 1858e1868. Chung, H.Y., Lee, D.J., 2003. Porosity and interior structure of flocculated activated sludge flocs. J. Colloid Interface Sci. 267, 136e143. C¸ic¸ek, N., Franco, J.P., Suidan, M.T., Urbain, V., Manem, J., 1999. Characterization and comparison of a membrane bioreactor and a conventional activated-sludge system in the treatment of wastewater containing high-molecular-weight compounds. Water Environ. Res. 71, 64e70. Comte, S., Guibaud, G., Baudu, M., 2006. Relations between extraction protocols for activated sludge extracellular polymeric substances (EPS) and EPS complexation properties Part I. Comparison of the efficiency of eight EPS extraction methods. Enzyme Microb. Technol. 38, 237e245. Cousin, C.P., Ganczarczyk, J.J., 1999. Effect of calcium ion concentration on the structure of activated sludge flocs. Environ. Technol. 20, 1129e1138. de Bruin, L.M.M., de Kreuk, M.K., van der Roest, H.F.R., Uijterlinde, C., van Loosdrecht, M.C.M., 2004. Aerobic granular sludge technology: an alternative to activated sludge. Water Sci. Technol. 49, 1e7. Deng, W., Li, X., Yan, J., Wang, F., Chi, Y., Cen, K., 2011. Moisture distribution in sludges based on different testing methods. J. Environ. Sci. 23, 875e880. Dominguez, L., Rodrı´guez, M., Prats, D., 2010. Effect of different extraction methods on bound EPS from MBR sludges. Part 2: influence of extraction methods over molecular weight distribution. Desalination 262, 106e109. Dominiak, D., Christensen, M.L., Keiding, K., Nielsen, P.H., 2011a. Sludge quality aspects of full-scale reed bed drainage. Water Res. 45, 6453e6460. Dominiak, D., Christensen, M.L., Keiding, K., Nielsen, P.H., 2011b. Gravity drainage of activated sludge: new experimental method and considerations of settling velocity, specific cake resistance and cake compressibility. Water Res. 45, 1941e1950. Geng, Z., Hall, E.R., 2007. A comparative study of fouling-related properties of sludge from conventional and membrane enhanced biological phosphorus removal processes. Water Res. 41, 4329e4338. Higgins, M.J., Novak, J.T., 1997. The effect of cations on the settling and dewatering of activated sludges: laboratory results. Water Environ. Res. 69, 215e224. Higgins, M.J., Tom, L.A., Sobeck, D.C., 2004a. Case study I: application of the divalent cation bridging theory to improve biofloc properties and industrial activated sludge system performance e direct addition of divalent ions. Wat. Environ. Res. 76, 344e352. Higgins, M.J., Sobeck, D.C., Owens, S.J., Szabo, L.M., 2004b. Case study II: application of the divalent cation bridging theory to improve biofloc properties and industrial activated sludge system performance e using alternatives to sodium-based chemicals. Wat. Environ. Res. 76, 353e359. How, Y., Ng, H.Y., Hermanowicz, S.W., 2005. Specific resistance to filtration of biomass from membrane bioreactor reactor and activated sludge: effects of exocellular polymeric substances and dispersed microorganisms. Water Environ. Res. 77 (2), 187e192. Hwang, K.-J., Hsueh, C.-J.L., 2003. Dynamic analysis of cake properties in microfiltration of soft colloids. J. Membr. Sci. 214, 259e273. Jarvis, P., Jefferson, B., Gregory, J., Parsons, S.A., 2005. A review of floc strength and breakage. Water Res. 39, 3121e3137. Karr, P.R., Keinath, T.M., 1978. Influence of particle size on sludge dewaterability. J. Water Pollut. Control Fed. 50 (8), 1911e1930.

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019


w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

Katsiris, N., Kouzeli-Katsiri, A., 1987. Bound water content of biological sludges in relation to filtration and dewatering. Water Res. 21 (11), 1319e1327. Keiding, K., Rasmussen, M.R., 2003. Osmotic effects in sludge dewatering. Adv. Env. Res. 7, 641e645. Keiding, K., Wybrandt, L., Nielsen, P.H., 2001. Remember the water e a comment on EPS colligative properties. Water Sci. Technol. 43 (6), 17e23. Klausen, M.M., Thomsen, T.R., Nielsen, J.L., Mikkelsen, L.H., Nielsen, P.H., 2004. Variations in microcolony-strength of probe-defined bacteria in activated sludge. FEMS Microbiol. Ecol. 50 (2), 123e132. Koenders, M.A., Wakeman, R.J., 1997. Filter cake formation from structured suspensions. Trans. IChemE 75, 309e320. Larsen, P., Eriksen, P.S., Lou, M.A., Thomsen, T.R., Kong, Y.H., Nielsen, J.L., Nielsen, P.H., 2006. Floc-forming properties of polyphosphate accumulating organisms in activated sludge. Water Sci. Technol. 54, 257e265. Larsen, P., Nielsen, J.L., Svendsen, T.C., Nielsen, P.H., 2008. Adhesion characteristics of nitrifying bacteria in activated sludge. Water Res. 24, 2814e2826. Lee, C.H., Liu, J.C., 2000. Enhanced sludge dewatering by dual polyelectrolytes conditioning. Water Res. 34, 4430e4436. Lee, S.A., Fane, A.G., Amal, R., Waite, T.D., 2003. The effect of floc size and structure on specific cake resistance and compressibility in dead-end microfiltration. Sep. Sci. Technol. 38, 869e887. Liao, B.Q., Allen, D.G., Leppard, G.G., Droppo, I.G., Liss, S.N., 2002. Interparticle interactions affecting the stability of sludge flocs. J. Colloid Interface Sci. 249, 372e380. Lindrea, K.C., Seviour, R.J., 2002. Activated sludge-the process. In: Bitton, G. (Ed.), Encyclopedia of Environmental Microbiology. Wiley, Chicester, United Kingdom. Lorenzen, S., Hinge, M., Christensen, M.L., Keiding, K., 2014. Filtration of core-shell colloids in studying the dewatering properties of water-swollen materials. Chem. Eng. Sci. 116, 558e566. , A., Spe randio, M., Cabassud, C., 2006. Comparison of Masse sludge characteristics and performance of a submerged membrane bioreactor and an activated sludge process at high solids retention time. Water Res. 40, 2405e2415. McIlroy, S.J., Saunders, A.M., Albertsen, M., Nierychlo, M., McIlroy, B., Hansen, A.A., Karst, S.M., Nielsen, J.L., Nielsen, P.H., 2015. MiDAS: the Field Guide to the Microbes of Activated Sludge. Database (accepted). Merlo, R.P., Shane Trussell, R., Hermanowicz, S.W., Jenkins, D., 2004. Physical, Chemical and Biological Properties of Submerged Membrane Bioreactor and Conventional Activated Sludges. WEFTEC, pp. 1e18. Mikkelsen, L.H., Keiding, K., 1999. Equilibrium aspects of the effects of shear and solids content on aggregate deflocculation. Adv. Colloid Interface 80, 151e182. Mikkelsen, L.H., Keiding, K., 2001. Effects of solids concentration on activated sludge deflocculation, conditioning and dewatering. Water Sci. Technol. 44, 417e425. Mikkelsen, L.H., Keiding, K., 2002. Physico-chemical characteristics of full scale sewage sludges with implication to dewatering. Water Res. 36, 2451e2462. Mikkelsen, L.H., Godtfredsen, A.K., Agerbæk, M.L., Nielsen, P.H., Keiding, K., 1996. Effects of colloidal stability on clarification and dewatering of activated sludge. Water Sci. Tech. 34, 449e457. Namer, J., Ganczarczyk, J.J., 1994. Fractal dimensions and shape factors of digested sludge particle aggregates. Water Poll. Res. J. Can. 29 (4), 441e455. Nielsen, P.H., 2002. Activated sludge: the floc. In: Bitton, G. (Ed.), Encyclopedia of Environmental Microbiology. Wiley, Chichester, England, pp. 54e61.

Nielsen, P.H., Jahn, A., 1999. Extraction of EPS. In: Wingender, J., Neu, T.R., Flemming, H.-C. (Eds.), Microbial Extracellular Polymeric Substances. Springer-Verlag, Berlin, pp. 49e72. Nielsen, P.H., Keiding, K., 1998. Disintegration of activated sludge flocs in the presence of sulfide. Water Res. 32 (2), 313e320. Nielsen, P.H., Thomsen, T.R., Nielsen, J.L., 2004. Bacterial composition of activated sludge e importance for floc and sludge properties. Water Sci. Technol. 49, 51e58. Nielsen, P.H., Mielczarek, A.T., Kragelund, C., Nielsen, J.L., Saunders, A.M., Kong, Y., Hansen, A.A., Vollertsen, J., 2010. A conceptual ecosystem model of microbial communities in enhanced biological phosphorus removal plants. Water Res. 44 (17), 5070e5088. Nielsen, P.H., Saunders, A.M., Hansen, A.A., Larsen, P., Nielsen, J.L., 2012. Microbial communities involved in enhanced biological phosphorus removal from wastewater e a model system in environmental biotechnology. Curr. Opin. Biotech. 23, 452e459. Niu, M., Zhang, W., Wang, D., Chen, Y., Chen, R., 2013. Correlation of physicochemical properties and sludge dewaterability under chemical conditioning using inorganic coagulants. Bioresour. Technol. 144, 337e343. Novak, J.T., 2006. Dewatering of sewage sludge. Dry. Technol. 24, 1257e1262. Novak, J.T., Goodman, G.L., Pariroo, A., Huang, J.-C., 1988. The blinding of sludges during filtration. J. Water Pollut. Control Fed. 60 (2), 206e214. Parker, D.G., Randall, C.W., King, P.H., 1972. Biological conditioning for improved sludge filterability. J. WPCF 44, 2066e2077. Peeters, B., Dewil, R., Lechat, D., Smets, I.Y., 2011. Quantification of the exchangeable calcium in activated sludge flocs and its implication to sludge settleability. Sep. Purif. Technol. 83, 1e8. Poon, C.S., Chu, C.W., 1999. The use of ferric chloride and anionic polymer in the chemically assisted primary sedimentation process. Chemosphere 39, 1573e1582. Rasmussen, H., Bruus, J.H., Keiding, K., Nielsen, P.H., 1994. Observations on dewaterability and physical, chemical and microbiological changes in anaerobically stored activatedsludge from a nutrient removal plant. Water Res. 28, 417e425. -Fauvel, E., Baudex, J.Raynaud, M., Vaxelaire, J., Oliver, J., Dieude C., 2012. Compression dewatering of municipal activated sludge: effects of salt and pH. Water Res. 46, 4448e4456. Sobech, D.C., Higging, M.J., 2002. Examination of three theories for mechanisms of cation-induced bioflocculation. Wat. Res. 36, 527e538. Shirato, M., Murase, T., Iwata, M., Nakatsuka, S., 1986. The Terzaghi-Voigt combined model for constant-pressure consolidation of filter cakes and homogeneous semi-solid materials. Chem. Eng. Sci. 41 (12), 3213e3218. Sørensen, B.L., Sørensen, P.B., 1997. Structure compression in cake filtration. J. Environ. Eng. ASCE 123, 345e353. Sørensen, P.B., Christensen, J.R., Bruus, J.H., 1995. Effect of small scale solids migration in filter cakes during filtration of wastewater solids suspensions. Water Environ. Res. 66, 25e32. Sørensen, P.B., Agerbaek, M.L., Sorensen, B.L., 1996. Predicting cake filtration using specific filtration flow rate. Wat. Environ. Res. 68, 1151e1155. randio, M., Masse , A., Espinosa-Bouchot, M.C., Cabassud, C., Spe 2005. Characterization of sludge structure and activity in submerged membrane bioreactor. Water Sci. Technol. 52, 401e408. Stickland, A.D., De Kretser, R.G., Scales, P.J., 2005. Nontraditional constant pressure filtration behavior. AIChE J. 51 (9), 2481e2488. Tiller, F.M., Yeh, C.S., 1987. The role of porosity in filtration II. Filtration followed by expression. AIChE J. 33, 1241e1256.

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019

w a t e r r e s e a r c h x x x ( 2 0 1 5 ) 1 e1 1

Tiravanti, G., Lore, F., Sonnante, G., 1985. Influence of the charge density of cationic polyelectrolytes on sludge conditioning. Water Res. 19, 93e97. Van den Broeck, R., Van Dierdonck, J., Caerts, B., Bisson, I., Kregersman, B., Nijskens, P., Dotremont, C., Van Impe, J.F., Smets, I.Y., 2010. The impact of deflocculation-reflocculation on fouling in membrane bioreactors. Sep. Sci. Technol. 71, 279e284. Van den Broeck, R., Van Dierdonck, J., Nijskens, P., Dotremont, C., Krzeminski, P., van der Graaf, J.H.J.M., van Lier, J.B., Van Impe, J.F., Smets, I.Y., 2012. The influence of solids retention time on activated sludge bioflocculation and membrane fouling in a membrane bioreactor (MBR). J. Membr. Sci. 401e402, 48e55. zac, P., 2004. Moisture distribution in activated Vaxelaire, J., Ce sludges: a review. Water Res. 38, 2215e2230. Vesilind, P.A., 1994. The role of water in sludge dewatering. Water Environ. Res. 66, 4e11. Wagner, J., Rosenwinkel, K.H., 2000. Sludge production in membrane bioreactors under different conditions. Water Sci. Technol. 41, 251e258.


n, B.-M., Keiding, K., Nielsen, P.H., 2000a. Anaerobic Wile deflocculation and aerobic reflocculation of activated sludge. Water Res. 34, 3933e3942. n, B.-M., Nielsen, J.L., Keiding, K., Nielsen, P.H., 2000b. Wile Influence of microbial activity on the stability of activated sludge flocs. Colloid Surf. B 18 (2), 145e156. n, B.-M., Lumley, D., Mattsson, A., Mino, T., 2008. Wile Relationship between floc composition and flocculation and settling properties studied at a full scale activated sludge plant. Water Res. 42, 4404e4418. Witzig, R., Manz, W., Rosenberger, S., Kru¨ger, U., Kraume, M., Szewzyk, U., 2002. Microbiological aspects of a bioreactor with submerged membranes for aerobic treatment of municipal wastewater. Water Res. 36, 394e402. Xu, W., Chellam, S., Clifford, D.A., 2004. Indirect evidence for deposit rearrangement during dead-end microfiltration of iron coagulated suspension. J. Membr. Sci. 239, 243e254. Yen, P.-S., Chen, L.C., Chien, C.Y., Wu, R.M., Lee, D.J., 2002. Network strength and dewaterability of flocculated activated sludge. Water Res. 36, 539e550.

Please cite this article in press as: Christensen, M.L., et al., Dewatering in biological wastewater treatment: A review, Water Research (2015), http://dx.doi.org/10.1016/j.watres.2015.04.019

Dewatering in biological wastewater treatment: A review.

Biological wastewater treatment removes organic materials, nitrogen, and phosphorus from wastewater using microbial biomass (activated sludge, biofilm...
1MB Sizes 0 Downloads 8 Views