2-1 Hazardous pollutants in sewage sludge
Sewage sludge is a complex mixture containing organic compounds, heavy metals, microorganisms and nutrients, respectively. Organic compounds such as biodegradable organics, recalcitrant organics, toxic organics and microorganisms (composed of beneficial microorganisms for final disposal and harmful pathogens) are generally found in sewage sludge. In order to remove the toxic pollutants from sewage sludge, many physical, chemical and biological treatments are carried out.
Several hazardous organic compounds such as bisphenol A (BPA), linear alkylbenzene sulfonates (LAS), nonylphenols (NPs), nonylphenol diethoxylates (NPDEs), nonylphenol ethoxylates (NPEOs), PAEs, polycyclic aromatic hydrocarbons (PAHs), polychlorinated biphenyls (PCBs) and dibenzodioxins/furans (PCDD/Fs) were observed in sewage sludge (Angelidaki et al., 2000; Fauser et al., 2003; Barnabe et al., 2007). Moreover, plasticizers, pharmaceuticals, personal care products, pesticide residues and flame retardants were also observed in sewage sludge. The gasoline additives had potential carcinogenic, teratogenic and endocrine disrupting properties (Barnabe et al., 2007). The release of recalcitrant and endocrine disrupting chemicals into the environment could cause a serious threat to the ecosystem (Barnabe et al., 2007). The releases of these toxic pollutants are due to the human activities, atmospheric deposition on the soil, urban runoff and industrial emissions. During wastewater treatment, the toxic pollutants could be accumulated in sewage sludge due to their non polar and hydrophobic nature that favored adsorption onto suspended solids. To decrease the harmfulness, toxic organic compounds must be controlled.
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In Denmark, the municipal sewage sludge were produced approximately 170,000 ton/dry-matter in 1994 and 140,000 ton/dry-matter in 2002. In the last decade, approximately 65% of sewage sludge was used for agriculture purpose (Jensen and Jepsen, 2005). European Union (EU) concerned about the available purpose for sewage sludge and reported the Directive 86/278 on environmental protection for agriculture. Therefore, EU and Denmark Environmental Protection Agency set the limit value of hazardous organics as shown in Table 2-1. Sewage sludge treatment with toxic organic compounds removal could protect the ecosystem, avoid public reluctances over beneficial use and favor commercialization of the final product.
Table 2-1 Limit value of hazardous compounds in sewage sludge (Spinosa, 2001;
Jensen and Jepsen, 2005)
Compound
Limit value (mg/kg-dw)
EU Denmark
DEHP 100 50
LAS 2,600 1,300
NPs and NPEOs 50 10
PAHs 6 3
PCBs 0.8* -
PCDD/Fs 100 -
*ng-TE/kg-dry matter
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2-2 Physical and chemical properties of PAEs
The structures and characteristics of DBP, DEHP and BBP are listed in Figure 2-1 and Table 2-2, respectively. In general, PAEs are liquids at room temperature. All of these PAEs have melting points below -25oC and boiling points at about 350oC. The high boiling point of these PAEs could prevent thermal decomposition in the ambient temperature. The low melting point and high boiling point of these PAEs contribute to their usefulness as plasticizers, heat transfer fluids and carriers. Water solubility is an extremely important property that influences the biodegradation, bioaccumulation potential and aquatic toxicity.
Water solubility is also a determining factor controlling the environmental distribution of chemicals. The more hydrophilic compounds with the shorter alkyl side-chains such as DBP are more soluble in water than those with the large alkyl-chains such as DEHP and BBP.
The equilibrium distribution of an organic chemical between water and octanol (Kow) is an important physical constant for predicting the tendency of a chemical to partition to water, sediment, sludge and soil. With increasing alkyl chain length, the log Kow increases indicating greater hydrophobicity. Most of the dialkyl phthalates are soluble in common organic solvents such as benzene, toluene, xylene, diethyl ether, chloroform and petroleum ether. Vapor pressure plays an important role in the fate of fugitive emissions and other releases of PAEs to the atmosphere. The vapor pressures of PAEs are declined with increasing alkyl chain length. Ideal plasticizers are highly compatible with polymers, stable in both high and low temperature environments, sufficiently lubricative over a wide temperature range, intensive to solar ultraviolet radiation, resistant for leaching and migration and inexpensive (Rahman and Brazel, 2004).
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Figure 2-1 PAEs structures (a) DBP, (b) DEHP and (c) BBP
Table 2-2 Physical and chemical properties of PAEs (Woodward, 1988; Staples et al., 1997)
Characteristics DBP DEHP BBP
CASno.a 84-74-2 117-81-7 85-68-7
Formula C16H22O4 C24H38O4 C19H20O4
Alkyl chain length 4 8 4, 6e
Specific gravity 1.047 0.986 1.116
Molecular weight 278 390 312
Vapor pressure (mmHg)d 2.7×10-5 1.0×10-7 5.0×10-6
aChemical abstracts service number
bAt atmosphere
cEquilibrium distribution of octanol/water partitioning
dAt 25oC
eAromatic ring
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2-3 Origins of PAEs into sewage sludge and its effect on sewage sludge treatment
PAEs enter the environment during production, manufacture, leaching, migration and volatilization by the usage and disposal of the products (Heise and Litz, 2004). Also, PAEs are released to wastewater and then transferred into wastewater treatment plant. When PAEs containing wastewater is introduced into a wastewater treatment plant: (1) one part of PAEs is degraded by physical, chemical and biological treatment during wastewater treatments and (2) the other part is strongly adsorbed on the surface of sludge (Marttinen et al., 2003;
Roslev et al., 2007; Dargnat et al., 2009). PAEs concentration of 2% was found in the treated water in which 70% was biodegraded and 28% was adsorbed in the sludge (Fauser et al., 2003). PAEs could be removed by chemical and biological methods; however, almost one-third of PAEs were still contained in the sewage sludge (Table 2-3).
Roslev et al. (2007) and Dargnat et al. (2009) proposed that the residual DEHP concentration was greater than DBP and BBP owing to its difficulty in biodegradation during sludge digestion. In addition, Cheng et al. (2000) analyzed the concentration of DEHP in three different wastewater treatment plants located in northern Taiwan and found that the concentration of DEHP in wastewater (including influent and effluent) and influent of sludge were lower than detection limits. However, after sludge treatment, the concentrations of DEHP were increased from 105.16 to 153.15 mg/kg-dw. These results indicate that the adsorbed DEHP in sludge would release in solution phase after sludge treatment.
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Table 2-3 Average PAEs concentration in wastewater treatment plant
Site Compounds
Wastewater treatment (g/L) Sludge treatment (mg/kg-dw)
Reference Influenta Effluent Influentb Dewatered
Min-Shen (Taiwan) DEHP NDc ND ND 142.86d
aWater containing municipal and industrial wastewater and returned sludge
bSludge containing primary and biological sludge before sludge treatment
cNot detected
dDewatered sludge after aerobic digestion
eDewatered sludge after anaerobic digestion
fDewatered sludge without pretreatment, digestion and adjusting
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2-4 Methods available for PAEs removal from sewage sludge
Many physical, chemical and biological methods are used for PAEs removal from sewage sludge in recent years. In physical and chemical catalogs, four methods such as hydrolysis, photodegradation, enzyme reaction and thermal reaction have been carried out to remove PAEs. PAEs could undergo two hydrolytic steps, producing first the mono-ester and one free alcohol moiety and a second hydrolytic step creating phthtalic acid and a second alcohol (Staple et al, 1997). But the degradation rate was very slow, especially in DEHP. The aqueous hydrolysis half-live of DEHP was 2,000 years while the aqueous hydrolysis half-lives of DBP and BBP were 22 and 0.3 years (Gledhill et al., 1980; Wolfe et al., 1980). Staple et al. (1997) also proposed that on aqueous photolysis occurred through absorption of UV light from sunlight in the region of 290-400 nm could be used to remove PAEs. Shorter wavelengths were attenuated by passage through the atmosphere and water column, so that the half-lives of photodegradation for PAEs removal were as much as shorter than hydrolysis (Lertsirisopon et al., 2009). Chen (2009) mentioned that the DEHP removal by the combination with UV light and hydrogen peroxide was better than that by direct UV catalysis. More acid or more alkaline PAEs containing aqueous solution could get better photodegradation than the neutral aqueous solution (Kaneco et al., 2006;
Lertsirisopon et al., 2009). Enzymatic treatment could effectively remove PAEs of sludge.
Gavala et al. (2004) contrasted the PAEs removal between 100 and 1,000 enzymic units/L of enzyme concentration, in which the PAEs degradation rates of 1,000 enzymic units/L of enzyme reaction was faster than of 100 enzymic units/L. For 1,000 enzymic units/L of enzyme reaction, PAEs could be degraded more than 50% after a week at 28oC. Although thermal treatment could remove PAEs but the degradation rate was very slow. Gavala et al.
(2004) demonstrated that PAEs were degraded less than 20% by thermal pretreatment at
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70oC for one week.
Since the PAEs were removed from sludge by biological treatments, several operation conditions affected the degradation rates. Shelton et al. (1984) and Wang et al. (1996) demonstrated that the lower molecular weight PAEs were easily biodegraded than the higher ones. In anaerobic digestion, removal of DBP was higher than DEHP owing to the long side-chains in DEHP (Alatriste-Mondragon et al., 2003). PAEs could be biodegraded in both aerobic and anaerobic conditions, but degradation rates were higher in aerobic than that in anaerobic condition (Angelidaki et al., 2000). Banat et al. (1999) proposed that the higher oxygen aeration rate was effective on the increase of DEHP removal; it was estimated that the DEHP could contact with more dissolved oxygen at the short time.
Alatriste-Mondragon et al. (2003) and Gavala et al. (2003) showed that DEHP accumulation in sludge coincided with a decrease in both gas production and in the efficiency of DBP removal. Chang et al. (2007) tested the single PAE aerobic biodegradation at different initial concentration. According to the first order kinetics of DBP biodegradation, the degradation rate constant (k) and half-lives (t1/2) of 100 mg/kg-dw DBP of sludge were 0.794 day-1 and 0.9 day, while the values of k and t1/2 of 1,000 mg/kg-dw DBP of sludge were 0.198 day-1 and 3.5 day, respectively. The results indicated that the higher initial concentration of DEHP led to the lower biodegradation rate. Chang et al. (2005) proposed that enough alkalinity could facilitate PAEs biodegradation even though alkalinity was increasing during biological reaction. Sludge might contain mesophilic bacteria and thermophilic bacteria.
Mesophilic bacteria could live at 35oC environment while thermophilic bacteria could adapt at 55oC environment. Chang et al. (2005) and Roslev et al. (2007) conducted PAEs biological removal at different temperature in the range of 30 and 50oC, in which the best temperature of PAEs removal was 50oC. In other words, thermophilic bacteria could degrade PAEs faster than mesophilic bacteria.
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2-5 Pretreatment for PAEs removal from sewage sludge
Many researchers pointed out that the goals of sludge pretreatment were to break the cell wall to facilitate the release of intracellular matter in the aqueous phase and break down many toxic and recalcitrant organic pollutants (Neis, 2002; Bien et al., 2004; Chang et al., 2007). Thus, sludge pretreatments were helpful for biological sludge digestion to get the better quality of sludge (Chiu et al., 1997; Lin et al., 1999; Kim et al., 2002; Gonze et al., 2003; Gronroos et al., 2005; Ding et al., 2006; Kim et al, 2009).
2-5-1 Alkalization pretreatment
The hydroxyl ions produced by alkalization could attack the cell walls of microorganisms and then release intracellular organics to liquid phase. Therefore, the types of alkaline reagents used in alkalization affected the efficiency of sludge pretreatment. Li et al.
(2008) used sodium hydroxide (NaOH) and calcium hydroxide [Ca(OH)2] to demonstrate the efficiency of SCOD increase in pretreatment and the results indicated that NaOH was more applicable for sludge pretreatment than Ca(OH)2. Kim et al. (2003) used NaOH, potassium hydroxide (KOH), magnesium hydroxide [Mg(OH)2] and Ca(OH)2 as alkalization reagents in sludge alkalization pretreatment. When adding the same concentration of these reagents to sludge individually, the order of high efficiency of SCOD increase was: NaOH > KOH >
Ca(OH)2 > Mg(OH)2. In Ca(OH)2 and Mg(OH)2 pretreatment, the disintegrated floc fragments and soluble organic polymers could be re-flocculated with the help of calcium and magnesium cations, so Ca(OH)2 and Mg(OH)2 were not applicable to conduct the sludge pretreatment.
In alkalization pretreatment, more NaOH concentration could get more SCOD increase
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of sludge because higher hydroxyl ions enhanced the reaction rate between hydroxyl ions and organics. In addition to hydroxyl ion concentration, total solids (TS) contents of sludge also affected the alkalization efficiency. Higher sludge TS concentration could provide higher organics in sludge alkalization. Hence, the more TS of sludge increased more SCOD after alkalization (Lin et al., 1998; Kim et al., 2009).
Table 2-4 SCOD increase in different alkalization method
Authors
aCombined with 20 kGy gamma-ray irradiation
bNot detected
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2-5-2 Ultrasound pretreatment
1. Physical mechanisms in sonication
During sonication, energy transportation is facilitated as electrical energy, acoustic energy, and chemical energy (Adewayi, 2001). When the ultrasound wave is propagated in the sludge, it generates a repeating pattern of compressions and rarefactions in sludge. As a result of reduced pressure, micro-bubbles are formed in the rarefaction regions. These micro-bubbles are known as cavitation bubbles containing vaporized liquid and gas that could be previously dissolved in the liquid phase. When the wave propagated, micro-bubbles oscillated under the influence of positive pressure and rapidly collapsed.
Cavitation was the phenomenon where micro-bubbles were formed in the aqueous phase and expand to unstable size, then rapidly collapsed. The collapsing of the bubbles resulted in localized temperature up to 5,000 K and pressures up to 180 MPa. The sudden and violent collapse of huge numbers of micro-bubbles generated powerful hydro-mechanical shear forces in the bulk liquid surrounding the bubbles. The collapsing bubbles disrupted adjacent microorganisms by extreme shear forces, rupturing the cell wall and membranes.
Hence, SCOD in sludge could increase during ultrasound reaction (Adewayi, 2001; Khanal et al., 2007).
2. Chemical mechanisms in sonication
Chemical mechanisms in sonication are listed below. When the ultrasound wave is propagated in the sludge, the heat could decompose H2O into hydrogen radicals (H·) and hydroxyl radicals (HO·). During quick cooling, H· and HO· could recombine to hydrogen peroxide (H2O2) and H2. At the same time, H· could react with dissolved oxygen to form
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HO2· and it could transform to H2O2. Hence, H·, HO·, HO2· and H2O2 could react with recalcitrant organics during sonication (Riesz et al., 1985; Suslick, 1989).
H2O + ultrasound wave → H· + HO· (1) increase (Bougrier et al., 2005; Dewill et al., 2006). Many researchers pretreated sludge by different sonication methods and discussed the effect of experimental condition on the increase of SCOD (Table 2-5). Thiem et al. (2001) demonstrated the effects of degree of disintegration for SCOD at different ultrasound frequency, where the lower ultrasound frequency got better degree of disintegration. In other words, lower ultrasound frequency increased more soluble organics of sludge. Chu et al. (2001) and Zhang et al. (2007) have investigated the ultrasound reaction at different power densities, which showed that the more ultrasound density the more SCOD increase. In addition, Chiu et al. (1997) combined alkalization and sonication pretreatments and demonstrated that the more NaOH concentration to sludge or the more sludge pH, the more SCOD increase. It was explained that the hydroxyl ions addition could attack and weaken the bacterial cell-wall then facilitate better destruction by following ultrasound reaction (Wang et al., 2005). On the other hand, Wang et al. (2005) showed that sodium bicarbonate (NaHCO3) addition could mask the oxidizing effect of OH· to disturb ultrasound reaction. In brief, addition of hydroxyl ion instead of HCO3- could facilitate SCOD increase in pretreatments. Under higher TS contents, the violent collapsing of micro-bubbles could accelerate the particles in vicinity of the bubbles, which could bombard the adjacent particles. Particles at a higher TS
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contents could facilitate the sludge disruption due to particle-to-particle collision. Wang et al. (2005), Dewil et al. (2006) and Show et al. (2007) reported higher SCOD increase under higher TS of sludge.
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Table 2-5 SCOD changes after sonication Authors Frequency
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Table 2-5 SCOD changes after sonication (Continue) Authors Frequency after 1M NaOH treated for 24 hours before sonication, and SCOD0 was original SCOD
bNot detected
c1M NaOH treated before sonication
dCombination with NaOH alkalization and sonication
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