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Effect of organic loading rate on the stability, operational parameters and performance of a secondary upflow anaerobic sludge bed reactor treating piggery waste
Bioresource Technology 96 (2005) 335–344
Effect of organic loading rate on the stability, operational parameters and performance of a secondary upflow anaerobic sludge bed reactor treating piggery waste E. S anchez a, R. Borja b
a,*
, L. Travieso a, A. Martın b, M.F. Colmenarejo
c
a Instituto de la Grasa (C.S.I.C.), Avda. Padre, Garcıa Tejero 4, E-41012 Sevilla, Spain Departamento de Ingenierıa Quımica, Facultad de Ciencias, Campus Universitario de Rabanales, Ctra. Madrid-Cadiz, Km. 396, E-14071 Cordoba, Spain c Centro de Ciencias Medioambientales (C.S.I.C.), Serrano 115-duplicado, 28006 Madrid, Spain
Received in revised form 25 March 2004; accepted 2 April 2004 Available online 14 May 2004
Abstract A study of anaerobic digestion of piggery wastewater was carried out in a laboratory-scale sludge bed reactor as a secondary treatment. The effect of organic volumetric loading rates (BV ) in the range of 1.0–8.1 g TCOD/l d on the process performance was evaluated. The best results were obtained at BV equal to or lower than 4 g TCOD/l d. At higher BV values, the removal efficiency of the process decreased suddenly. A linear relationship was found between the effluent SCOD and the TVFA/alkalinity ratio (P ). A relationship was found among the different operational variables (BV , removal efficiency, effluent soluble COD, soluble COD removal rate (R), retention factor (/), specific microbial growth rate (l), methane production rate per volume of reactor and per volume of waste treated––QM and qM , respectively) and the corresponding regression equations were obtained. An increase of BV determined a decrease of removal efficiency, / and qM and an increase of effluent soluble COD, l, R and QM . The value of the maximum specific microbial growth rate (lM ) determined through the equation that correlated BV and l was found to be 0.19 d1 . This value was of the same magnitude as those reported in other works of anaerobic digestion of piggery waste. Ó 2004 Elsevier Ltd. All rights reserved. Keywords: Piggery waste; Upflow anaerobic sludge bed reactor; Organic volumetric loading rate (BV ); Operational parameters; Correlation
1. Introduction Piggery waste is characterized by a high content of organic matter and pathogenic organisms. The disposal of piggery waste without adequate treatment can cause a drastic effect on the environment and human health. This waste is formed by a mixture of manure (feces and urine) and food wastage such as swill and sugar cane molasses (S anchez et al., 2001). According to the most common characteristics of this waste, anaerobic digestion could be considered one of the most promising treatment alternatives (Kimchie et al., 1988; Hobson and Shaw, 1973; Hobson, 1981, 1985, 1992; S anchez
* Corresponding author. Tel.: +34-95-4689654; fax: +34-954691262. E-mail address: [email protected] (R. Borja).
0960-8524/$ - see front matter Ó 2004 Elsevier Ltd. All rights reserved. doi:10.1016/j.biortech.2004.04.003
et al., 1995; Baader, 1990; Yang and Gan, 1998; Parkin and Owen, 1986; Sanchez and Travieso, 1994). Piggery waste treatment by anaerobic digestion has several advantages, the most important are as follows: 1. The process destroys pathogenic and parasitic organisms. 2. Methane as a valuable by-product can be used as a source of energy, and liquid and sludge effluents can be used as a soil conditioner. 3. The low biomass production determines smaller volumes of sludge than in the aerobic processes, thus lowering the cost of sludge management. 4. Capacity to stabilize large volumes of diluted organic slurries at low costs. 5. This process is more economical than the aerobic process, especially at strengths higher than 4 g/l of COD.
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E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
Hobson and Shaw (1973, 1974) studied the mechanism and flora of the anaerobic digestion of piggery waste in a complete mixed digester. They observed that Enterobacteria streptococci made up about 50% of the total content of digester organisms. The predominant groups of organisms were identified as Clostridium butyricum, Bacteroides and gram negative Coccobacilli, and the strict anaerobic ones did not predominate over the facultative ones. Stevens and Schulte (1979) studied the effect of the temperature at solids retention times of between 6 and 55 days at organic volumetric loading rates of between 0.61 and 4.81 kg VS/m3 d, in a complete mixed anaerobic digester. They concluded that at organic rates in the range of 0.61–1.80 kg VS/m3 d and temperatures lower than 25 °C, the operation proceeded satisfactorily. Low temperature digestion was found to require twice as long a retention time with satisfactory production and composition of gas. A previous study (S anchez et al., 2001) looked at the influence of temperature and substrate concentration on the anaerobic batch digestion of piggery wastewater. This study compared the process at mesophilic temperature (35 °C) with temperatures in the range of 16.8–29.5 °C, and influent concentrations in the range of 3.3–26.3 g TCOD/l. The process at mesophilic temperature was more stable than at ambient temperature, obtaining higher values of removal efficiency. Batch or intermittent feeding, plugflow, completely-mixed fixed bed and upflow bed anaerobic digesters have been used for piggery waste treatment at laboratory, pilot and full-scale plants with good results (Baader, 1990; Chen, 1983; Hobson, 1992; I~ niguez et al., 1994; Kimchie et al., 1988; Lapp et al., 1995; Lomas, 1999; Montalvo, 1995; Rodrıguez and Lomas, 1999; S anchez and Travieso, 1994; S anchez et al., 1995, 2001, 2002; Yang and Gan, 1998). The effect of different parameters such as temperature, hydraulic retention time, organic volumetric loading rate, mixing and other factors have been previously studied. However, the upflow anaerobic sludge bed (UASB) reactor has seldom been used for piggery waste treatment because the composition of this waste, rich in nitrogenous compounds, makes the active biomass densification, the granulation process, and therefore the microorganisms retention in the reactor difficult. UASB reactors have
been applied mainly to carbohydrate type wastewaters with success and limited work has been done on the application of this reactor configuration for piggery waste stabilization. Moreover, pigs in Cuba are fed with a mixture of swill and molasses (Sanchez et al., 2001). This could determine different behaviour and performance patterns in Anaerobic Sludge Bed Reactors. Montalvo (1995) studied the separate anaerobic digestion of liquid and solid fractions of piggery waste, respectively, and compared this system to the digestion of the whole waste in a single sludge bed reactor. He found that separate digestion was more effective than that carried out using the single stage reactor. S anchez et al. (1995) compared the performance of sludge bed and anaerobic fixed bed reactors treating piggery wastewaters. The best results were obtained in the latter. The aim of the present work was to evaluate the behaviour and performance of a UASB reactor treating piggery wastewater (rich in carbohydrate) after a preliminary screening and primary sedimentation. A study of the variations of the operational parameters with organic volumetric loading rates and their interrelations by means of different equations was also carried out.
2. Methods 2.1. Piggery waste The piggery waste used in the experiment was obtained from a farm situated 20 km west of Havana City, Cuba. The waste was screened through a 2 mm sieve. After screening, the waste settled for 1 h. The average characteristics and range of the supernatants obtained in primary sedimentation are presented in Table 1. The supernatant after settling was diluted to a TCOD around 8 g/l, value within the range of primary sedimentation supernatant effluents, and this final settled effluent was used to feed the reactor. 2.2. Equipment The experiments were carried out in a laboratoryscale upflow anaerobic sludge bed (UASB) reactor,
Table 1 Characteristics of the raw supernatant piggery waste after a settling period of 60 min Parameter
Average value
Standard deviation
Number of samples
Total COD (mg/l) Total organic carbon (TOC) (mg/l) Total solids (TS) (mg/l) Total volatile solids (TVS) (mg/l) Total suspended solids (TSS) (mg/l) Volatile suspended solids (VSS) (mg/l) Total nitrogen (TN) (mg/l) Phosphates (P) (mg/l) pH
10,189 4100 7210 5122 1637 1166 341 419 6.0
4911 1748 2430 2178 886 813 135 228 0.7
44 44 70 70 70 70 41 41 60
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
consisting essentially of a 5 l cylindrical glass flask 15 cm in diameter and 30 cm tall. The flask was hermetically sealed at the top by a rubber cap with two holes 5 mm in diameter required for biogas outlet and effluent collection, respectively. The biogas was collected in a gas holder submerged in a saturated solution of water and sodium chloride to prevent the absorption of CO2 . The lower part of the reactor was fed by means of a pipe 10 mm in diameter connected to the centre of the reactor allowing the influent to be introduced by using a peristaltic pump. 2.3. Inoculum The upflow anaerobic sludge reactor (UASB) was inoculated with 1 l of anaerobic sludge obtained from a batch piggery sludge digester after 40 days of digestion time. The characteristics of the inoculum are summarized in Table 2.
was analysed by an Orsat analyser to determine the concentration of CO2 . The temperature was maintained at a range of 30–35 °C during the experimental period. 2.5. Analyses Samples of effluent were taken and analysed three times a week and influent once a week. Hence, each run had a duration of seven weeks at steady-state conditions. Analytical determinations of total and soluble chemical oxygen demand (TCOD and SCOD, respectively), total and soluble organic carbon (TOC and SOC, respectively), total suspended solids (TSS), volatile suspended solids (VSS), total volatile fatty acids (TVFA), alkalinity and pH were carried out according to Standard Methods for the Examination of Water and Wastewater (APHA, 1989).
3. Results and discussion
2.4. Experimental procedure Once the inoculum was added to the UASB reactor, the operational volume was completed with distilled water to obtain 5 l of working volume. After this step, the influent was continuously fed by a peristaltic pump at a flow-rate of 0.625 l/d (hydraulic retention time–– HRT––of eight days) until steady-state conditions were obtained. Steady-state was considered after a period equivalent to three times the HRT (24 days). Continuous-mode experiments were carried out at influent flowrates of: 0.625, 0.833, 1.00, 1.25, 1.66, 2.50 and 5.00 l/d, which corresponded to HRT values of 8, 6, 5, 4, 3, 2 and 1 days, respectively, and organic volumetric loading rates (BV ) of 1.0, 1.4, 1.6, 2.0, 2.7, 4.1 and 8.1 g TCOD/ l d, respectively. The experiment was performed by increasing the influent flow-rate and consequently, reducing the HRT at a constant influent concentration. A total of 20 samples were taken for each operational condition and average values were determined for each parameter analysed. Biogas production was determined daily. The gas accumulated in the measurement flask Table 2 Characteristics of the inoculum useda Parameter
Value
TCOD (g/l) TS (g/l) TSS VS (g/l) VSS (g/l) TOC pH
47.5 50.0 41.0 26.1 21.4 15.7 7.3
a
337
Values are averages of five determinations. There was virtually no variation (<3%) among these analyses.
Table 3 shows the steady-state results obtained under the different experimental conditions studied. The increase of BV at fixed value of influent TCOD caused a decrease in the removal efficiency expressed in TCOD, TOC, SCOD, SOC, TSS and VSS. A sudden decrease in removal efficiencies was observed at a BV of 4.1 g TCOD/l d (HRT of two days). The reduction of removal efficiencies was a consequence of the increase in total volatile fatty acids (TVFA) and a simultaneous decrease in the alkalinity causing an increase in the TVFA/alkalinity ratio (P ) and a reduction in the pH value. This situation had a strong influence on the biogas quality, increasing the CO2 concentration on all the range of BV studied. Fig. 1 shows the strong effect of the HRT on the TVFA/alkalinity ratio and, therefore, on the process stability. An increase in the HRT value caused a decrease in P , favouring the process stability. At hydraulic retention times in the range of 4–8 days the process was well balanced and removal efficiencies higher than 70% were obtained in most cases. However, when the HRT was lower than or equal to two days, the removal efficiencies suddenly decreased and values of P increased over 0.5, producing unbalanced conditions. Fig. 2 illustrates the variation of the CO2 concentration as a function of the values of P . The concentration of CO2 increased when the value of P increased from 0.2 to 0.6 and was practically constant at higher P values. Similar behaviour was observed in the acidogenesis of dairy and gelatin-rich wastewaters using upflow anaerobic reactors (Yu and Fang, 2002, 2003). A decrease in the HRT produced an increase in the effluent concentration because the anaerobic waste stabilization was less complete and TVFA concentration and P increased. Fig. 3 illustrates the relationship between the soluble
338
HRT (d)
8
6
5
4
3
2
1
BV (g TCOD/l d)
1
1.4
1.6
2
2.7
4.1
8.1
Parameter
Influent
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
TCOD SCOD TOC SOC TSS VSS TVFA Alkalinity Pb pH Q (l/l w d) CH4 CO2 qM (l/l w d) QM (l/l r d)
8117 5782 3272 2338 1678 1309 1855 1397 1.1 5.9
1205 785 513 378 306 223
85.4 86.0 84.3 83.8 84.1 83.4
1726 1444 681 590 368 268
78.4 73.5 79.2 74.8 77.3 75.6
2030 1583 706 639 540 353
75.0 72.6 78.4 72.7 67.8 73.0
2141 1772 815 696 692 390
73.6 69.4 75.1 70.2 61.3 70.8
3512 2412 1030 792 804 469
56.7 58.3 68.5 66.1 52.1 64.2
4941 3187 1686 1152 990 548
39.1 44.9 48.5 50.7 41.0 58.1
6611 4644 1755 1410 1223 752
18.6 19.7 34.2 39.7 27.1 42.6
a b
573 1986 0.24 7.4 5.5 66 32.2 3.63 0.45
595 1720 0.29 7.3 4.5 63.3 34.2 2.85 0.47
750 1838 0.34 7.3 4.1 57.8 40.1 2.37 0.47
810 1660 0.41 7.3 3.5 54.7 42.0 1.91 0.48
844 1435 0.49 7.2 3.3 54.2 43.0 1.79 0.60
858 1386 0.52 7.2 2.9 46.6 61.3 1.35 0.68
1211 1030 0.98 6.4 2.6 33.3 63.0 0.86 0.86
Values are averages of 20 determinations taken over three weeks after the steady-state conditions had been reached. The differences between the observed values were less than 3% in all cases. TVFA/alkalinity ratio is expressed in equivalents of acetic acid/equivalents of calcium carbonate.
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
Table 3 Steady-state results obtained under different experimental conditionsa
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344 1.2
100
339
1.2
E % TCOD
90
E (% TCOD removal)
70 0.8 60
0.6
50
40 0.4 30
20 0.2
1
P (eq. acetic acid/eq. calcium carbonate)
1 80
P (eq. acetic acid/eq. calcium carbonate)
P
0.8
0.6
0.4
0.2
10
0 0
2
4
6
0 10
8
0 0
Fig. 1. Variation of the percentage of TCOD removal and TVFA/ alkalinity ratio (P ) with the HRT.
Volumetric fraction of CO2 in the biogas
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0 0
0.2
0.4
0.6
0.8
1
1.2
P (eq. acetic acid/eq. calcium carbonate) Fig. 2. Variation of the volumetric fraction of CO2 in the biogas with the TVFA/alkalinity (P ) ratio.
substrate concentration (SCOD) and the value of P . A straight line was obtained with the following equation: P ¼ 0:182ðSÞ þ 0:056
ð1Þ 2
1
2
3
4
5
6
7
S (g SCOD/l)
HRT (d)
the regression coefficient, R , being equal to 0.97.
Fig. 3. Effect of the effluent SCOD concentration on the TVFA/ alkalinity (P ) ratio.
Eq. (1) showed a strong influence of substrate concentration on the TVFA/alkalinity ratio and, therefore, on process stability. As we already know, TVFA/alkalinity ratio (P ) can be used as a measure of process stability (Fannin, 1987): when this ratio is less than 0.3– 0.4, the process is considered to be operating favourably without the risk of acidification. As was observed in this study, unstable conditions (P > 0:5) appeared in substrate concentrations higher than 2.3 g SCOD/l. The intercept value can be attributed to the presence of mineral acids that have a slight influence and contribution on the P value. As regards the alkalinity values obtained in the effluents of the process, it was observed (Table 3) that the buffering capacity of the experimental system was maintained at favourable levels with appropriate total alkalinity present at BV values in the range of 1.0–4.1 g TCOD/l d. The experimental data obtained in this study indicate that a total alkalinity in the range of 1390–1990 mg/l as CaCO3 was sufficient to prevent the pH from dropping to below 7.2 at BV values of up to 4.1 g TCOD/l d. The pH in the reactor remained more or less constant at HRTs in the range of 8.0–2.0 days, with 7.4 and 7.2 as extreme values. This stability can be attributed to carbonate/bicarbonate buffering. This is produced by the generation of CO2 in the digestion process, which is not completely removed from the reactor as gas (Wheatley, 1990).
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E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
the uptakes of soluble organic carbonaceous compounds were always higher than those obtained for particulate organic carbonaceous compounds. From the results obtained, it can also be demonstrated that suspended COD was previously transformed to soluble COD and then transformed to methane. Fig. 5 shows the variation of the experimental effluent substrate concentrations, S, expressed as g SCOD/l, with the organic loading rate BV . As can be seen, a hyperbolic relationship appears to relate both parameters, which can be described by the following equation:
1 TCOD
0.9
SCOD TOC SOC
0.7
0.6
0.5
S ¼ K1 ½BV =ðK2 þ BV Þ 0.4
0.3
0.2
0.1
0 0
2
4
6
8
10
Bv (g TCOD/l d)
Fig. 4. Variation of the fractional removal efficiency with the organic volumetric loading rate (BV ).
The effect of BV on the removal efficiencies, expressed in TCOD, SCOD, TOC and SOC is given in Fig. 4. As can be seen, an increase in the BV determined a progressive decrease in the removal efficiency, the best removal efficiencies were obtained at BV values in the range of 1.0–4.1 TCOD/l d. At BV of 8.1 g TCOD/l d the removal efficiencies dropped suddenly, and this was concomitant with an increase in P ratio, which achieved values over 0.5. An exponential relationship between BV and the removal efficiencies was observed. The typical equation that correlated both variables is given as follows: E¼e
kBV
ð2Þ
where E, is the fractional removal efficiency and k is a constant rate given in d l/g, its value depending on parameters selected for expressing the substrate concentration (TCOD, SCOD, TOC and SOC). Therefore, the k values obtained for TCOD, SCOD, TOC and SOC were found to be 0.21, 0.20, 0.14 and 0.13 d l/g, respectively. The correlation coefficients were 0.99, 0.99, 0.96 and 0.94 for TCOD, SCOD, TOC and SOC, respectively, and the variability coefficients for the corresponding constants were: 5%, 5%, 11% and 13%, respectively. The values of the constant obtained demonstrated that the rates of diminution of removal efficiencies were decreased with BV in the following order: SOC > TOC > SCOD > TCOD. The results showed that
ð3Þ
where K1 is given in g SCOD/l and represents the theoretical maximum SCOD concentration at the effluent when BV is too high in comparison to K2 . The constant K2 is expressed in g TCOD/l d and represents the BV value for which the effluent SCOD concentration is equivalent to half the value of K1 . A plot of the inverse of S versus the inverse of BV gave a straight line with an intercept equal to 1=K1 and slope equal to K2 =K1 with a correlation factor (R2 ) equal to 0.99. So the values of K1 and K2 were found to be 8.7 ± 0.4 g SCOD/l and 7.3 ± 0.3 g TCOD/l d, respectively. Fig. 5 also shows the theoretical S values, obtained by Eq. (3), and its variation with the BV values. As can be seen, deviations lower
6
5
4
S (g SCOD/l)
Fractional removal efficiency
0.8
3
2
1
0 0
2
4
6
8
10
Bv (g TCOD/l d)
Fig. 5. Effect of the organic volumetric loading rate (BV ) on the experimental effluent SCOD values and those theoretical ones predicted by Eq. (3).
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
ð4Þ
where RM , is the maximum SCOD removal rate (g SCOD/l d), and K is a kinetic constant that represents the value of BV for which the value of R is equal to 1=2RM . The value of the kinetic constants was calculated by means of the plot of the inverse of BV versus the inverse of R. A straight line with a correlation coefficient of 0.97 was obtained. From this linearised equation, RM can be calculated from the intercept on the y-axis and K can be calculated from the slope of the straight line. From this regression, the values of the constants RM and K were estimated to be 2.1 ± 0.2 g SCOD/l d and 2.4 ± 0.3 g TCOD/l d, respectively. Substitution of these values into Eq. (4) allowed the theoretical R values to be determined. Differences between experimental and theoretical values were lower than 10% in all cases. The biogas production and its composition were also a function of the BV used. Fig. 7 shows the effect of BV
1.8
1.6
1.4
R (g SCOD/l d)
1.2
4
1
3.5 0.8 3
2.5
0.6
2 0.4
1.5
1 0.2 0.5
0 0
2
4
6
8
0 10
Bv (g TCOD/l d)
Fig. 7. Effect of the organic volumetric loading rate (BV ) on methane gas production per litre of influent (qM ) and per litre of reactor (QM ).
on the methane production rate (qM ), expressed in l/l waste feed. As can be seen, an increase in BV produced an increase in methane gas production per volume of the reactor (QM ) due to the increase in the total amount of organic matter provided to the reactor at increased BV values. However, a reduction of the biogas amount produced per volume of influent added (qM ) was observed in a parallel way when BV increased. The organic volumetric loading rate (BV ) has a great influence on the anaerobic digestion performance and is an important parameter to the process scale-up. BV is related to the HRT and the influent concentration by the following equation: BV ¼ S0 Q=V ¼ S0 =HRT
1
QM (l/l r.d)
QM (l/l r.d)
R ¼ RM ½BV =ðK þ BV Þ
qm (l/l w.d)
qm (l/l w.d)
than 5% between the experimental and theoretical values of S were always observed. For low BV values the second term of the denominator of this expression can be neglected and Eq. (3) can be transformed into a linear equation, as follows: S ¼ 1:20BV . Fig. 6 illustrates the effect of BV on the SCOD removal rate, R. As can be seen, the SCOD removal rate increased with the BV according to a hyperbolic model, as follows:
341
ð5Þ
where S0 , is the influent concentration (g TCOD/l); Q, is the influent flow-rate (l/d); V , is the volume of the reactor (l); and HRT, is the hydraulic retention time (days), which is equal to the quotient V =Q. At a fixed influent concentration, an increase in BV could be only obtained by an increase in the influent flow-rate (Q) which causes a reduction in HRT. The HRT is related to the microorganisms or solids retention time (SRT) as follows:
0.8
0.6
0.4
0.2
SRT ¼ ½X0 ðHRTÞ=XE
0 0
1
2
3
4
5
6
7
8
9
Bv (g TCOD/l d)
Fig. 6. Variation of the SCOD removal rate (R) with the organic volumetric loading rate (BV ).
ð6Þ
where X0 , is the concentration of microorganisms in the reactor, given in g VSS/l; XE , is the concentration of microorganisms in the effluent of the reactor, given in g VSS/l. The ðX0 =XE Þ ratio is defined as the biomass
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
retention factor ð/Þ (McCarty and Mossey, 1991). The increase in / determines the increase in the SRT and a better adaptation of the microorganisms to the substrate favouring methanogenesis and improving process performance. The increase in the SRT in the anaerobic process can be obtained by increasing the HRT or the / value. Although the increase in the HRT can be achieved by using greater reactor volume or reducing the influent flow, the increase in / using alternatives to prevent the biomass to escape from the process such as sludge recycling or biomass immobilization are the best options. Therefore, HRT can also be expressed as HRT ¼ SRT=/
25
20
Retention factor (φ)
342
15
10
ð7Þ
from which the following equation can be obtained: 5
1=HRT ¼ /=SRT ¼ l/
ð8Þ
taking into account that l, defined as the microbial specific growth rate (d1 ), is numerically equal to the inverse of SRT. Therefore, by combining Eqs. (5) and (8), the following equation was obtained: l ¼ BV =S0 /
ð9Þ
Eq. (9) demonstrates that the microbial specific growth rate and the biomass factor retention are inversely proportional. Therefore, to achieve good operational conditions, the retention factor must be increased, improving the microorganisms separation and reducing the concentration of VSS in the effluent. The maximum value of l will be obtained for a maximum value of BV and a minimum value of /. Fig. 8 shows the effect of BV on the values of /. According to Table 2, the value of X0 was 4.3 g VSS/l, the value of XE being equal to the VSS concentration in the effluent. The concentration of microorganisms in the influent was assumed to be neglected. The value of / decreased for increased BV values, due to the increase in influent flow-rate which in turn caused an increase in the upflow velocity in the reactor. This fact provoked a higher microorganisms concentration in suspension, reducing the SRT. Moreover, an increase in the amount of organic matter added to the reactor produced the development of non-methanogenic microorganisms. As can be seen in Fig. 8, an exponential relationship can be established between BV and /. Eq. (10) demonstrates a strong influence of the organic volumetric loading rate on the retention factor: / ¼ 17:63BV0:57
ð10Þ
the regression coefficient, R2 , being 0.98. Combining Eqs. (9) and (10), a relationship was found that correlates l with BV for the range of the organic loading rates studied:
0 0
1
2
3
4
5
6
7
8
9
Bv (g TCOD/l d)
Fig. 8. Effect of the organic volumetric loading rate (BV ) on the retention factor (/).
l ¼ 7 103 B1:57 V
ð11Þ
This equation allows the theoretical l values to be calculated. The value of lM for the experimental range studied was obtained using Eq. (11). This value was 0.18 d1 and is very close to that obtained by Rodrıguez and Lomas (1999). Hence, the minimum theoretical values of /, SRT and HRT obtained were: 5.4, 5.4 and 1.0 days, respectively. Fig. 9 illustrates the effect of the parameter / on the methane production per volume of influent added to the process (qM ). As can be seen, a straight line was obtained, which is given by the following equation: qM ¼ 0:196/ 0:163
ð12Þ
the regression coefficient, R2 , being 0.98. By combining Eqs. (10) and (12), the following equation is obtained: qM ¼ 3:455BV0:57 0:163
ð13Þ
Finally, Fig. 10 shows the effect of l on the substrate removal rate given in SCOD. As can be clearly observed, an increase in l produced an increase in the substrate removal rate. The curve follows a Michaelis– Menten equation type, given by the following expression: R ¼ 2:99½l=ð0:016 þ lÞ
ð14Þ
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
Acknowledgements
4
The authors wish to express their gratitude to the ‘‘Ministerio de Ciencia y Tecnologıa’’ (project REN 2001-0472/TECNO), and ‘‘Ministerio de Educaci on, Cultura y Deportes’’ for providing financial support.
3.5
3
References
qm (l/l w.d)
2.5
2
1.5
1
0.5
0 0
5
10
15
20
25
Retention factor (φ)
Fig. 9. Effect of the retention factor (/) on the methane production per litre of waste influent added (qM ).
1.6
1.4
1.2
R (g SCOD/l d)
343
1
0.8
0.6
0.4
0.2
0 0
0.02 0.04 0.06 0.08
0.1
0.12 0.14 0.16 0.18
0.2
µ (1/d) Fig. 10. Effect of the specific microbial growth rate (l) on the SCOD removal rate (R).
the regression coefficient, R2 , and standard error being equal to 0.97 and 10%, respectively.
APHA (American Public Health Association), 1989. Standard Methods for the Examination of Water and Wastewater, 17th ed. APHA, Washington, DC, USA. Baader, W., 1990. Biogas technology and implementation in the Federal Republic of Germany. In: Proceedings of the International Conference on Biogas, Pune, India, January. Chen, Y.R., 1983. Kinetic analysis of anaerobic digestion of pig manure and its design implications. Agric. Wastes 8, 65–81. Fannin, K.F., 1987. Start-up, operation, stability and control. In: Chynoweth, D.P., Isaacson, R. (Eds.), Anaerobic Digestion of Biomass. Elsevier, London, UK, pp. 171–195. Hobson, P.N., Shaw, B.G., 1973. The anaerobic digestion of wastes from an intensive pig unit. Water Res. 7, 437–449. Hobson, P.N., Shaw, B.G., 1974. The bacterial population of piggery waste anaerobic digesters. Water Res. 8, 507–516. Hobson, P.N., 1981. Anaerobic digestion of animal excreta and other agricultural wastes. Trib. Cebedeau 34 (455), 437–441. Hobson, P.N., 1992. Treatment of animal wastes and uses of treated residues. Biotechnologies for pollution control and energy. In: Proceedings of the Third Workshop of the Working Group on Biogas Production Technologies. CNREE Network on Biomass Production and Conversion for Energy, Braunschweig, Germany, pp. 48–70. I~ niguez, C.G., Robles, C.A., Franco, G.M., 1994. Continuous solidsubstrate fermentation of swine waste recovered solids for pig feed. Biores. Technol. 50, 139–147. Kimchie, S., Lumbroso, R.E., Geller, Z., Abramovitch, D., Shelef, G., 1988. An integrative treatment process for piggery wastes based on anaerobic digestion. In: Tilche, A., Rozzi, A. (Eds.), Proceeding of the Fifth International Symposium on Anaerobic Digestion. Monduzzi Editore S.P.A., Bologna, Italy, pp. 635–639. Lapp, H.M., Schulte, D., Kroeker, E.J., Sparling, A.B., Topnik, B.H., 1995. Start-up of pilot-scale swine manure digesters for methane production. Managing Livestock Wastes 275, 234–243. Lomas, J.M., 1999. Evaluation of a pilot scale downflow stationary fixed film anaerobic reactor treating piggery slurry in the mesophilic range. Biomass Bioenergy 17 (1), 49–58. McCarty, P.L., Mossey, F.E., 1991. Modelling of anaerobic digestion process (a discussion concept). Water Sci. Technol. 24, 17–33. Montalvo, S.J., 1995. Treatment of swine wastes by a high-ratemodified-anaerobic-process (HRAMP). Biores. Technol. 53, 207– 210. Parkin, G., Owen, W., 1986. Fundamentals of anaerobic digestion of wastewater sludges. J. Env. Eng. 112 (5), 867–1120. Rodrıguez, A., Lomas, J.M., 1999. Kinetic study of the anaerobic digestion of the solid fraction of piggery slurries. Biomass Bioenergy 17 (5), 435–443. Sanchez, E., Travieso, L., 1994. Effect of the organic volumetric loading in SCOD removal in downflow fixed bed reactors. Biores. Technol. 47, 173–176. Sanchez, E., Monroy, O., Ca~ nizares, R.O., Travieso, L., 1995. Comparative study of piggery waste treatment by upflow sludge beds anaerobic reactors and packed bed reactors. J. Agric. Eng. Res. 62, 71–76.
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S anchez, E., Borja, R., Weiland, P., Travieso, L., 2001. Effect of substrate concentration and temperature on the anaerobic digestion of piggery waste in tropical climates. Process Biochem. 37, 483–489. S anchez, E., Travieso, L., Weiland, P., Borja, R., Nikolaeva, S., 2002. Effect of influent strength changes on the performance of a downflow anaerobic fixed bed reactor treating piggery waste. Resour., Conserv. Recycl. 36, 73–82. Stevens, M.A., Schulte, D.D., 1979. Low temperature digestion of swine manure. Journal of Environmental Engineering Division ASCE 105 (EE1), 33–42.
Wheatley, A., 1990. Anaerobic digestion: a waste treatment technology. SCI, Elsevier, London, UK. Yu, H.Q., Fang, H.H.P., 2002. Acidogenesis of dairy wastewaters at various pH levels. Water Sci. Technol. 43 (10), 201–206. Yu, H.Q., Fang, H.H.P., 2003. Acidogenesis of gelatin-rich wastewaters in an up-flow anaerobic reactor: influence of pH and temperature. Water Res. 37, 55–66, Water Sci. Technol. 43 (10) 201–206. Yang, P.Y., Gan, C., 1998. An on-farm swine waste treatment system in Hawaii. Biores. Technol. 65, 21–27.
Effect of organic loading rate on the stability, operational parameters and performance of a secondary upflow anaerobic sludge bed reactor treating piggery waste E. S anchez a, R. Borja b
a,*
, L. Travieso a, A. Martın b, M.F. Colmenarejo
c
a Instituto de la Grasa (C.S.I.C.), Avda. Padre, Garcıa Tejero 4, E-41012 Sevilla, Spain Departamento de Ingenierıa Quımica, Facultad de Ciencias, Campus Universitario de Rabanales, Ctra. Madrid-Cadiz, Km. 396, E-14071 Cordoba, Spain c Centro de Ciencias Medioambientales (C.S.I.C.), Serrano 115-duplicado, 28006 Madrid, Spain
Received in revised form 25 March 2004; accepted 2 April 2004 Available online 14 May 2004
Abstract A study of anaerobic digestion of piggery wastewater was carried out in a laboratory-scale sludge bed reactor as a secondary treatment. The effect of organic volumetric loading rates (BV ) in the range of 1.0–8.1 g TCOD/l d on the process performance was evaluated. The best results were obtained at BV equal to or lower than 4 g TCOD/l d. At higher BV values, the removal efficiency of the process decreased suddenly. A linear relationship was found between the effluent SCOD and the TVFA/alkalinity ratio (P ). A relationship was found among the different operational variables (BV , removal efficiency, effluent soluble COD, soluble COD removal rate (R), retention factor (/), specific microbial growth rate (l), methane production rate per volume of reactor and per volume of waste treated––QM and qM , respectively) and the corresponding regression equations were obtained. An increase of BV determined a decrease of removal efficiency, / and qM and an increase of effluent soluble COD, l, R and QM . The value of the maximum specific microbial growth rate (lM ) determined through the equation that correlated BV and l was found to be 0.19 d1 . This value was of the same magnitude as those reported in other works of anaerobic digestion of piggery waste. Ó 2004 Elsevier Ltd. All rights reserved. Keywords: Piggery waste; Upflow anaerobic sludge bed reactor; Organic volumetric loading rate (BV ); Operational parameters; Correlation
1. Introduction Piggery waste is characterized by a high content of organic matter and pathogenic organisms. The disposal of piggery waste without adequate treatment can cause a drastic effect on the environment and human health. This waste is formed by a mixture of manure (feces and urine) and food wastage such as swill and sugar cane molasses (S anchez et al., 2001). According to the most common characteristics of this waste, anaerobic digestion could be considered one of the most promising treatment alternatives (Kimchie et al., 1988; Hobson and Shaw, 1973; Hobson, 1981, 1985, 1992; S anchez
* Corresponding author. Tel.: +34-95-4689654; fax: +34-954691262. E-mail address: [email protected] (R. Borja).
0960-8524/$ - see front matter Ó 2004 Elsevier Ltd. All rights reserved. doi:10.1016/j.biortech.2004.04.003
et al., 1995; Baader, 1990; Yang and Gan, 1998; Parkin and Owen, 1986; Sanchez and Travieso, 1994). Piggery waste treatment by anaerobic digestion has several advantages, the most important are as follows: 1. The process destroys pathogenic and parasitic organisms. 2. Methane as a valuable by-product can be used as a source of energy, and liquid and sludge effluents can be used as a soil conditioner. 3. The low biomass production determines smaller volumes of sludge than in the aerobic processes, thus lowering the cost of sludge management. 4. Capacity to stabilize large volumes of diluted organic slurries at low costs. 5. This process is more economical than the aerobic process, especially at strengths higher than 4 g/l of COD.
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Hobson and Shaw (1973, 1974) studied the mechanism and flora of the anaerobic digestion of piggery waste in a complete mixed digester. They observed that Enterobacteria streptococci made up about 50% of the total content of digester organisms. The predominant groups of organisms were identified as Clostridium butyricum, Bacteroides and gram negative Coccobacilli, and the strict anaerobic ones did not predominate over the facultative ones. Stevens and Schulte (1979) studied the effect of the temperature at solids retention times of between 6 and 55 days at organic volumetric loading rates of between 0.61 and 4.81 kg VS/m3 d, in a complete mixed anaerobic digester. They concluded that at organic rates in the range of 0.61–1.80 kg VS/m3 d and temperatures lower than 25 °C, the operation proceeded satisfactorily. Low temperature digestion was found to require twice as long a retention time with satisfactory production and composition of gas. A previous study (S anchez et al., 2001) looked at the influence of temperature and substrate concentration on the anaerobic batch digestion of piggery wastewater. This study compared the process at mesophilic temperature (35 °C) with temperatures in the range of 16.8–29.5 °C, and influent concentrations in the range of 3.3–26.3 g TCOD/l. The process at mesophilic temperature was more stable than at ambient temperature, obtaining higher values of removal efficiency. Batch or intermittent feeding, plugflow, completely-mixed fixed bed and upflow bed anaerobic digesters have been used for piggery waste treatment at laboratory, pilot and full-scale plants with good results (Baader, 1990; Chen, 1983; Hobson, 1992; I~ niguez et al., 1994; Kimchie et al., 1988; Lapp et al., 1995; Lomas, 1999; Montalvo, 1995; Rodrıguez and Lomas, 1999; S anchez and Travieso, 1994; S anchez et al., 1995, 2001, 2002; Yang and Gan, 1998). The effect of different parameters such as temperature, hydraulic retention time, organic volumetric loading rate, mixing and other factors have been previously studied. However, the upflow anaerobic sludge bed (UASB) reactor has seldom been used for piggery waste treatment because the composition of this waste, rich in nitrogenous compounds, makes the active biomass densification, the granulation process, and therefore the microorganisms retention in the reactor difficult. UASB reactors have
been applied mainly to carbohydrate type wastewaters with success and limited work has been done on the application of this reactor configuration for piggery waste stabilization. Moreover, pigs in Cuba are fed with a mixture of swill and molasses (Sanchez et al., 2001). This could determine different behaviour and performance patterns in Anaerobic Sludge Bed Reactors. Montalvo (1995) studied the separate anaerobic digestion of liquid and solid fractions of piggery waste, respectively, and compared this system to the digestion of the whole waste in a single sludge bed reactor. He found that separate digestion was more effective than that carried out using the single stage reactor. S anchez et al. (1995) compared the performance of sludge bed and anaerobic fixed bed reactors treating piggery wastewaters. The best results were obtained in the latter. The aim of the present work was to evaluate the behaviour and performance of a UASB reactor treating piggery wastewater (rich in carbohydrate) after a preliminary screening and primary sedimentation. A study of the variations of the operational parameters with organic volumetric loading rates and their interrelations by means of different equations was also carried out.
2. Methods 2.1. Piggery waste The piggery waste used in the experiment was obtained from a farm situated 20 km west of Havana City, Cuba. The waste was screened through a 2 mm sieve. After screening, the waste settled for 1 h. The average characteristics and range of the supernatants obtained in primary sedimentation are presented in Table 1. The supernatant after settling was diluted to a TCOD around 8 g/l, value within the range of primary sedimentation supernatant effluents, and this final settled effluent was used to feed the reactor. 2.2. Equipment The experiments were carried out in a laboratoryscale upflow anaerobic sludge bed (UASB) reactor,
Table 1 Characteristics of the raw supernatant piggery waste after a settling period of 60 min Parameter
Average value
Standard deviation
Number of samples
Total COD (mg/l) Total organic carbon (TOC) (mg/l) Total solids (TS) (mg/l) Total volatile solids (TVS) (mg/l) Total suspended solids (TSS) (mg/l) Volatile suspended solids (VSS) (mg/l) Total nitrogen (TN) (mg/l) Phosphates (P) (mg/l) pH
10,189 4100 7210 5122 1637 1166 341 419 6.0
4911 1748 2430 2178 886 813 135 228 0.7
44 44 70 70 70 70 41 41 60
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
consisting essentially of a 5 l cylindrical glass flask 15 cm in diameter and 30 cm tall. The flask was hermetically sealed at the top by a rubber cap with two holes 5 mm in diameter required for biogas outlet and effluent collection, respectively. The biogas was collected in a gas holder submerged in a saturated solution of water and sodium chloride to prevent the absorption of CO2 . The lower part of the reactor was fed by means of a pipe 10 mm in diameter connected to the centre of the reactor allowing the influent to be introduced by using a peristaltic pump. 2.3. Inoculum The upflow anaerobic sludge reactor (UASB) was inoculated with 1 l of anaerobic sludge obtained from a batch piggery sludge digester after 40 days of digestion time. The characteristics of the inoculum are summarized in Table 2.
was analysed by an Orsat analyser to determine the concentration of CO2 . The temperature was maintained at a range of 30–35 °C during the experimental period. 2.5. Analyses Samples of effluent were taken and analysed three times a week and influent once a week. Hence, each run had a duration of seven weeks at steady-state conditions. Analytical determinations of total and soluble chemical oxygen demand (TCOD and SCOD, respectively), total and soluble organic carbon (TOC and SOC, respectively), total suspended solids (TSS), volatile suspended solids (VSS), total volatile fatty acids (TVFA), alkalinity and pH were carried out according to Standard Methods for the Examination of Water and Wastewater (APHA, 1989).
3. Results and discussion
2.4. Experimental procedure Once the inoculum was added to the UASB reactor, the operational volume was completed with distilled water to obtain 5 l of working volume. After this step, the influent was continuously fed by a peristaltic pump at a flow-rate of 0.625 l/d (hydraulic retention time–– HRT––of eight days) until steady-state conditions were obtained. Steady-state was considered after a period equivalent to three times the HRT (24 days). Continuous-mode experiments were carried out at influent flowrates of: 0.625, 0.833, 1.00, 1.25, 1.66, 2.50 and 5.00 l/d, which corresponded to HRT values of 8, 6, 5, 4, 3, 2 and 1 days, respectively, and organic volumetric loading rates (BV ) of 1.0, 1.4, 1.6, 2.0, 2.7, 4.1 and 8.1 g TCOD/ l d, respectively. The experiment was performed by increasing the influent flow-rate and consequently, reducing the HRT at a constant influent concentration. A total of 20 samples were taken for each operational condition and average values were determined for each parameter analysed. Biogas production was determined daily. The gas accumulated in the measurement flask Table 2 Characteristics of the inoculum useda Parameter
Value
TCOD (g/l) TS (g/l) TSS VS (g/l) VSS (g/l) TOC pH
47.5 50.0 41.0 26.1 21.4 15.7 7.3
a
337
Values are averages of five determinations. There was virtually no variation (<3%) among these analyses.
Table 3 shows the steady-state results obtained under the different experimental conditions studied. The increase of BV at fixed value of influent TCOD caused a decrease in the removal efficiency expressed in TCOD, TOC, SCOD, SOC, TSS and VSS. A sudden decrease in removal efficiencies was observed at a BV of 4.1 g TCOD/l d (HRT of two days). The reduction of removal efficiencies was a consequence of the increase in total volatile fatty acids (TVFA) and a simultaneous decrease in the alkalinity causing an increase in the TVFA/alkalinity ratio (P ) and a reduction in the pH value. This situation had a strong influence on the biogas quality, increasing the CO2 concentration on all the range of BV studied. Fig. 1 shows the strong effect of the HRT on the TVFA/alkalinity ratio and, therefore, on the process stability. An increase in the HRT value caused a decrease in P , favouring the process stability. At hydraulic retention times in the range of 4–8 days the process was well balanced and removal efficiencies higher than 70% were obtained in most cases. However, when the HRT was lower than or equal to two days, the removal efficiencies suddenly decreased and values of P increased over 0.5, producing unbalanced conditions. Fig. 2 illustrates the variation of the CO2 concentration as a function of the values of P . The concentration of CO2 increased when the value of P increased from 0.2 to 0.6 and was practically constant at higher P values. Similar behaviour was observed in the acidogenesis of dairy and gelatin-rich wastewaters using upflow anaerobic reactors (Yu and Fang, 2002, 2003). A decrease in the HRT produced an increase in the effluent concentration because the anaerobic waste stabilization was less complete and TVFA concentration and P increased. Fig. 3 illustrates the relationship between the soluble
338
HRT (d)
8
6
5
4
3
2
1
BV (g TCOD/l d)
1
1.4
1.6
2
2.7
4.1
8.1
Parameter
Influent
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
Effluent
E (%)
TCOD SCOD TOC SOC TSS VSS TVFA Alkalinity Pb pH Q (l/l w d) CH4 CO2 qM (l/l w d) QM (l/l r d)
8117 5782 3272 2338 1678 1309 1855 1397 1.1 5.9
1205 785 513 378 306 223
85.4 86.0 84.3 83.8 84.1 83.4
1726 1444 681 590 368 268
78.4 73.5 79.2 74.8 77.3 75.6
2030 1583 706 639 540 353
75.0 72.6 78.4 72.7 67.8 73.0
2141 1772 815 696 692 390
73.6 69.4 75.1 70.2 61.3 70.8
3512 2412 1030 792 804 469
56.7 58.3 68.5 66.1 52.1 64.2
4941 3187 1686 1152 990 548
39.1 44.9 48.5 50.7 41.0 58.1
6611 4644 1755 1410 1223 752
18.6 19.7 34.2 39.7 27.1 42.6
a b
573 1986 0.24 7.4 5.5 66 32.2 3.63 0.45
595 1720 0.29 7.3 4.5 63.3 34.2 2.85 0.47
750 1838 0.34 7.3 4.1 57.8 40.1 2.37 0.47
810 1660 0.41 7.3 3.5 54.7 42.0 1.91 0.48
844 1435 0.49 7.2 3.3 54.2 43.0 1.79 0.60
858 1386 0.52 7.2 2.9 46.6 61.3 1.35 0.68
1211 1030 0.98 6.4 2.6 33.3 63.0 0.86 0.86
Values are averages of 20 determinations taken over three weeks after the steady-state conditions had been reached. The differences between the observed values were less than 3% in all cases. TVFA/alkalinity ratio is expressed in equivalents of acetic acid/equivalents of calcium carbonate.
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
Table 3 Steady-state results obtained under different experimental conditionsa
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344 1.2
100
339
1.2
E % TCOD
90
E (% TCOD removal)
70 0.8 60
0.6
50
40 0.4 30
20 0.2
1
P (eq. acetic acid/eq. calcium carbonate)
1 80
P (eq. acetic acid/eq. calcium carbonate)
P
0.8
0.6
0.4
0.2
10
0 0
2
4
6
0 10
8
0 0
Fig. 1. Variation of the percentage of TCOD removal and TVFA/ alkalinity ratio (P ) with the HRT.
Volumetric fraction of CO2 in the biogas
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0 0
0.2
0.4
0.6
0.8
1
1.2
P (eq. acetic acid/eq. calcium carbonate) Fig. 2. Variation of the volumetric fraction of CO2 in the biogas with the TVFA/alkalinity (P ) ratio.
substrate concentration (SCOD) and the value of P . A straight line was obtained with the following equation: P ¼ 0:182ðSÞ þ 0:056
ð1Þ 2
1
2
3
4
5
6
7
S (g SCOD/l)
HRT (d)
the regression coefficient, R , being equal to 0.97.
Fig. 3. Effect of the effluent SCOD concentration on the TVFA/ alkalinity (P ) ratio.
Eq. (1) showed a strong influence of substrate concentration on the TVFA/alkalinity ratio and, therefore, on process stability. As we already know, TVFA/alkalinity ratio (P ) can be used as a measure of process stability (Fannin, 1987): when this ratio is less than 0.3– 0.4, the process is considered to be operating favourably without the risk of acidification. As was observed in this study, unstable conditions (P > 0:5) appeared in substrate concentrations higher than 2.3 g SCOD/l. The intercept value can be attributed to the presence of mineral acids that have a slight influence and contribution on the P value. As regards the alkalinity values obtained in the effluents of the process, it was observed (Table 3) that the buffering capacity of the experimental system was maintained at favourable levels with appropriate total alkalinity present at BV values in the range of 1.0–4.1 g TCOD/l d. The experimental data obtained in this study indicate that a total alkalinity in the range of 1390–1990 mg/l as CaCO3 was sufficient to prevent the pH from dropping to below 7.2 at BV values of up to 4.1 g TCOD/l d. The pH in the reactor remained more or less constant at HRTs in the range of 8.0–2.0 days, with 7.4 and 7.2 as extreme values. This stability can be attributed to carbonate/bicarbonate buffering. This is produced by the generation of CO2 in the digestion process, which is not completely removed from the reactor as gas (Wheatley, 1990).
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E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
the uptakes of soluble organic carbonaceous compounds were always higher than those obtained for particulate organic carbonaceous compounds. From the results obtained, it can also be demonstrated that suspended COD was previously transformed to soluble COD and then transformed to methane. Fig. 5 shows the variation of the experimental effluent substrate concentrations, S, expressed as g SCOD/l, with the organic loading rate BV . As can be seen, a hyperbolic relationship appears to relate both parameters, which can be described by the following equation:
1 TCOD
0.9
SCOD TOC SOC
0.7
0.6
0.5
S ¼ K1 ½BV =ðK2 þ BV Þ 0.4
0.3
0.2
0.1
0 0
2
4
6
8
10
Bv (g TCOD/l d)
Fig. 4. Variation of the fractional removal efficiency with the organic volumetric loading rate (BV ).
The effect of BV on the removal efficiencies, expressed in TCOD, SCOD, TOC and SOC is given in Fig. 4. As can be seen, an increase in the BV determined a progressive decrease in the removal efficiency, the best removal efficiencies were obtained at BV values in the range of 1.0–4.1 TCOD/l d. At BV of 8.1 g TCOD/l d the removal efficiencies dropped suddenly, and this was concomitant with an increase in P ratio, which achieved values over 0.5. An exponential relationship between BV and the removal efficiencies was observed. The typical equation that correlated both variables is given as follows: E¼e
kBV
ð2Þ
where E, is the fractional removal efficiency and k is a constant rate given in d l/g, its value depending on parameters selected for expressing the substrate concentration (TCOD, SCOD, TOC and SOC). Therefore, the k values obtained for TCOD, SCOD, TOC and SOC were found to be 0.21, 0.20, 0.14 and 0.13 d l/g, respectively. The correlation coefficients were 0.99, 0.99, 0.96 and 0.94 for TCOD, SCOD, TOC and SOC, respectively, and the variability coefficients for the corresponding constants were: 5%, 5%, 11% and 13%, respectively. The values of the constant obtained demonstrated that the rates of diminution of removal efficiencies were decreased with BV in the following order: SOC > TOC > SCOD > TCOD. The results showed that
ð3Þ
where K1 is given in g SCOD/l and represents the theoretical maximum SCOD concentration at the effluent when BV is too high in comparison to K2 . The constant K2 is expressed in g TCOD/l d and represents the BV value for which the effluent SCOD concentration is equivalent to half the value of K1 . A plot of the inverse of S versus the inverse of BV gave a straight line with an intercept equal to 1=K1 and slope equal to K2 =K1 with a correlation factor (R2 ) equal to 0.99. So the values of K1 and K2 were found to be 8.7 ± 0.4 g SCOD/l and 7.3 ± 0.3 g TCOD/l d, respectively. Fig. 5 also shows the theoretical S values, obtained by Eq. (3), and its variation with the BV values. As can be seen, deviations lower
6
5
4
S (g SCOD/l)
Fractional removal efficiency
0.8
3
2
1
0 0
2
4
6
8
10
Bv (g TCOD/l d)
Fig. 5. Effect of the organic volumetric loading rate (BV ) on the experimental effluent SCOD values and those theoretical ones predicted by Eq. (3).
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
ð4Þ
where RM , is the maximum SCOD removal rate (g SCOD/l d), and K is a kinetic constant that represents the value of BV for which the value of R is equal to 1=2RM . The value of the kinetic constants was calculated by means of the plot of the inverse of BV versus the inverse of R. A straight line with a correlation coefficient of 0.97 was obtained. From this linearised equation, RM can be calculated from the intercept on the y-axis and K can be calculated from the slope of the straight line. From this regression, the values of the constants RM and K were estimated to be 2.1 ± 0.2 g SCOD/l d and 2.4 ± 0.3 g TCOD/l d, respectively. Substitution of these values into Eq. (4) allowed the theoretical R values to be determined. Differences between experimental and theoretical values were lower than 10% in all cases. The biogas production and its composition were also a function of the BV used. Fig. 7 shows the effect of BV
1.8
1.6
1.4
R (g SCOD/l d)
1.2
4
1
3.5 0.8 3
2.5
0.6
2 0.4
1.5
1 0.2 0.5
0 0
2
4
6
8
0 10
Bv (g TCOD/l d)
Fig. 7. Effect of the organic volumetric loading rate (BV ) on methane gas production per litre of influent (qM ) and per litre of reactor (QM ).
on the methane production rate (qM ), expressed in l/l waste feed. As can be seen, an increase in BV produced an increase in methane gas production per volume of the reactor (QM ) due to the increase in the total amount of organic matter provided to the reactor at increased BV values. However, a reduction of the biogas amount produced per volume of influent added (qM ) was observed in a parallel way when BV increased. The organic volumetric loading rate (BV ) has a great influence on the anaerobic digestion performance and is an important parameter to the process scale-up. BV is related to the HRT and the influent concentration by the following equation: BV ¼ S0 Q=V ¼ S0 =HRT
1
QM (l/l r.d)
QM (l/l r.d)
R ¼ RM ½BV =ðK þ BV Þ
qm (l/l w.d)
qm (l/l w.d)
than 5% between the experimental and theoretical values of S were always observed. For low BV values the second term of the denominator of this expression can be neglected and Eq. (3) can be transformed into a linear equation, as follows: S ¼ 1:20BV . Fig. 6 illustrates the effect of BV on the SCOD removal rate, R. As can be seen, the SCOD removal rate increased with the BV according to a hyperbolic model, as follows:
341
ð5Þ
where S0 , is the influent concentration (g TCOD/l); Q, is the influent flow-rate (l/d); V , is the volume of the reactor (l); and HRT, is the hydraulic retention time (days), which is equal to the quotient V =Q. At a fixed influent concentration, an increase in BV could be only obtained by an increase in the influent flow-rate (Q) which causes a reduction in HRT. The HRT is related to the microorganisms or solids retention time (SRT) as follows:
0.8
0.6
0.4
0.2
SRT ¼ ½X0 ðHRTÞ=XE
0 0
1
2
3
4
5
6
7
8
9
Bv (g TCOD/l d)
Fig. 6. Variation of the SCOD removal rate (R) with the organic volumetric loading rate (BV ).
ð6Þ
where X0 , is the concentration of microorganisms in the reactor, given in g VSS/l; XE , is the concentration of microorganisms in the effluent of the reactor, given in g VSS/l. The ðX0 =XE Þ ratio is defined as the biomass
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
retention factor ð/Þ (McCarty and Mossey, 1991). The increase in / determines the increase in the SRT and a better adaptation of the microorganisms to the substrate favouring methanogenesis and improving process performance. The increase in the SRT in the anaerobic process can be obtained by increasing the HRT or the / value. Although the increase in the HRT can be achieved by using greater reactor volume or reducing the influent flow, the increase in / using alternatives to prevent the biomass to escape from the process such as sludge recycling or biomass immobilization are the best options. Therefore, HRT can also be expressed as HRT ¼ SRT=/
25
20
Retention factor (φ)
342
15
10
ð7Þ
from which the following equation can be obtained: 5
1=HRT ¼ /=SRT ¼ l/
ð8Þ
taking into account that l, defined as the microbial specific growth rate (d1 ), is numerically equal to the inverse of SRT. Therefore, by combining Eqs. (5) and (8), the following equation was obtained: l ¼ BV =S0 /
ð9Þ
Eq. (9) demonstrates that the microbial specific growth rate and the biomass factor retention are inversely proportional. Therefore, to achieve good operational conditions, the retention factor must be increased, improving the microorganisms separation and reducing the concentration of VSS in the effluent. The maximum value of l will be obtained for a maximum value of BV and a minimum value of /. Fig. 8 shows the effect of BV on the values of /. According to Table 2, the value of X0 was 4.3 g VSS/l, the value of XE being equal to the VSS concentration in the effluent. The concentration of microorganisms in the influent was assumed to be neglected. The value of / decreased for increased BV values, due to the increase in influent flow-rate which in turn caused an increase in the upflow velocity in the reactor. This fact provoked a higher microorganisms concentration in suspension, reducing the SRT. Moreover, an increase in the amount of organic matter added to the reactor produced the development of non-methanogenic microorganisms. As can be seen in Fig. 8, an exponential relationship can be established between BV and /. Eq. (10) demonstrates a strong influence of the organic volumetric loading rate on the retention factor: / ¼ 17:63BV0:57
ð10Þ
the regression coefficient, R2 , being 0.98. Combining Eqs. (9) and (10), a relationship was found that correlates l with BV for the range of the organic loading rates studied:
0 0
1
2
3
4
5
6
7
8
9
Bv (g TCOD/l d)
Fig. 8. Effect of the organic volumetric loading rate (BV ) on the retention factor (/).
l ¼ 7 103 B1:57 V
ð11Þ
This equation allows the theoretical l values to be calculated. The value of lM for the experimental range studied was obtained using Eq. (11). This value was 0.18 d1 and is very close to that obtained by Rodrıguez and Lomas (1999). Hence, the minimum theoretical values of /, SRT and HRT obtained were: 5.4, 5.4 and 1.0 days, respectively. Fig. 9 illustrates the effect of the parameter / on the methane production per volume of influent added to the process (qM ). As can be seen, a straight line was obtained, which is given by the following equation: qM ¼ 0:196/ 0:163
ð12Þ
the regression coefficient, R2 , being 0.98. By combining Eqs. (10) and (12), the following equation is obtained: qM ¼ 3:455BV0:57 0:163
ð13Þ
Finally, Fig. 10 shows the effect of l on the substrate removal rate given in SCOD. As can be clearly observed, an increase in l produced an increase in the substrate removal rate. The curve follows a Michaelis– Menten equation type, given by the following expression: R ¼ 2:99½l=ð0:016 þ lÞ
ð14Þ
E. Sanchez et al. / Bioresource Technology 96 (2005) 335–344
Acknowledgements
4
The authors wish to express their gratitude to the ‘‘Ministerio de Ciencia y Tecnologıa’’ (project REN 2001-0472/TECNO), and ‘‘Ministerio de Educaci on, Cultura y Deportes’’ for providing financial support.
3.5
3
References
qm (l/l w.d)
2.5
2
1.5
1
0.5
0 0
5
10
15
20
25
Retention factor (φ)
Fig. 9. Effect of the retention factor (/) on the methane production per litre of waste influent added (qM ).
1.6
1.4
1.2
R (g SCOD/l d)
343
1
0.8
0.6
0.4
0.2
0 0
0.02 0.04 0.06 0.08
0.1
0.12 0.14 0.16 0.18
0.2
µ (1/d) Fig. 10. Effect of the specific microbial growth rate (l) on the SCOD removal rate (R).
the regression coefficient, R2 , and standard error being equal to 0.97 and 10%, respectively.
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