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Effect of total solids on fecal coliform regrowth in anaerobically digested biosolids
water research 42 (2008) 3817–3825
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Effect of total solids on fecal coliform regrowth in anaerobically digested biosolids Yinan Qia, Steven K. Dentela,*, Diane S. Hersonb a
Department of Civil and Environmental Engineering, University of Delaware, Newark, DE 19716, USA Department of Biological Sciences, University of Delaware, Newark, DE 19716, USA
b
article info
abstract
Article history:
Fecal coliform (FC) concentrations in anaerobically digested biosolids can increase during
Received 6 November 2007
centrifugal dewatering and afterwards in storage of dewatered cake. The immediate in-
Received in revised form 4 May 2008
crease after centrifugation (reactivation) has been demonstrated to be the revitalization of
Accepted 3 June 2008
fecal coliforms that had become non-culturable. The increase during storage (regrowth)
Available online 21 June 2008
has been regarded as a subsequence of reactivated bacteria growing in a favorable environment. In this paper, however, regrowth is demonstrated without preceding reactivation,
Keywords:
using intensive laboratory centrifugation to duplicate the levels of regrowth seen in full-
Fecal coliform
scale centrifugation. Higher total solids (TS) levels of the dewatered biosolids lead to
Regrowth
greater magnitudes of FC increase. The final TS level appears much more important
Reactivation
than the level of shear imposed during centrifugation, based on comparison of different
Methanogens
centrifugation/dilution procedures used to obtain similar TS levels. The greater TS levels
Sludge
also reduce methane production, suggesting that methanogens compete with, or inhibit,
Biosolids
the fecal coliforms. The addition of bromoethanesulfonate as a methanogen-specific inhib-
Centrifugation
itor decreased the production of methane gas, and also increased the number of fecal
Dewatering
coliforms. ª 2008 Elsevier Ltd. All rights reserved.
1.
Introduction
Anaerobic digestion processes have long been known to successfully reduce the number of pathogens and indicator organisms if operated properly (Puchajda and Oleszkiewicz, 2006). However, researchers in the last a few years have reported higher fecal coliform (FC) population, on a dry solids basis, in centrifugally dewatered biosolids compared to the digester effluents (Iranpour et al., 2002, 2003; Erdal et al., 2003; Hendrickson et al., 2004; Monteleone et al., 2004; Qi et al., 2004, 2007; Cooper et al., 2005; Jolis, 2006). For example, Iranpour et al. (2002, 2003) found that after dewatering and placement in a silo, the most probable number (MPN) of FC in the thermophilically digested
biosolids increased from less than 103 MPN/g dry solids (DS) to 106 MPN/g DS. With this type of drastic increase, the dewatered biosolids no longer meet the US EPA Class A biosolids limits. Erdal et al. (2003) reported an increase of FC MPN of more than one order of magnitude after centrifugation of mesophilically digested biosolids. Moreover, the FC MPN further increased after 1 day of storage. Regardless of the differences in the digestion processes and centrifuges used, these observed increases in the numbers of indicator microorganisms can be categorized as either reactivation or regrowth depending on when the increase occurs. In practice, reactivation is referred to as an increase in FC or Escherichia coli density in the biosolids immediately after
* Corresponding author. Department of Civil and Environmental Engineering, University of Delaware, 348 DuPont Hall, Newark, DE 19716, USA. Tel.: þ1 302 831 8120; fax: þ1 302 831 3640. E-mail address: [email protected] (S.K. Dentel). 0043-1354/$ – see front matter ª 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2008.06.001
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centrifugation or other dewatering processes compared to the feed to the dewatering equipment. Regrowth has been regarded as an additional increase in FC or E. coli density in dewatered biosolids during further storage over a period of time. Although the observed FC increase could be the growth during centrifugation, the residence time of the process is normally less than 15 min (Dentel et al., 2000), which is less than the doubling time for E. coli. Therefore, order of magnitude increases in FC density cannot be solely due to growth during the centrifugation process. One possible explanation is that the FC enters a so-called viable but non-culturable (VBNC) state during digestion in response to environmental stresses such as low nutrient levels, low or high temperature, or the presence of toxic substances (Lisle et al., 1998; Rockabrand et al., 1999; Grey and Steck, 2001). Microorganisms in the VBNC state are still viable but cannot be detected using standard culturing methods. Thus Higgins et al. (2007a,b) combined enumeration of E. coli by culturing and by competitive polymerase chain reaction (cPCR) to indicate that some of the E. coli after digestion had become VBNC, with centrifugation evidently triggering their reactivation. However, increases in E. coli and FC densities during the storage of dewatered biosolids could not be linked to VBNC populations. These increases should then be considered as a separate phenomenon. This distinction between reactivation and regrowth has been previously shown (Qi et al., 2007), by measuring FC density in mesophilic, anaerobically digested biosolids, before and after dewatering, from four wastewater treatment plants. No FC regrowth was found in two of these plants immediately after centrifugation, but after 24 h incubation at 37 C, the FC MPNs in dewatered biosolids from these plants increased by more than two orders of magnitude. The digested biosolids that had not been dewatered showed no such increase after the same incubation. Apparently, the fecal coliforms did not become VBNC during the digestion in those two treatment plants, or even if some of them did, they were not reactivated by centrifugation. The regrowth during storage, though rapid, was within the bounds of documented growth kinetics (Dentel et al., 2007). Thus, regrowth, even in the absence of reactivation, can be an important concern with regard to biosolids management. When FC regrowth occurs after dewatering, it might be attributable to either of two phenomena: (1) a change in the aqueous phase constituents, or (2) the increased solids concentration. Centrifugal dewatering could change aqueous phase characteristics due to shear, compression, or a microbial response to the process. For example, volatile fatty acids (VFAs) that can accumulate in the digester during anaerobic digestion have been reported to be toxic to some microorganisms (such as fecal coliforms, methanogens, and Salmonella) present in biosolids (Kunte et al., 2000; Puchajda and Oleszkiewicz, 2006; Puchajda et al., 2006). Other studies have shown that the shearing force during centrifugation may break down the protein into more labile and bioavailable form (Murthy et al., 2002, 2003; Higgins et al., 2003). Moreover, it has been hypothesized (e.g. Hendrickson et al., 2004; Higgins et al., 2007a,b) that an activating substance (such as an autoinducer) is released during centrifugation. Thus, the regrowth of FC in dewatered biosolids might be
attributed to (1) the relief from the stress of inhibitory substances such as VFAs and sulfide compounds with the removal of centrate; (2) the release of extra nutrients or activating substance by the shearing force during centrifugation. On the other hand, the greater solids (TS) concentration (or the lower available water concentration) could affect microbial mobility or nutrients and enzymes transport. In biosolids, the availability or chemical activity of water may decrease due to capillary effects or close association to hydrophilic constituents such as extracellular proteins and polysaccharides. Consequently, the free water might be considerably less than the water concentration as indicated by a total solids analysis. This phenomenon has been quantified in terms of water activity (aw), which is defined as the vapor pressure of water divided by that of pure water at the same temperature. Microbial activity is reduced at certain levels of aw, and the limiting level differs by microbe (e.g. Beuchat, 1981; Beuchat and Scouten, 2002; Web Ref 1). For example, E. coli has a higher aw tolerance (0.95, Web Ref 1) than that of methanogens (0.97, Hiraishi et al., 2005). As dewatering decreases aw, species that are less tolerant will be inhibited. If fecal coliforms have higher tolerance of aw than other competitive species, their growth would be favored in an environment that has higher solids concentration such as in dewatered biosolids.
2.
Materials and methods
2.1.
Sample collection and storage
The wastewater treatment plants where samples were collected employ mesophilic anaerobic digestion. Digested, dewatered biosolids and centrate were collected according to EPA Method 1681 (USEPA, 2005). Digested samples were collected 5–10 min before the corresponding dewatered samples to approximate the solids residence time in the centrifuge. Consistent with Method 1681, samples were transported to the laboratory on ice and maintained at 4 C until enumeration, which was conducted within 24 h of sampling. To avoid non-uniformity through sedimentation, digested biosolids and centrate were well mixed by hand shaking before obtaining subsamples.
2.2.
Media and reagent preparation
A-1 media (Difco) and EPA dilution water (Hach) were prepared according to manufacture’s instructions. After autoclaving, A-1 media were kept in cold (4 C) and brought back to room temperature 1–2 h before use; EPA dilution water was kept at room temperature.
2.3.
Experimental procedures and methods
Various treatments were applied to the samples to investigate different hypotheses. Detailed procedures are as follows.
2.3.1.
Sample preparation
All samples were homogenized before they were enumerated for fecal coliforms. Digested biosolids samples were homogenized by blending at high speed (15K rpm) for 3 min in
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a conventional kitchen blender. Dewatered biosolids samples were also blended by this method, but first diluted to a liquid consistency in the same step by combining 15 g of the dewatered biosolids with 135 mL of distilled water. Blending was conducted in a 4 C controlled temperature room to prevent temperature increases during the procedure. Samples were then serially diluted (1:10 ratio) with EPA dilution water to desired concentrations for FC enumeration. Some experiments included adding additional chemicals, centrifugation of digested biosolids, or incubation of samples for specified times and temperatures. The flocculant polymer (Clarifloc NE-694 by SNF-Polydyne) obtained from the treatment plant was added to digested biosolids at a 1:10 polymer: dry solids ratio to assess the effect of a high polymer dose. 2Bromoethanesulfonate (BES) obtained from Aldrich (Milwaukee) was added to some digested samples as well, with the final concentrations ranging from 10 mM to 500 mM.
2.3.2.
Variation of solute and solids properties
Three different methods were used to vary the solute and solids contents in samples. These also served to vary the solute properties.
distilled, deionized (DI) water to achieve TS similar to samples using Method 2. These conditions are also shown in Table 1. All centrifugations were conducted in a 4 C environment to minimize possible heating of sample. Sufficient volumes of lab-centrifuged samples were generated to create similar total sample and headspace volumes at all solids concentrations.
2.3.3.
2.3.4. 2.3.2.1. Method 1. To investigate centrate effects, digested biosolids (designated D) were centrifuged in the laboratory for 15 min at 18K rpm (39,100 g), the supernatant removed, and the remaining solids rediluted with DI water to the original volume (this treatment is designated as DCDI). Dewatered biosolids from the same plant (designated DW) were also diluted with plant centrate to achieve a TS level similar to the digested biosolids (this treatment is designated as DWC). D, DCDI, DW, and DWC samples were assayed for fecal coliform immediately after preparation and also after 24 h incubation at 37 C.
2.3.2.2. Method 2. To obtain a range of solids concentrations from the original digested biosolids, samples were produced by centrifuging digested biosolids at different rpm and time combinations, and in two cases, by recombining centrate with the spun down solids. Thus, all these solids were combinations of the original solid and liquid phases, but at different relative amounts. In increasing TS levels, these samples were designated DC-1 through DC-4, as detailed in Table 1.
liquid but produce samples at various solids concentrations, lab-centrifuged digested biosolids were then rediluted with
Purging
In some experiments, digested biosolids were purged with air to remove volatile substances and gases. Purging was conducted with 200 mL samples for 3 h at a flow rate of 200 mL/min.
2.3.5.
FC enumeration
The 5-tube most probable number (MPN) method with A-1 media was used for fecal coliform enumeration, as described previously (Qi et al., 2007). The protocols of EPA Method 1681 (USEPA, 2005) were followed for sample collection, transport, storage, and incubation. In order to normalize comparisons, all MPNs are expressed per gram dry solids (with total solids (TS) measured according to Standard Methods, 2005). Upper and lower 95% confidence intervals (CIs) were used for error bars. Previous work assessed other possible influences on the enumeration procedure, such as media type, diluent used, interference by Bacillus spp., and artifacts due to aggregation of cells; none of these factors were found to be significant (Qi et al., 2007).
2.3.6. 2.3.2.3. Method 3. To remove the majority of the centrate
Incubation
Samples were incubated at 37 C and 25 C to investigate the effects of increased temperature and prolonged incubation on FC regrowth. If no gas production measurements were involved, 40 mL of liquid samples were incubated in 50 mL centrifuge tubes with the cap loosely closed; solids samples were added to the tubes to the 40 mL mark without compression. For gas production tests, 200 mL liquid samples were incubated in 250 mL brown glass bottles; solids samples were placed in 30 mL centrifuge tubes with less than 10 mL headspace left. Headspaces were purged with helium gas to assure anaerobic conditions.
Gas production measurement
Methane and carbon dioxide concentrations were measured using a gas chromatograph (HP 5890; Restek column 19715, 15 m, 0.53 mm ID, Rt-QPLOT packing) using thermal
Table 1 – Centrifugation and recombining conditions used to generate samples shown in Fig. 3a and b Centrifuge conditions None 40 mL in 50 mL tubes for 15 min at 6K rpm (5860 g) 30 mL in 40 mL tubes for 15 min at 18K rpm (39,100 g)
Treatment
Final TS (3a) (%)
None 20 mL centrate removed, shaken 25 mL centrate removed, shaken All centrate removed, shaken All centrate removed, shaken All centrate removed, DI water added, shaken
2.2 4.3 5.8 12.8 21.7
Final TS (3b) (%)
2.3 3.4 16.1 21 2.4 4.2 19
Sample designation D DC-1 DC-2 DC-3 DC-4 DC-1-DI DC-2-DI DC-3-DI
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3.
Results and discussion
3.1.
Regrowth in the absence of reactivation
Fig. 1 shows the change in FC MPNs in digested and dewatered biosolids during a period of 31 days incubation at 25 C. The error bars represent the 95% CIs. Mesophilic digestion and centrifugal dewatering were used in the treatment plant where the samples were collected. The solids contents before and after dewatering were 2.1% and 19.1%, respectively. As shown in Fig. 1, digested and dewatered biosolids had similar FC MPNs immediately before and after centrifugation. In other words, no reactivation of fecal coliforms was observed. However, during incubation at 25 C, the FC MPNs in dewatered biosolids increased from 105 to nearly 108 per gram dry solids after 6 days (exceeded the Class B limitation of 2 106/g DS), gradually decreasing thereafter. In contrast, the FC MPNs in digested biosolids remained constant in the first few days and gradually decreased thereafter. The remarkable increases in dewatered biosolids MPNs demonstrate that regrowth is not necessarily associated with reactivation, and that the dewatered biosolids have much more regrowth potential than do their pre-dewatered counterparts. In view of public health and environmental risks, this indicates that the point of sampling for enumeration should be reconsidered since the potential for regrowth is not taken into account in current practices.
3.2.
Solute concentrations and characteristics
Fig. 2 shows the results of generating different solute and solids from digested biosolids using Method 1. Comparing
10 Digested Biosolids
FC log MPN/g dry solids
Dewatered Biosolids 8
6
4
2
0 0
1
6
10
31
Days Fig. 1 – FC MPNs during long term storage in digested and dewatered biosolids at 25 8C. Error bars represent 95% confidence intervals (CIs).
8 Before Incubation After Incubation
FC log MPN/g dry solids
conductivity detector (Valco Instruments). Helium was used as the carrier gas at a flow rate of 480 mL/min. An integrator was coupled with the detector to record and analyze data. Standard methane and carbon dioxide gases (Alltech) were used for calibration.
7
6
5
4
3 D
DCDI
DW
DWC
Fig. 2 – Centrate toxicity test. MPNs were measured before and after incubation at 37 8C for 24 h. Error bars represent 95% CIs. D – digested biosolids; DCDI – lab-centrifuged digested biosolids, rediluted with DI water; DW – dewatered biosolids; DWC – dewatered biosolids rediluted with centrate.
the four samples, there were no significant differences in fecal coliform density before incubation. This indicates that there was no reactivation due to centrifugation, obviating the hypothesis that centrifugation releases a fecal coliform autoinducer. However, the 24-h incubation did lead to significant effects on fecal coliform numbers. The DWC sample showed an increase as significant as the dewatered biosolids, whereas no regrowth was observed in the DCDI sample or the digested sludge. The separation and removal of toxic or inhibitory substances by centrifugation is therefore contradicted, because this would have led to regrowth in the DCDI sample and prevented it in the DWC sample. Actually, a solid–liquid separation process such as centrifugation cannot change the concentration of a soluble species in either the centrate or the dewatered cake. Consequently, the effect of an inducer, inhibitor, or toxic agent should not be altered by centrifugation, because the species concentration, rather than its amount relative to the solids mass, provides the effect. The same may be said for an inhibitor or toxic agent of which the effect is concentration dependent. There are some situations in which differences in a soluble chemical species concentration relative to the solids mass may have an effect, even if the initial solution concentrations are identical. If the species serves as a nutrient, it will be depleted sooner in the presence of a greater solids concentration due to the greater biomass activity. Similarly, a product of microbial metabolism will accumulate faster when the solids concentration is greater. An autoinducer might also accumulate faster only if it is generated after solid–liquid separation, rather than before: in this case, it would be at significantly higher concentrations in the liquid phase of the dewatered cake than in the centrate. However, the results are consistent with an introduction of a nutrient during the full-scale centrifugation process. For example, shear of the microbial solids could lead to the
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3.3.
7.5
a DC-3 6.5 DC-2
6.0
DC-1 5.5
D
5.0 4.5
b DC-2 5.5 DC-2-DI
The higher fecal coliform counts after intensive centrifugation might have stemmed from either the centrifugation process itself, or from the higher solids concentrations it produced. In order to make this distinction, Methods 2 and 3 were used to generate various samples from digested biosolids. Fig. 3a shows the results using Method 2: after 24 h incubation at 37 C, all centrifuged samples showed an increase in fecal coliform MPNs, from less than one order of magnitude to as many as two orders of magnitudes. The methods used to generate samples DC-1 and DC-2 should be noted. The solids in these samples were subjected to the same centrifugation conditions used for sample DC-3, but then diluted with centrate to different extents. That the regrowth amounts differed suggests that the solids concentration or solute properties are more critical than the centrifugation conditionsdsuch as degree of compressiondin determining the extent of regrowth. Results using Method 3 are shown in Fig. 3b. The indicated samples were rediluted with distilled deionized water to obtain the desired solids concentrations, yet the FC MPNs after incubation were almost the same as their counterparts from direct centrifugation. This also suggests that the intensity of centrifugation is not related to the extent of FC regrowth, and again indicates that the solute properties are not crucial; rather, the concentration of solids, or the liquid:solid ratio, governs regrowth.
Table 2 – COD measurements for biosolids samples after plant and lab centrifugation and filtration processes
Filtered lab centrate (0.45 mm)
COD (mg/L)
TS (%)
490 520 400 14,100 370 330 350 330
0.112 0.122 1.99
DC-3-DI
DC-1
5.0
Plant centrate Lab centrate, high g centrifugation Filtrate from digested biosolids (0.45 mm) Digested biosolids Filtered plant centrate (0.45 mm)
DC-3
6.0
Effect of TS on fecal coliform regrowth
Sample
DC-4
7.0
FC log MPN/g dry solids
separation of gel-like extracellular material with a lesser density than the actual bacterial cells. If this were to release some soluble fraction, then the liquid phase in both the dewatered and centrate fractions would include it, allowing regrowth. The failure of lab-scale centrifugation to duplicate this phenomenon would be due to the lack of shear in a desktop centrifuge. To evaluate the above possibility, COD measurements were made of various samples with the same treatment used for samples shown in Fig. 2. Table 2 shows that the centrifuged and filtered samples were uniformly low in COD. The plant centrate COD was essentially the same as the lab centrate whether filtered or not, and both were the same as filtrate obtained directly from the digested biosolids. These findings make it unlikely that a source of COD, released during centrifugation, would serve either as a nutrient for fecal coliforms or as an inhibitor to a competing population.
4.5 DC-1-DI
4.0 3.5 0
5
10
15
20
25
TS (%) Fig. 3 – FC MPNs in lab-centrifuged samples with different TS after incubation for 24 h at 37 8C. Error bars represents the 95% CIs. (a) Digested biosolids at various concentrations created by lab centrifugation then recombination of centrate and solids. (b) Digested biosolids centrifuged as in (a) or centrifuged to a high concentration then rediluted with distilled water. Table 1 provides details on preparation and concentration of these samples.
This finding is at odds with prior work suggesting shearing effects as a major factor in FC reactivation and regrowth. For example, Erdal et al. (2003) enumerated FC in biosolids dewatered by high solids centrifuge, low solids centrifuge, and belt filter press, finding more reactivation and regrowth in the high solids cakes than in the low solids cakes. No reactivation was found after dewatering by belt filter press, and only moderate regrowth (<10) with 1 day of storage. Yet our results show that a lab centrifuge engenders regrowth comparable in magnitude to that in a full-scale decanter centrifuge (2 logs vs. 3 logs), even though the shear occurring in the latter is undoubtedly much more severe. This apparent contradiction could be due to several factors: (1) the belt filter press may have produced cake with lower TS than from the centrifuges (cake TS values were not reported); (2) influential pretreatments such as the added polymer dose might have differed; or (3) differing exposures to oxygen in each process could have played a role. While the effects of shear, polymer, and oxygen on FC regrowth still need to be differentiated, the results in Fig. 3 clearly demonstrate that the final solids concentration plays an important role in FC regrowth.
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As indicated above, a higher solids concentration should lead to more rapid depletion of the lesser supply of nutrients available in the liquid phase. This might be linked to the more rapid growth of fecal coliforms as follows. Since counts of fecal coliforms indicate them to be a minor constituent of the microbial biomass (Dentel et al., 2007), available nutrients are used primarily by other populations which generate reduced organics and a low redox potentialdconditions that halt fecal coliform activity, but only until usable organics and an electron acceptor are again available. This phenomenon has been clearly identified for two enteric bacteria: Shigella (Freter, 1962) and E. coli (Freter and Ozawa, 1963). Savage (1977) suggested that a variety of obligate anaerobes could serve as the competing population. In anaerobically digested sludge, the anaerobic population of acetogens and methanogens would be the likely cause.
3.4.
Effect of TS on methane production
If an anaerobic consortium of acetogens and methanogens represents competition to fecal coliforms, then the production of methane should be indicative of its extent, and this should be (inversely) related to fecal coliform regrowth. Digested biosolids were therefore centrifuged as described in Table 1 to reach 10% (DC-3) and 19% (DC-4) TS. They were incubated together with uncentrifuged digested biosolids (D) and plant-dewatered biosolids (DW) at 25 C and 37 C. The methane production of each sample was measured to indicate relative methanogen activity after 24 h incubation. The results are shown in Fig. 4. Methane production is normalized to the total volume present during incubation. Fig. 4 shows greater methane production at higher solids concentrations. The linear results suggest substrate saturation for DW and DC-4 samples, so that the rate of methane production is determined by the microbial density rather than the substrate concentration. When this is the case, the methane productions should be divided by the solids mass present in the sample in order to normalize to unit microbial activity. The digested (D) and slightly dewatered (DC-3) samples show some curvature. The lower solids concentrations
translate into lesser substrate availability. Normalization by solids mass is appropriate in these cases because the substrate concentration is well represented by the solids mass. Fig. 5 shows the relative methane productions using this basis. It should first be noted that the plant-dewatered solids, whether incubated for 24 h at the dewatered concentration, or diluted to the original digested concentration, show essentially the same methane production after this normalization, supporting this means of representation. Also in Fig. 5, methane production from the sample taken after plant dewatering is matched by the sample obtained by lab centrifugation (DC-4), which was close to the same solids concentration. It has been postulated (Chen et al., 2005) that centrifugation disrupts methanogens through the shearing that occurs in the full-scale, butdassuming that lab centrifugation is less shear-intensive than dewatering centrifugationdthis conjecture is not supported by these data. The sample DC-3, however, exhibits a higher methane production rate than DC-4 when centrifuged under lab conditions but only to a lower TS level. This again suggests that the solids level is significant rather than the amount of shear imparted. Normalizing on a solids basis shows a surprisingly high methane production for the digested sample, especially for the initial period. This sample was at the lowest solids concentration and the initially high rate might be attributable to more readily degradable substrate in the liquid phase (soluble species or smaller size materials with greater specific surface area). This source being depleted, the rate then slowed over time. Nonetheless, the final methane production was still higher than for any dewatered sample. Thus, although methane production as depicted in Fig. 4 increases with solids concentration, the normalized trend in Fig. 5 is the opposite. In other words, the methanogens face an increasingly limiting or competitive environment due to the greater solids concentration, which includes a greater amount of active biomass as well as substrate. If the higher initial rate in the digested sample stemmed from more readily degradable substrate in solution, the addition of a polymer flocculant would mitigate this effect by
D DC-4 DC-3 DW-S DW-L DA BES 10mM BES 50mM Polymer
CH4 production umol/g DS
CH4 production umol/mL sample
120 D DC-4 DC-3 DW
100
80
60
40
600
400
200
20
0 0 0
5
10
15
20
25
Time (hr) Fig. 4 – Effect of TS on methane production during incubation at 37 8C. Table 1 gives sample details. Methane production is normalized to volume of each sample.
5
10
15
20
25
Time (hr) Fig. 5 – Effect of various sample treatments on methane production during incubation at 37 8C. Table 1 gives sample details. Methane production is normalized to mass of solids in each sample.
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Table 3 – CH4 production of different samples (digested, centrifuged and dewatered) after 24 h incubation at 37 8C and 25 8C
D DC-4 DC-3 DW-L DW-S DA BES 10 mM BES 50 mM BES 200 mM BES 500 mM
CH4 production (mM/g dry solids)
Solids content (%)
Digested biosolids Digested, lab-centrifuged, high solids Digested, lab-centrifuged, low solids Dewatered, rediluted w/EPADW Dewatered biosolids Digested, aerated Digested, w/10 mM BES Digested, w/50 mM BES Digested, w/200 mM BES Digested, w/500 mM BES
capturing this fraction and incorporating it into the solids. Fig. 5 shows that this was the case when a high dose of cationic flocculant was added to the digested sample prior to incubation. The later increase in gas production with polymer addition may stem from biodegradation of the polymer itself (Chang and Dentel, 1999; Chang et al., 2001). Aeration of an anaerobically digested sample (DA) led to much more severe drop in methane production as seen in Fig. 5. Likewise, samples dosed with 2-bromoethanesulfonate (BES) exhibited inhibition. BES is a structural analog of coenzyme M (CoM) (Gunsalus et al., 1978), which is responsible for the terminal step of methane biosynthesis (Vogels et al., 1988) and present only in methanogens (DiMarco et al., 1990). Thus, BES is known to be a methanogen-specific inhibitor. The doses used were high but this was evidently necessary, as shown by the greater effect of the 50 mM dose compared to the 10 mM dose. The low methane production for the higher dose may have been residual gas produced prior to BES addition. Likewise, samples incubated at 25 C instead of 37 C were not amenable to methanogens but produced 30–75 mM methane/g dry solids (Table 3), perhaps for this same reason. It was hypothesized that limitations on the specific growth rate of methanogens might be related to regrowth of fecal coliform. In this case, the addition of BES would not only decrease methane production, but also lead to greater fecal coliform levels. Fig. 6 shows the FC MPNs in biosolids with different BES concentrations after 24 h incubations at 25 C and 37 C. The differences in FC MPNs at 25 C are very small among the samples, indicating that BES has little effect on fecal coliforms because of the low microbial metabolism at low temperature. This is consistent with the methane production results. At 37 C, however, the FC MPNs were noticeably affected. MPNs increased as the BES concentration increased up to 200 mM, where the MPN increased by nearly three orders of magnitude. BES additions in excess of 200 mM led to MPN decreases, indicating that high BES hindered the growth of fecal coliforms. When the BES concentration was increased to 500 mM, the FC MPN decreased to lower than that without BES addition. Chiu and Lee (2001) have shown that BES can alter or inhibit microbial communities beyond the methanogenic fraction. The relationship between fecal coliforms and methanogens is interesting. When the methanogen activity decreases,
1.7–1.9 20–24 10–11 2.6–2.9 26–29 1.7–1.9 1.7–1.9 1.7–1.9 1.7–1.9 1.7–1.9
37 C
25 C
620.73 436.26 613.43 356.31 429.21 177.55 119.60 130.70 37.12 49.85
74.05 73.35 71.23 30.45 43.36 – – 64.07 54.33 29.78
the FC density increases. A decrease of methanogen activity could be achieved by either increasing the solids concentration or adding BES to inhibit methanogens. Methanogens and fecal coliforms do not compete directly for most substrates. Methanogens, as obligate anaerobes, utilize simple reduced organics as well as hydrogen and carbon dioxide. Fecal coliform bacteria, such as E. coli, utilize simple sugars and amino acids and are more likely to compete with other obligate anaerobes such as Clostridium (Gujer and Zehnder, 1983). In an environment where most organic substrates are in the solid phase, requiring hydrolysis and degradation before being imbibed, the fecal coliforms may possess a greater repertoire of appropriate enzymes. This population would thus be favored in the environment of a higher solids concentration, explaining fecal coliform regrowth after dewatering. Some researchers have shown that the centrifugal dewatering process increases odor emissions from biosolids (Murthy et al., 2002, 2003; Higgins et al., 2003). These studies have implicated shearing forces during centrifugation. Our results suggest
8
FC log MPN/g dry solids
Sample description
37 oC 25 oC
7 6 5 4 3 2
D
M
M
M
0m
S1
BE
0m
S5
BE
0m
10
S BE
mM
00
S2 BE
M
0m
30
S BE
M 0m
50
S BE
Fig. 6 – Effects of BES on FC MPNs after biosolids samples were incubated for 24 h at 37 8C and 25 8C. Error bars represent 95% CIs. The down arrow on the last column of 37 8C means the data are below the detection limit for the dilutions used. D – digested biosolids; BES 10–500 mM – digested biosolids with BES additions at different concentrations.
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an alternative explanation: the higher solids concentration in centrifugal dewatering reduces the activity of methanogens. Methanogens are responsible for demethylating volatile sulfur compounds (VSCs), which are odorous, thereby deodorizing the biosolids (Higgins et al., 2003). With methanogen activity reduced during centrifugation, VSCs build up and odor levels increase.
4.
Conclusions
Significant regrowth of fecal coliform bacteria can occur in the absence of reactivation, suggesting that these phenomena need to be differentiated. High speed laboratory centrifugation, without provision for simulating the shear in a decanter centrifuge, can lead to the same magnitude of fecal coliform regrowth. The solution phase of biosolids does not appear to contain any constituents (autoinducer, inhibitor, or substrate) that lead to regrowth, either by affecting fecal coliforms directly or competing populations. Higher solids concentrations in digested biosolids lead to greater overall methane production, but lesser specific methanogens activity. The inhibition of methanogens by BES exacerbates fecal coliform regrowth, suggesting that high solids levels have the same effect.
references
Beuchat, L.R., Scouten, A.J., 2002. Combined effects of water activity, temperature and chemical treatments on the survival of Salmonella and Escherichia coli O157:H7 on alfalfa seeds. J. Appl. Microbiol. 92, 382–395. Beuchat, L.R., 1981. Microbial stability as affected by water activity. Cereal Foods World 26 (7), 345–349. Chang, L.L., Dentel, S.K., 1999. Effects of a cationic polymer on biological treatment processes, Proceedings of the WEFTEC’99, Water Environment Federation. Chang, L.L., Raudenbush, D.L., Dentel, S.K., 2001. Aerobic and anaerobic biodegradability of a flocculant polymer. Water Sci. Technol. 44 (2–3), 461–468. Chen, Y., Higgins, M.J., Maas, N.A., Murthy, S.N., Toffey, W., Foster, D.J., 2005. Roles of methanogens on volatile organic sulfur compound production in anaerobically digested wastewater biosolids. Water Sci. Technol. 62 (1–2), 67–72. Chiu, P.C., Lee, M., 2001 May. 2-Bromoethanesulfonate affects bacteria in a trichloroethene-dechlorinating culture. Appl. Environ. Microbiol. 67 (5), 2371–2374. Cooper, J., Cass, J., Rogers, M., Smith, S.R., 2005. Update on the effects of sludge dewatering on E. coli enumeration. In: Proceedings of the 10th European Biosolids and Biowastes Conference Workshop, November 13–16, Wakefield, UK. Dentel, S.K., Abu-Orf, M.M., Walker, C.A., 2000. Optimization of slurry flocculation and dewatering based on electrokinetic and rheological phenomena. Chem. Eng. J. 80 (1–3), 65–72. Dentel, S.K., Qi, Y.N., Herson, D.S., 2007. Improving the assessment of risk from pathogens in biosolids: fecal coliform regrowth, survival, enumeration, and assessment. In: Proceedings Moving Forward Wastewater Biosolids Sustainability: Technical, Managerial, and Public Synergy, IWA Specialist Conference, June 25–27, 2007, Moncton, Canada.
DiMarco, A.A., Bobik, Bobik, T.A., Wolfe, R.S., 1990. Unusual coenzymes of methanogenesis. Annu. Rev. Biochem. 59, 355–394. Erdal, Z.K., Mendenhall, T.C., Neely, S.K., Wagoner, D.L., Quigley, C., 2003. Implementing improvements in a North Carolina residuals management program. In: Proceedings of the WEF/AWWA Joint Residuals and Biosolids Management Conference. Water Environment Federation, Alexandria, VA. Freter, R., 1962. In vivo and in vitro antagonism of intestinal bacteria against Shigellaflexneri. II. The inhibitory mechanism. J. Infect. Dis. 110, 38–46. Freter, R., Ozawa, A., 1963. Explanation for limitation of populations of Escherichia coli in broth cultures. J. Bacteriol. 86, 904–910. Grey, B., Steck, T.R., 2001. Concentrations of copper thought to be toxic to E. coli can induce the viable but nonculturable condition. Appl. Environ. Microbiol. 67, 5325–5327. Gujer, W., Zehnder, A.J.B., 1983. Conversion processes in anaerobic digestion. Water Sci. Technol. 15, 127–167. Gunsalus, R.P., Romesser, J.A., Wolfe, R.S., 1978. Preparation of coenzyme M analogues and their activity in the methylcoenzyme M reductase system of Methanobacterium thermoautotrophicum. Biochemistry 17, 2374–2377. Hendrickson, D.A., Denard, D., Farrell, J., Higgins, M.J., Murthy, S.N., 2004. Reactivation of fecal coliforms after anaerobic digestion and dewatering. In: Proceedings of the WEF/WEAU Residuals Biosolids Conference. Water Environment Federation, Alexandria, VA. Higgins, M.J., Chen, Y.C., Hendrickson, D.A., Murthy, S.N., 2007a. Evaluation of bacterial pathogen and indicator densities after dewatering of anaerobically digested biosolids phase II and phase III. WERF Final Report. Higgins, M.J., Chen, Y.C., Murthy, S.N., Hendrickson, D.A., Farrel, J., Schafer, P., 2007b. Reactivation and growth of nonculturable indicator bacteria in anaerobically digested biosolids after centrifuge dewatering. Water Res. 41, 665–673. Higgins, M.J., Yarosz, D., Chen, Y., Murthy, S.N., 2003. Mechanisms for volatile sulfur compound and odor production in digested biosolids. In: Proceedings of Water Environment Federation and AWWA Annual Biosolids and Residuals Conference, Baltimore, MD. Hiraishi, A., Nakaya, T., Takaoka, K., Toyama, M., 2005. Biological treatment for decomposition of pollutants by microorganism under controlled water activity. Jpn. Kokai Tokkyo Koho 12 pp. CODEN: JKXXAF JP 2005270695. Iranpour, R., Cox, H.H.J., Oh, S., Ardent, T., Mohamed, F., Netto, H., Fan, S., Kearney, R.J., 2003. Occurrence of fecal coliform and Salmonella spp. following thermophilic digestion and postdigestion processing at the Hyperion treatment plant. In: Proceedings of the WEF/AWWA/CWEA Joint Residuals and Biosolids Management Conference. Water Environment Federation, Alexandria, VA. Iranpour, R., Oh, S., Cox, H.J., Samar, P., Taylor, D., Mohamed, F., Hagekhalil, A., Kearney, R.J., 2002. Effects of dewatering on bacteria inactivation: centrifuge simulation and field tests at the Hyperion treatment plant. In: Proceedings of the WEFTEC 2002. Water Environment Federation, Alexandria, VA. Jolis, D., 2006. Regrowth of fecal coliforms in Class A biosolids. Water Environ. Res. 78 (4), 442–445. Kunte, D.P., Yeole, T.Y., Ranade, D.R., 2000. Effect of volatile fatty acids on Shigella dysenteriae during anaerobic digestion of human night soil. World J. Microbiol. Biotechnol. 16, 519–552. Lisle, J.T., Broadaway, S.C., Prescott, A.M., Pyle, B.H., Fricker, C., McFeters, G.A., 1998. Effects of starvation on physiological activity and chlorine disinfection resistance in Escherichia coli O157:H7. Appl. Environ. Microbiol. 64, 4658–4662.
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Monteleone, M.C., Furness, D., Jefferson, B., Cartmell, E., 2004. Fate of E. coli across mechanical dewatering processes. Environ. Technol. 25, 825–831. Murthy, S.N., Forbes, B., Burrowes, P., Esqueda, T., Glindemann, D., Novak, J., Higgins, M.J., Mendenhall, T., Toffey, W., Peot, C., 2002. Impact of high shear solids processing on odor production from anaerobically digested biosolids. WEFTEC 2002, Chicago, IL. Murthy, S.N., Higgins, M.J., Chen, Y., Toffey, W., Golembeski, J., 2003. Influence of solids characteristics and dewatering process on volatile sulfur compound production from anaerobically digested biosolids. In: Proceedings of the Water Environment Federation and AWWA Biosolids and Residuals Conference, Baltimore, MD. Puchajda, B., Oleszkiewicz, J., 2006. Extended acid digestion for inactivation of fecal coliforms. Water Environ. Res. 78 (12), 2389–2396. Puchajda, B., Oleszkiewicz, J., Sparling, R., Reimers, R., 2006. Lowtemperature inactivation of fecal coliforms in sludge digestion. Water Environ. Res. 78 (7), 680–685. Qi, Y.N., Dentel, S.K., Herson, D.S., 2007. Increases in fecal coliform bacteria resulting from centrifugal dewatering of digested biosolids. Water Res. 41, 571–580.
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Qi, Y.N., Gillow, S., Herson, D.S., Dentel, S.K., 2004. Reactivation and/or growth of fecal coliform bacteria during centrifugal dewatering of anaerobically digested biosolids. Water Sci. Technol. 50 (9), 115–120. Rockabrand, D., Austin, T., Kaiser, R., Blum, P., 1999. Bacterial growth state distinguished by single-cell protein profiling: does chlorination kill coliforms in municipal effluent? Appl. Environ. Microbiol. 65, 4181–4188. Savage, D., 1977. Microbial ecology of the gastrointestinal tract. Ann. Rev. Microbiol. 31, 107–133. Standard Methods, 2005. In: Eaton, A.D., Clesceri, L.S., Rice, E.W., Greenberg, A.E. (Eds.), Standard Methods for the Examination of Water and Wastewater, 21st ed. APHA–AWWA–WEF, Washington, DC. United States EPA (US EPA), 2005. Method 1681: fecal coliforms in sewage sludge (biosolids) by multiple-tube fermentation using A-1 medium. EPA-821-R-04-027. Vogels, G.D., Keltjens, J.T., van der Drift, C., 1988. Biochemistry of methane production. In: Zehnder, A.J.B. (Ed.), Biology of Anaerobic Microorganisms. John Wiley & Sons, New York, NY, pp. 707–770. Web Reference 1: http://www.foodtechsource.com/rcenter/tech_ data/td_water.htm.
Available at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/watres
Effect of total solids on fecal coliform regrowth in anaerobically digested biosolids Yinan Qia, Steven K. Dentela,*, Diane S. Hersonb a
Department of Civil and Environmental Engineering, University of Delaware, Newark, DE 19716, USA Department of Biological Sciences, University of Delaware, Newark, DE 19716, USA
b
article info
abstract
Article history:
Fecal coliform (FC) concentrations in anaerobically digested biosolids can increase during
Received 6 November 2007
centrifugal dewatering and afterwards in storage of dewatered cake. The immediate in-
Received in revised form 4 May 2008
crease after centrifugation (reactivation) has been demonstrated to be the revitalization of
Accepted 3 June 2008
fecal coliforms that had become non-culturable. The increase during storage (regrowth)
Available online 21 June 2008
has been regarded as a subsequence of reactivated bacteria growing in a favorable environment. In this paper, however, regrowth is demonstrated without preceding reactivation,
Keywords:
using intensive laboratory centrifugation to duplicate the levels of regrowth seen in full-
Fecal coliform
scale centrifugation. Higher total solids (TS) levels of the dewatered biosolids lead to
Regrowth
greater magnitudes of FC increase. The final TS level appears much more important
Reactivation
than the level of shear imposed during centrifugation, based on comparison of different
Methanogens
centrifugation/dilution procedures used to obtain similar TS levels. The greater TS levels
Sludge
also reduce methane production, suggesting that methanogens compete with, or inhibit,
Biosolids
the fecal coliforms. The addition of bromoethanesulfonate as a methanogen-specific inhib-
Centrifugation
itor decreased the production of methane gas, and also increased the number of fecal
Dewatering
coliforms. ª 2008 Elsevier Ltd. All rights reserved.
1.
Introduction
Anaerobic digestion processes have long been known to successfully reduce the number of pathogens and indicator organisms if operated properly (Puchajda and Oleszkiewicz, 2006). However, researchers in the last a few years have reported higher fecal coliform (FC) population, on a dry solids basis, in centrifugally dewatered biosolids compared to the digester effluents (Iranpour et al., 2002, 2003; Erdal et al., 2003; Hendrickson et al., 2004; Monteleone et al., 2004; Qi et al., 2004, 2007; Cooper et al., 2005; Jolis, 2006). For example, Iranpour et al. (2002, 2003) found that after dewatering and placement in a silo, the most probable number (MPN) of FC in the thermophilically digested
biosolids increased from less than 103 MPN/g dry solids (DS) to 106 MPN/g DS. With this type of drastic increase, the dewatered biosolids no longer meet the US EPA Class A biosolids limits. Erdal et al. (2003) reported an increase of FC MPN of more than one order of magnitude after centrifugation of mesophilically digested biosolids. Moreover, the FC MPN further increased after 1 day of storage. Regardless of the differences in the digestion processes and centrifuges used, these observed increases in the numbers of indicator microorganisms can be categorized as either reactivation or regrowth depending on when the increase occurs. In practice, reactivation is referred to as an increase in FC or Escherichia coli density in the biosolids immediately after
* Corresponding author. Department of Civil and Environmental Engineering, University of Delaware, 348 DuPont Hall, Newark, DE 19716, USA. Tel.: þ1 302 831 8120; fax: þ1 302 831 3640. E-mail address: [email protected] (S.K. Dentel). 0043-1354/$ – see front matter ª 2008 Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2008.06.001
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centrifugation or other dewatering processes compared to the feed to the dewatering equipment. Regrowth has been regarded as an additional increase in FC or E. coli density in dewatered biosolids during further storage over a period of time. Although the observed FC increase could be the growth during centrifugation, the residence time of the process is normally less than 15 min (Dentel et al., 2000), which is less than the doubling time for E. coli. Therefore, order of magnitude increases in FC density cannot be solely due to growth during the centrifugation process. One possible explanation is that the FC enters a so-called viable but non-culturable (VBNC) state during digestion in response to environmental stresses such as low nutrient levels, low or high temperature, or the presence of toxic substances (Lisle et al., 1998; Rockabrand et al., 1999; Grey and Steck, 2001). Microorganisms in the VBNC state are still viable but cannot be detected using standard culturing methods. Thus Higgins et al. (2007a,b) combined enumeration of E. coli by culturing and by competitive polymerase chain reaction (cPCR) to indicate that some of the E. coli after digestion had become VBNC, with centrifugation evidently triggering their reactivation. However, increases in E. coli and FC densities during the storage of dewatered biosolids could not be linked to VBNC populations. These increases should then be considered as a separate phenomenon. This distinction between reactivation and regrowth has been previously shown (Qi et al., 2007), by measuring FC density in mesophilic, anaerobically digested biosolids, before and after dewatering, from four wastewater treatment plants. No FC regrowth was found in two of these plants immediately after centrifugation, but after 24 h incubation at 37 C, the FC MPNs in dewatered biosolids from these plants increased by more than two orders of magnitude. The digested biosolids that had not been dewatered showed no such increase after the same incubation. Apparently, the fecal coliforms did not become VBNC during the digestion in those two treatment plants, or even if some of them did, they were not reactivated by centrifugation. The regrowth during storage, though rapid, was within the bounds of documented growth kinetics (Dentel et al., 2007). Thus, regrowth, even in the absence of reactivation, can be an important concern with regard to biosolids management. When FC regrowth occurs after dewatering, it might be attributable to either of two phenomena: (1) a change in the aqueous phase constituents, or (2) the increased solids concentration. Centrifugal dewatering could change aqueous phase characteristics due to shear, compression, or a microbial response to the process. For example, volatile fatty acids (VFAs) that can accumulate in the digester during anaerobic digestion have been reported to be toxic to some microorganisms (such as fecal coliforms, methanogens, and Salmonella) present in biosolids (Kunte et al., 2000; Puchajda and Oleszkiewicz, 2006; Puchajda et al., 2006). Other studies have shown that the shearing force during centrifugation may break down the protein into more labile and bioavailable form (Murthy et al., 2002, 2003; Higgins et al., 2003). Moreover, it has been hypothesized (e.g. Hendrickson et al., 2004; Higgins et al., 2007a,b) that an activating substance (such as an autoinducer) is released during centrifugation. Thus, the regrowth of FC in dewatered biosolids might be
attributed to (1) the relief from the stress of inhibitory substances such as VFAs and sulfide compounds with the removal of centrate; (2) the release of extra nutrients or activating substance by the shearing force during centrifugation. On the other hand, the greater solids (TS) concentration (or the lower available water concentration) could affect microbial mobility or nutrients and enzymes transport. In biosolids, the availability or chemical activity of water may decrease due to capillary effects or close association to hydrophilic constituents such as extracellular proteins and polysaccharides. Consequently, the free water might be considerably less than the water concentration as indicated by a total solids analysis. This phenomenon has been quantified in terms of water activity (aw), which is defined as the vapor pressure of water divided by that of pure water at the same temperature. Microbial activity is reduced at certain levels of aw, and the limiting level differs by microbe (e.g. Beuchat, 1981; Beuchat and Scouten, 2002; Web Ref 1). For example, E. coli has a higher aw tolerance (0.95, Web Ref 1) than that of methanogens (0.97, Hiraishi et al., 2005). As dewatering decreases aw, species that are less tolerant will be inhibited. If fecal coliforms have higher tolerance of aw than other competitive species, their growth would be favored in an environment that has higher solids concentration such as in dewatered biosolids.
2.
Materials and methods
2.1.
Sample collection and storage
The wastewater treatment plants where samples were collected employ mesophilic anaerobic digestion. Digested, dewatered biosolids and centrate were collected according to EPA Method 1681 (USEPA, 2005). Digested samples were collected 5–10 min before the corresponding dewatered samples to approximate the solids residence time in the centrifuge. Consistent with Method 1681, samples were transported to the laboratory on ice and maintained at 4 C until enumeration, which was conducted within 24 h of sampling. To avoid non-uniformity through sedimentation, digested biosolids and centrate were well mixed by hand shaking before obtaining subsamples.
2.2.
Media and reagent preparation
A-1 media (Difco) and EPA dilution water (Hach) were prepared according to manufacture’s instructions. After autoclaving, A-1 media were kept in cold (4 C) and brought back to room temperature 1–2 h before use; EPA dilution water was kept at room temperature.
2.3.
Experimental procedures and methods
Various treatments were applied to the samples to investigate different hypotheses. Detailed procedures are as follows.
2.3.1.
Sample preparation
All samples were homogenized before they were enumerated for fecal coliforms. Digested biosolids samples were homogenized by blending at high speed (15K rpm) for 3 min in
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a conventional kitchen blender. Dewatered biosolids samples were also blended by this method, but first diluted to a liquid consistency in the same step by combining 15 g of the dewatered biosolids with 135 mL of distilled water. Blending was conducted in a 4 C controlled temperature room to prevent temperature increases during the procedure. Samples were then serially diluted (1:10 ratio) with EPA dilution water to desired concentrations for FC enumeration. Some experiments included adding additional chemicals, centrifugation of digested biosolids, or incubation of samples for specified times and temperatures. The flocculant polymer (Clarifloc NE-694 by SNF-Polydyne) obtained from the treatment plant was added to digested biosolids at a 1:10 polymer: dry solids ratio to assess the effect of a high polymer dose. 2Bromoethanesulfonate (BES) obtained from Aldrich (Milwaukee) was added to some digested samples as well, with the final concentrations ranging from 10 mM to 500 mM.
2.3.2.
Variation of solute and solids properties
Three different methods were used to vary the solute and solids contents in samples. These also served to vary the solute properties.
distilled, deionized (DI) water to achieve TS similar to samples using Method 2. These conditions are also shown in Table 1. All centrifugations were conducted in a 4 C environment to minimize possible heating of sample. Sufficient volumes of lab-centrifuged samples were generated to create similar total sample and headspace volumes at all solids concentrations.
2.3.3.
2.3.4. 2.3.2.1. Method 1. To investigate centrate effects, digested biosolids (designated D) were centrifuged in the laboratory for 15 min at 18K rpm (39,100 g), the supernatant removed, and the remaining solids rediluted with DI water to the original volume (this treatment is designated as DCDI). Dewatered biosolids from the same plant (designated DW) were also diluted with plant centrate to achieve a TS level similar to the digested biosolids (this treatment is designated as DWC). D, DCDI, DW, and DWC samples were assayed for fecal coliform immediately after preparation and also after 24 h incubation at 37 C.
2.3.2.2. Method 2. To obtain a range of solids concentrations from the original digested biosolids, samples were produced by centrifuging digested biosolids at different rpm and time combinations, and in two cases, by recombining centrate with the spun down solids. Thus, all these solids were combinations of the original solid and liquid phases, but at different relative amounts. In increasing TS levels, these samples were designated DC-1 through DC-4, as detailed in Table 1.
liquid but produce samples at various solids concentrations, lab-centrifuged digested biosolids were then rediluted with
Purging
In some experiments, digested biosolids were purged with air to remove volatile substances and gases. Purging was conducted with 200 mL samples for 3 h at a flow rate of 200 mL/min.
2.3.5.
FC enumeration
The 5-tube most probable number (MPN) method with A-1 media was used for fecal coliform enumeration, as described previously (Qi et al., 2007). The protocols of EPA Method 1681 (USEPA, 2005) were followed for sample collection, transport, storage, and incubation. In order to normalize comparisons, all MPNs are expressed per gram dry solids (with total solids (TS) measured according to Standard Methods, 2005). Upper and lower 95% confidence intervals (CIs) were used for error bars. Previous work assessed other possible influences on the enumeration procedure, such as media type, diluent used, interference by Bacillus spp., and artifacts due to aggregation of cells; none of these factors were found to be significant (Qi et al., 2007).
2.3.6. 2.3.2.3. Method 3. To remove the majority of the centrate
Incubation
Samples were incubated at 37 C and 25 C to investigate the effects of increased temperature and prolonged incubation on FC regrowth. If no gas production measurements were involved, 40 mL of liquid samples were incubated in 50 mL centrifuge tubes with the cap loosely closed; solids samples were added to the tubes to the 40 mL mark without compression. For gas production tests, 200 mL liquid samples were incubated in 250 mL brown glass bottles; solids samples were placed in 30 mL centrifuge tubes with less than 10 mL headspace left. Headspaces were purged with helium gas to assure anaerobic conditions.
Gas production measurement
Methane and carbon dioxide concentrations were measured using a gas chromatograph (HP 5890; Restek column 19715, 15 m, 0.53 mm ID, Rt-QPLOT packing) using thermal
Table 1 – Centrifugation and recombining conditions used to generate samples shown in Fig. 3a and b Centrifuge conditions None 40 mL in 50 mL tubes for 15 min at 6K rpm (5860 g) 30 mL in 40 mL tubes for 15 min at 18K rpm (39,100 g)
Treatment
Final TS (3a) (%)
None 20 mL centrate removed, shaken 25 mL centrate removed, shaken All centrate removed, shaken All centrate removed, shaken All centrate removed, DI water added, shaken
2.2 4.3 5.8 12.8 21.7
Final TS (3b) (%)
2.3 3.4 16.1 21 2.4 4.2 19
Sample designation D DC-1 DC-2 DC-3 DC-4 DC-1-DI DC-2-DI DC-3-DI
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3.
Results and discussion
3.1.
Regrowth in the absence of reactivation
Fig. 1 shows the change in FC MPNs in digested and dewatered biosolids during a period of 31 days incubation at 25 C. The error bars represent the 95% CIs. Mesophilic digestion and centrifugal dewatering were used in the treatment plant where the samples were collected. The solids contents before and after dewatering were 2.1% and 19.1%, respectively. As shown in Fig. 1, digested and dewatered biosolids had similar FC MPNs immediately before and after centrifugation. In other words, no reactivation of fecal coliforms was observed. However, during incubation at 25 C, the FC MPNs in dewatered biosolids increased from 105 to nearly 108 per gram dry solids after 6 days (exceeded the Class B limitation of 2 106/g DS), gradually decreasing thereafter. In contrast, the FC MPNs in digested biosolids remained constant in the first few days and gradually decreased thereafter. The remarkable increases in dewatered biosolids MPNs demonstrate that regrowth is not necessarily associated with reactivation, and that the dewatered biosolids have much more regrowth potential than do their pre-dewatered counterparts. In view of public health and environmental risks, this indicates that the point of sampling for enumeration should be reconsidered since the potential for regrowth is not taken into account in current practices.
3.2.
Solute concentrations and characteristics
Fig. 2 shows the results of generating different solute and solids from digested biosolids using Method 1. Comparing
10 Digested Biosolids
FC log MPN/g dry solids
Dewatered Biosolids 8
6
4
2
0 0
1
6
10
31
Days Fig. 1 – FC MPNs during long term storage in digested and dewatered biosolids at 25 8C. Error bars represent 95% confidence intervals (CIs).
8 Before Incubation After Incubation
FC log MPN/g dry solids
conductivity detector (Valco Instruments). Helium was used as the carrier gas at a flow rate of 480 mL/min. An integrator was coupled with the detector to record and analyze data. Standard methane and carbon dioxide gases (Alltech) were used for calibration.
7
6
5
4
3 D
DCDI
DW
DWC
Fig. 2 – Centrate toxicity test. MPNs were measured before and after incubation at 37 8C for 24 h. Error bars represent 95% CIs. D – digested biosolids; DCDI – lab-centrifuged digested biosolids, rediluted with DI water; DW – dewatered biosolids; DWC – dewatered biosolids rediluted with centrate.
the four samples, there were no significant differences in fecal coliform density before incubation. This indicates that there was no reactivation due to centrifugation, obviating the hypothesis that centrifugation releases a fecal coliform autoinducer. However, the 24-h incubation did lead to significant effects on fecal coliform numbers. The DWC sample showed an increase as significant as the dewatered biosolids, whereas no regrowth was observed in the DCDI sample or the digested sludge. The separation and removal of toxic or inhibitory substances by centrifugation is therefore contradicted, because this would have led to regrowth in the DCDI sample and prevented it in the DWC sample. Actually, a solid–liquid separation process such as centrifugation cannot change the concentration of a soluble species in either the centrate or the dewatered cake. Consequently, the effect of an inducer, inhibitor, or toxic agent should not be altered by centrifugation, because the species concentration, rather than its amount relative to the solids mass, provides the effect. The same may be said for an inhibitor or toxic agent of which the effect is concentration dependent. There are some situations in which differences in a soluble chemical species concentration relative to the solids mass may have an effect, even if the initial solution concentrations are identical. If the species serves as a nutrient, it will be depleted sooner in the presence of a greater solids concentration due to the greater biomass activity. Similarly, a product of microbial metabolism will accumulate faster when the solids concentration is greater. An autoinducer might also accumulate faster only if it is generated after solid–liquid separation, rather than before: in this case, it would be at significantly higher concentrations in the liquid phase of the dewatered cake than in the centrate. However, the results are consistent with an introduction of a nutrient during the full-scale centrifugation process. For example, shear of the microbial solids could lead to the
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3.3.
7.5
a DC-3 6.5 DC-2
6.0
DC-1 5.5
D
5.0 4.5
b DC-2 5.5 DC-2-DI
The higher fecal coliform counts after intensive centrifugation might have stemmed from either the centrifugation process itself, or from the higher solids concentrations it produced. In order to make this distinction, Methods 2 and 3 were used to generate various samples from digested biosolids. Fig. 3a shows the results using Method 2: after 24 h incubation at 37 C, all centrifuged samples showed an increase in fecal coliform MPNs, from less than one order of magnitude to as many as two orders of magnitudes. The methods used to generate samples DC-1 and DC-2 should be noted. The solids in these samples were subjected to the same centrifugation conditions used for sample DC-3, but then diluted with centrate to different extents. That the regrowth amounts differed suggests that the solids concentration or solute properties are more critical than the centrifugation conditionsdsuch as degree of compressiondin determining the extent of regrowth. Results using Method 3 are shown in Fig. 3b. The indicated samples were rediluted with distilled deionized water to obtain the desired solids concentrations, yet the FC MPNs after incubation were almost the same as their counterparts from direct centrifugation. This also suggests that the intensity of centrifugation is not related to the extent of FC regrowth, and again indicates that the solute properties are not crucial; rather, the concentration of solids, or the liquid:solid ratio, governs regrowth.
Table 2 – COD measurements for biosolids samples after plant and lab centrifugation and filtration processes
Filtered lab centrate (0.45 mm)
COD (mg/L)
TS (%)
490 520 400 14,100 370 330 350 330
0.112 0.122 1.99
DC-3-DI
DC-1
5.0
Plant centrate Lab centrate, high g centrifugation Filtrate from digested biosolids (0.45 mm) Digested biosolids Filtered plant centrate (0.45 mm)
DC-3
6.0
Effect of TS on fecal coliform regrowth
Sample
DC-4
7.0
FC log MPN/g dry solids
separation of gel-like extracellular material with a lesser density than the actual bacterial cells. If this were to release some soluble fraction, then the liquid phase in both the dewatered and centrate fractions would include it, allowing regrowth. The failure of lab-scale centrifugation to duplicate this phenomenon would be due to the lack of shear in a desktop centrifuge. To evaluate the above possibility, COD measurements were made of various samples with the same treatment used for samples shown in Fig. 2. Table 2 shows that the centrifuged and filtered samples were uniformly low in COD. The plant centrate COD was essentially the same as the lab centrate whether filtered or not, and both were the same as filtrate obtained directly from the digested biosolids. These findings make it unlikely that a source of COD, released during centrifugation, would serve either as a nutrient for fecal coliforms or as an inhibitor to a competing population.
4.5 DC-1-DI
4.0 3.5 0
5
10
15
20
25
TS (%) Fig. 3 – FC MPNs in lab-centrifuged samples with different TS after incubation for 24 h at 37 8C. Error bars represents the 95% CIs. (a) Digested biosolids at various concentrations created by lab centrifugation then recombination of centrate and solids. (b) Digested biosolids centrifuged as in (a) or centrifuged to a high concentration then rediluted with distilled water. Table 1 provides details on preparation and concentration of these samples.
This finding is at odds with prior work suggesting shearing effects as a major factor in FC reactivation and regrowth. For example, Erdal et al. (2003) enumerated FC in biosolids dewatered by high solids centrifuge, low solids centrifuge, and belt filter press, finding more reactivation and regrowth in the high solids cakes than in the low solids cakes. No reactivation was found after dewatering by belt filter press, and only moderate regrowth (<10) with 1 day of storage. Yet our results show that a lab centrifuge engenders regrowth comparable in magnitude to that in a full-scale decanter centrifuge (2 logs vs. 3 logs), even though the shear occurring in the latter is undoubtedly much more severe. This apparent contradiction could be due to several factors: (1) the belt filter press may have produced cake with lower TS than from the centrifuges (cake TS values were not reported); (2) influential pretreatments such as the added polymer dose might have differed; or (3) differing exposures to oxygen in each process could have played a role. While the effects of shear, polymer, and oxygen on FC regrowth still need to be differentiated, the results in Fig. 3 clearly demonstrate that the final solids concentration plays an important role in FC regrowth.
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As indicated above, a higher solids concentration should lead to more rapid depletion of the lesser supply of nutrients available in the liquid phase. This might be linked to the more rapid growth of fecal coliforms as follows. Since counts of fecal coliforms indicate them to be a minor constituent of the microbial biomass (Dentel et al., 2007), available nutrients are used primarily by other populations which generate reduced organics and a low redox potentialdconditions that halt fecal coliform activity, but only until usable organics and an electron acceptor are again available. This phenomenon has been clearly identified for two enteric bacteria: Shigella (Freter, 1962) and E. coli (Freter and Ozawa, 1963). Savage (1977) suggested that a variety of obligate anaerobes could serve as the competing population. In anaerobically digested sludge, the anaerobic population of acetogens and methanogens would be the likely cause.
3.4.
Effect of TS on methane production
If an anaerobic consortium of acetogens and methanogens represents competition to fecal coliforms, then the production of methane should be indicative of its extent, and this should be (inversely) related to fecal coliform regrowth. Digested biosolids were therefore centrifuged as described in Table 1 to reach 10% (DC-3) and 19% (DC-4) TS. They were incubated together with uncentrifuged digested biosolids (D) and plant-dewatered biosolids (DW) at 25 C and 37 C. The methane production of each sample was measured to indicate relative methanogen activity after 24 h incubation. The results are shown in Fig. 4. Methane production is normalized to the total volume present during incubation. Fig. 4 shows greater methane production at higher solids concentrations. The linear results suggest substrate saturation for DW and DC-4 samples, so that the rate of methane production is determined by the microbial density rather than the substrate concentration. When this is the case, the methane productions should be divided by the solids mass present in the sample in order to normalize to unit microbial activity. The digested (D) and slightly dewatered (DC-3) samples show some curvature. The lower solids concentrations
translate into lesser substrate availability. Normalization by solids mass is appropriate in these cases because the substrate concentration is well represented by the solids mass. Fig. 5 shows the relative methane productions using this basis. It should first be noted that the plant-dewatered solids, whether incubated for 24 h at the dewatered concentration, or diluted to the original digested concentration, show essentially the same methane production after this normalization, supporting this means of representation. Also in Fig. 5, methane production from the sample taken after plant dewatering is matched by the sample obtained by lab centrifugation (DC-4), which was close to the same solids concentration. It has been postulated (Chen et al., 2005) that centrifugation disrupts methanogens through the shearing that occurs in the full-scale, butdassuming that lab centrifugation is less shear-intensive than dewatering centrifugationdthis conjecture is not supported by these data. The sample DC-3, however, exhibits a higher methane production rate than DC-4 when centrifuged under lab conditions but only to a lower TS level. This again suggests that the solids level is significant rather than the amount of shear imparted. Normalizing on a solids basis shows a surprisingly high methane production for the digested sample, especially for the initial period. This sample was at the lowest solids concentration and the initially high rate might be attributable to more readily degradable substrate in the liquid phase (soluble species or smaller size materials with greater specific surface area). This source being depleted, the rate then slowed over time. Nonetheless, the final methane production was still higher than for any dewatered sample. Thus, although methane production as depicted in Fig. 4 increases with solids concentration, the normalized trend in Fig. 5 is the opposite. In other words, the methanogens face an increasingly limiting or competitive environment due to the greater solids concentration, which includes a greater amount of active biomass as well as substrate. If the higher initial rate in the digested sample stemmed from more readily degradable substrate in solution, the addition of a polymer flocculant would mitigate this effect by
D DC-4 DC-3 DW-S DW-L DA BES 10mM BES 50mM Polymer
CH4 production umol/g DS
CH4 production umol/mL sample
120 D DC-4 DC-3 DW
100
80
60
40
600
400
200
20
0 0 0
5
10
15
20
25
Time (hr) Fig. 4 – Effect of TS on methane production during incubation at 37 8C. Table 1 gives sample details. Methane production is normalized to volume of each sample.
5
10
15
20
25
Time (hr) Fig. 5 – Effect of various sample treatments on methane production during incubation at 37 8C. Table 1 gives sample details. Methane production is normalized to mass of solids in each sample.
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Table 3 – CH4 production of different samples (digested, centrifuged and dewatered) after 24 h incubation at 37 8C and 25 8C
D DC-4 DC-3 DW-L DW-S DA BES 10 mM BES 50 mM BES 200 mM BES 500 mM
CH4 production (mM/g dry solids)
Solids content (%)
Digested biosolids Digested, lab-centrifuged, high solids Digested, lab-centrifuged, low solids Dewatered, rediluted w/EPADW Dewatered biosolids Digested, aerated Digested, w/10 mM BES Digested, w/50 mM BES Digested, w/200 mM BES Digested, w/500 mM BES
capturing this fraction and incorporating it into the solids. Fig. 5 shows that this was the case when a high dose of cationic flocculant was added to the digested sample prior to incubation. The later increase in gas production with polymer addition may stem from biodegradation of the polymer itself (Chang and Dentel, 1999; Chang et al., 2001). Aeration of an anaerobically digested sample (DA) led to much more severe drop in methane production as seen in Fig. 5. Likewise, samples dosed with 2-bromoethanesulfonate (BES) exhibited inhibition. BES is a structural analog of coenzyme M (CoM) (Gunsalus et al., 1978), which is responsible for the terminal step of methane biosynthesis (Vogels et al., 1988) and present only in methanogens (DiMarco et al., 1990). Thus, BES is known to be a methanogen-specific inhibitor. The doses used were high but this was evidently necessary, as shown by the greater effect of the 50 mM dose compared to the 10 mM dose. The low methane production for the higher dose may have been residual gas produced prior to BES addition. Likewise, samples incubated at 25 C instead of 37 C were not amenable to methanogens but produced 30–75 mM methane/g dry solids (Table 3), perhaps for this same reason. It was hypothesized that limitations on the specific growth rate of methanogens might be related to regrowth of fecal coliform. In this case, the addition of BES would not only decrease methane production, but also lead to greater fecal coliform levels. Fig. 6 shows the FC MPNs in biosolids with different BES concentrations after 24 h incubations at 25 C and 37 C. The differences in FC MPNs at 25 C are very small among the samples, indicating that BES has little effect on fecal coliforms because of the low microbial metabolism at low temperature. This is consistent with the methane production results. At 37 C, however, the FC MPNs were noticeably affected. MPNs increased as the BES concentration increased up to 200 mM, where the MPN increased by nearly three orders of magnitude. BES additions in excess of 200 mM led to MPN decreases, indicating that high BES hindered the growth of fecal coliforms. When the BES concentration was increased to 500 mM, the FC MPN decreased to lower than that without BES addition. Chiu and Lee (2001) have shown that BES can alter or inhibit microbial communities beyond the methanogenic fraction. The relationship between fecal coliforms and methanogens is interesting. When the methanogen activity decreases,
1.7–1.9 20–24 10–11 2.6–2.9 26–29 1.7–1.9 1.7–1.9 1.7–1.9 1.7–1.9 1.7–1.9
37 C
25 C
620.73 436.26 613.43 356.31 429.21 177.55 119.60 130.70 37.12 49.85
74.05 73.35 71.23 30.45 43.36 – – 64.07 54.33 29.78
the FC density increases. A decrease of methanogen activity could be achieved by either increasing the solids concentration or adding BES to inhibit methanogens. Methanogens and fecal coliforms do not compete directly for most substrates. Methanogens, as obligate anaerobes, utilize simple reduced organics as well as hydrogen and carbon dioxide. Fecal coliform bacteria, such as E. coli, utilize simple sugars and amino acids and are more likely to compete with other obligate anaerobes such as Clostridium (Gujer and Zehnder, 1983). In an environment where most organic substrates are in the solid phase, requiring hydrolysis and degradation before being imbibed, the fecal coliforms may possess a greater repertoire of appropriate enzymes. This population would thus be favored in the environment of a higher solids concentration, explaining fecal coliform regrowth after dewatering. Some researchers have shown that the centrifugal dewatering process increases odor emissions from biosolids (Murthy et al., 2002, 2003; Higgins et al., 2003). These studies have implicated shearing forces during centrifugation. Our results suggest
8
FC log MPN/g dry solids
Sample description
37 oC 25 oC
7 6 5 4 3 2
D
M
M
M
0m
S1
BE
0m
S5
BE
0m
10
S BE
mM
00
S2 BE
M
0m
30
S BE
M 0m
50
S BE
Fig. 6 – Effects of BES on FC MPNs after biosolids samples were incubated for 24 h at 37 8C and 25 8C. Error bars represent 95% CIs. The down arrow on the last column of 37 8C means the data are below the detection limit for the dilutions used. D – digested biosolids; BES 10–500 mM – digested biosolids with BES additions at different concentrations.
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an alternative explanation: the higher solids concentration in centrifugal dewatering reduces the activity of methanogens. Methanogens are responsible for demethylating volatile sulfur compounds (VSCs), which are odorous, thereby deodorizing the biosolids (Higgins et al., 2003). With methanogen activity reduced during centrifugation, VSCs build up and odor levels increase.
4.
Conclusions
Significant regrowth of fecal coliform bacteria can occur in the absence of reactivation, suggesting that these phenomena need to be differentiated. High speed laboratory centrifugation, without provision for simulating the shear in a decanter centrifuge, can lead to the same magnitude of fecal coliform regrowth. The solution phase of biosolids does not appear to contain any constituents (autoinducer, inhibitor, or substrate) that lead to regrowth, either by affecting fecal coliforms directly or competing populations. Higher solids concentrations in digested biosolids lead to greater overall methane production, but lesser specific methanogens activity. The inhibition of methanogens by BES exacerbates fecal coliform regrowth, suggesting that high solids levels have the same effect.
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