Bioresource Technology 116 (2012) 170–178

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Bioresource Technology

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Batch anaerobic co-digestion of proteins and carbohydrates

Elsayed Elbeshbishy a,b,⇑, George Nakhla a

a Department of Chemical and Biochemical Engineering, University of Western Ontario, London, Ontario, Canada N6A 5B9b Trojan Technologies, London, Ontario, Canada N5V 4T7

a r t i c l e i n f o

Article history:Received 23 February 2012Received in revised form 10 April 2012Accepted 12 April 2012Available online 21 April 2012

Keywords:Co-digestionProteinsCarbohydrateMixing ratioHydrolysis coefficient

0960-8524/$ - see front matter � 2012 Elsevier Ltd. Ahttp://dx.doi.org/10.1016/j.biortech.2012.04.052

Abbreviations: %BD, percentage biodegradabilitmethanisation; BSA, bovine serum albumin; C:N, carbammonia; MPRs, methane production rates; SCODdemand; TA, total ammonia; TCOD, total chemicalsuspended solids; TVFAs, total volatile fatty acids; VS⇑ Corresponding author at: Department of Chemica

ing, University of Western Ontario, London, Ontario, C860 3556.

E-mail addresses: sayedbesh@yahoo.com, eelbetrojan.com (E. Elbeshbishy).

a b s t r a c t

Batch anaerobic studies were conducted using five mixtures (M1–M5) of bovine serum albumin (BSA)and starch. The results showed that co-digestion of BSA and starch had a positive impact on the methaneproduction. The highest methane production of 288 mL, the highest methane yield of 360 mL CH4/gCODadded, and the highest maximum methane production rate of 62 mL CH4/d were achieved for M4(20% BSA and 80% starch). Most of the particulate proteins (90%) as well as particulate carbohydrates(95%) were degraded in the first 3 days. The hydrolysis coefficients of particulate proteins and particulatecarbohydrates ranged from 0.65 to 1.01 d�1 and from 0.53 to 1.06 d�1, respectively. The highest methaneproduction was achieved at C:N ratio of 12.8 for M4. For BSA only, propionic acid was the main volatilefatty acid (VFA), while for the starch only, butyric acid was the predominant VFA.

� 2012 Elsevier Ltd. All rights reserved.

1. Introduction

Anaerobic digestion has been proven to be a reliable and eco-nomically feasible technology in full scale operations (Ten Brumm-eler, 2000). Although anaerobic digestion of organic solid wastes isan established technology, it represents on average only 27.5% of allof the biological waste treatment processes used for organic solidwastes (De Baere, 2006). In anaerobic degradation, carbohydratesare first hydrolyzed by enzymes to sugars, which are then degradedby acidogens to volatile fatty acids (VFAs), prior to further conver-sion by acetogens to acetate, carbon dioxide and hydrogen. Lastly,acetate and H2/CO2 are converted by acetoclastic and H2-utilizingmethanogens, respectively, to methane, while proteins are firsthydrolyzed and degraded by proteolytic enzymes into peptidesand individual amino acids (McInerney, 1998). The peptides andamino acids are then acidified into VFAs, hydrogen, ammonium,and reduced sulfur. The VFAs are further converted by acetogensinto acetate and H2/CO2, both of which are lastly converted to meth-ane by methanogens (Fang and Yu, 2002). Acidogens grow faster

ll rights reserved.

y; %M, COD methanised oron to nitrogen ratio; FA, free, soluble chemical oxygenoxygen demand; TSS, total

S, volatile suspended solids.l and Biochemical Engineer-anada N6A 5B9. Tel.: +1 519

shb@uwo.ca, selbeshbishy@

and are less sensitive to pH variation than acetogens and methano-gens, resulting in the accumulation of VFAs and pH decrease, andleading to the suppression of methanogenic activities and, in somecases, even process failure (Batstone et al., 2004). Therefore, it isimportant to have buffering capacity in the system, i.e. productsthat will counteract the effects of the VFAs need also to be formed.It is known that carbohydrate-rich substrates are good producers ofVFAs and those protein-rich substrates yield good buffering capac-ity due to production of ammonia (Mata-Alvarez et al., 2000), thusco-digestion of carbohydrate-rich waste and protein-rich wastes isfeasible. Additionally, co-digestion also allows for digestion ofpoorly biodegradable materials such as fat or protein wastes thatcannot be digested unless mixed with other, more degradablewastes such as carbohydrate-rich wastes (Alatriste-Mondragonet al., 2006).

The main issue for co-digestion process lies in balancing severalparameters in the co-substrate mixture: macro and micronutri-ents, carbon to nitrogen (C:N) ratio, pH, inhibitors/toxic com-pounds, biodegradable organic matter and solids content(Hartmann et al., 2003). Moreover, anaerobic co-digestion can in-crease CH4 production by 50–200%, depending on the operatingconditions and the co-substrates used (Alatriste-Mondragonet al., 2006). Furthermore, in practical terms, the amounts of organ-ic waste generated at particular sites may not be sufficient to makedigestion cost-effective. However, the establishment of a centra-lised facility which co-digested a number of waste products canbe economically viable (Converti et al., 1997).

There are few studies addressing the anaerobic biodegradabilityof mixtures of protein, carbohydrate, and lipids. Breure et al.

Table 1Substrate composition of different mixtures as percentage and amount of BSA andstarch used in each bottle.

Mixture % BSA % Starch Weight of BSA (mg) Weight of starch (mg)

M1 100 0 670 0M2 80 20 536 150M3 50 50 335 375M4 20 80 134 600M5 0 100 0 750

E. Elbeshbishy, G. Nakhla / Bioresource Technology 116 (2012) 170–178 171

(1986) investigated the influence of high concentrations of carbo-hydrate (up to 5 g/L) in the feed medium on the hydrolysis and aci-dogenic fermentation of gelatin (up to 5 g/L) in chemostat cultures.After reaching steady state conditions, a relatively high concentra-tion of glucose or lactose was added to the gelatin-containinggrowth medium as a second substrate. The results showed thatthe degradation of the protein was progressively retarded withthe increase of carbohydrates concentrations in the feed. The re-moval efficiencies of the gelatin were 90% and 77% when gelatinand gelatin plus carbohydrate were fed, respectively. Kuang(2002) investigated the influence of canola oil, starch, and yeast ex-tract using different mixing ratios on methanogenesis in an upflowanaerobic sludge blanket reactor. The results showed that whenthe concentration of starch exceeded that of oil and yeast extractsin co-digestion, methanation enhanced was. Percentage methana-tion of 80.7% and 30% were achieved for mixing ratios of 1:5:1 and1:1:5 of canola oil: starch: yeast extract on COD basis, respectively.In another study, Tommaso et al. (2003) studied the influence ofthe carbohydrate and lipids on anaerobic degradation of bovineserum albumin (BSA) in a horizontal-flow anaerobic immobilizedbiomass reactor fed with BSA based substrates. The results showedthat the initial protein degradation rates were negatively affectedby the presence of other organic compounds such as carbohy-drates. Initial protein degradation rates of 72.8 and 66.4 mg BSA/L.h were observed for the BSA only and for the mixture of BSAand carbohydrate, respectively, although the system that was fedwith protein-plus-carbohydrates showed greater process stability.

It must be asserted that while the aforementioned studies andother literature reports addressed the methanation of mixtures ofcarbohydrates, proteins, and lipids, the understanding of the ki-netic parameters and microbial mechanisms of co-digestion is stilllargely unknown despite the availability of single-substratebiokinetics i.e. anaerobic digestion model (ADM). Thus, a study ofco-digestion of carbohydrates and proteins to find the optimumco-digestion condition is necessary. Also, the relation betweenthe ratio of carbohydrates and proteins and kinetic coefficients ofthe anaerobic digestion needs further examination. Therefore, theprimary objective of this study was to evaluate the performanceof anaerobic co-digestion of BSA and starch in batch processesusing five different mixing ratios.

2. Methods

2.1. Seed and substrate

Anaerobic sludge was collected from the primary anaerobicdigester at St Mary’s wastewater treatment plant (St Mary’s, Ontar-io) and used as seed sludge for all batches used in this study. Thetotal suspended solids (TSS) and volatile suspended solids (VSS)concentrations of the sludge were 12.2 and 9.8 g/L, respectively.The model protein, off-white to beige lyophilized powder bovineserum albumin (CAS 9048-46-8), and the model carbohydrate, so-lid white fine powder starch, (CAS 9005-84-9), were purchasedfrom Sigma Aldrich (Ontario, Canada) and used as substrates.

2.2. Batch anaerobic digestion

Batch anaerobic studies were conducted using BSA and starchas substrates. Five different substrate mixtures of BSA and starch,M1–M5, were used. Table 1 shows the substrate composition offive different mixtures as percentage (on COD basis) and amountof BSA and starch. Because the exact chemical formula of BSA is un-known, the chemical oxygen demand concentration was measuredexperimentally, in triplicate, using six different BSA concentrations(0.25, 0.5, 1, 2, 3, 4, 5, and 10 g/L). A linear relationship between the

BSA concentration and COD value was observed i.e. 1 g BSA wasequivalent to 1.2 g COD. Batch experiments were conducted intriplicates in a series of serum bottles (liquid volume of 200 mland headspace volume of 100 ml). To each bottle, 40 mL of seed,159 mL of deionized water with the required amount of substrateequivalent to 5 g COD/L based on deionized water only (4.2 g ofBSA/L or 4.7 g starch/L), and 1 mL of nutrient stock solution wereadded. Each liter of nutrient stock solution contained 1000 g NaH-CO3, 250 g of K2HPO4, 100 g of MgSO4�7H2O, 10 g of CaCl2�2H2O, 2 gof FeCl2�4H2O, 0.05 g of H3BO3, 0.05 g of ZnCl2, 0.03 g of CuCl2, 0.5 gof MnCl2�4H2O, 0.05 g of (NH4)6Mo7O24, 0.05 g of AlCl3, 0.05 g ofCoCl2�6H2O, and 0.05 g of NiCl2. The initial pH value for the mixedsolution in each bottle was adjusted to 7.2 ± 0.2 using 1 N NaOHand HCl. The headspace was flushed with oxygen-free nitrogengas for a period of 3 min and capped tightly with rubber stoppers.The bottles were then placed in a swirling-action shaker (MaxQ4000, Incubated and Refrigerated Shaker, Thermo Scientific, CA)operating at 180 rpm and maintained at a temperature of 37 �C.Blank bottles were prepared using the sludge with addition ofdeionized water and nutrient stock solution only (no substrateadded). The background methane production from the seed(blank), determined in blank assays with medium and no sub-strate, was subtracted from the methane production obtained inthe substrate assays (Angelidaki et al., 2009). Forty two bottleswere used for each mixture in addition to three for the blank. Threebottles were sacrificed each time for liquid parameter analysis at0.5, 1, 2, 3, 4, 5, 7, 9, 11, 13, 16, 19, and 21 days.

2.3. Analytical methods

Liquid samples were analyzed for TSS, VSS, and alkalinity usingstandard methods (APHA, 1998). Total and soluble chemical oxy-gen demand (TCOD, SCOD) and total ammonia (NH4-N) were mea-sured using HACH methods and test kits (HACH Odyssey DR/2500).Protein was determined by micro-bicinchoninic acid protein assay(Pierce, Rockford, USA) using standard solution of bovine serumalbumin, which was modified from Lowry method (Lowry et al.,1951). Carbohydrate and glucose were determined by the colori-metric method (Dubois et al., 1956) with UV wavelength of490 nm using glucose as standard. Soluble parameters were deter-mined after filtering the samples through 0.45 lm filter paper.Specific surface area was determined by Malvern Mastersizer2000 (version 5.22) laser beam diffraction granulometer. The totalgas volume was measured by releasing the gas pressure in the vialsusing appropriately sized glass syringes (Perfektum, Popper & SonsInc., NY, USA) in the 5–100 mL range to equilibrate with the ambi-ent pressure as recommended by Owen et al. (1979). Biogascomposition including hydrogen, methane, and nitrogen wasdetermined by a gas chromatograph (Model 310, SRI Instruments,Torrance, CA) equipped with a thermal conductivity detector (TCD)and a molecular sieve column (Molesieve 5A, mesh 80/100,6 ft � 1/8 in). The temperatures of the column and the TCD detec-tor were 90 and 105 �C, respectively. Argon was used as the carriergas at a flow rate of 30 mL/min. The concentrations of VFAs wereanalyzed after filtering the sample through 0.45 lm using a gas

(a)

172 E. Elbeshbishy, G. Nakhla / Bioresource Technology 116 (2012) 170–178

chromatograph (Varian 8500, Varian Inc., Toronto, Canada) with aflame ionization detector (FID) equipped with a fused silica column(30 m � 0.32 mm). Helium was used as the carrier gas at a flowrate of 5 mL/min. The temperatures of the column and detectorwere 110 and 250 �C, respectively.

3. Calculations

3.1. Gompertz model

The following modified Gompertz model has been used to de-scribe the progression of cumulative methane production in thebatch tests (Lay et al., 1999):

H ¼ p � exp � expRe

m

pðk� tÞ þ 1

� �� �ð1Þ

where H is the cumulative methane production (mL), P is the max-imum methane production (mL), Rm is the initial methane produc-tion rate (mL/h), k is the lag phase time (h), t is the incubation time(h), and e = exp (1) = 2.718.

3.2. COD methanised (%M)

The percentage of CODmethanised (%M) was calculated by theequations:

%M ¼ ðCODCH4=CODinitialÞ � 100 ð2Þ

(b)

(c)

Fig. 1. Methane production for the five mixtures; (a) cumulative methane, (b)methane production rate, and (c) methane content.

3.3. Anaerobic biodegradability percentage (%BD)

The Anaerobic biodegradability percentage (%BD) was calcu-lated by the equations:

%BD ¼ %Rþ%VFAs ð3Þ

%BD ¼ ½ðCODinitial � CODfinalÞ=CODinitial� � 100

þ ðCODVFA=CODinitialÞ � 100 ð4Þ

where %BD includes the fraction of COD removed and the fraction ofCOD acidified contained in the vial at the conclusion of theexperiment.

3.4. Free ammonia

Free ammonia (FA) concentrations were calculated from thetotal ammonia (TA) concentration in the liquid and the fractionof FA (fN), using the following equation (Omil et al., 1995):

FN ¼ FA=TA ¼ 1=½1þ ðkb � 10�pHÞ=kw� ð5Þ

where kb and kw are the dissociation constants for ammonia andwater, respectively (1.855 � 10�5 and 2.355 � 10�14 mol/L at37 �C).

3.5. COD mass balance

The COD mass balance was calculated by the equation:

%CODbalance ¼ ðCODfinal þ CODCH4Þ=CODinitial � 100 ð6Þ

where CODfinal is the final TCOD at the end of the experiment,CODCH4 is the COD of the cumulative methane produced based on0.395 L CH4/g COD, CODinitial is the initial TCOD.

3.6. Hydrolysis coefficient

Hydrolysis of organic polymers is often described by a first-order kinetic model:

rS ¼ dS=dt ¼ kh � S ð7Þ

where kh is the hydrolysis first-order coefficient and S is the partic-ulate substrate concentration at any time t.

4. Results and discussion

4.1. Methane production

Cumulative methane production, methane production rate, andmethane content are presented in Fig. 1a–c, respectively, error barsare not shown as error was less than 10%. As depicted from Fig. 1a,for all substrate mixtures (M1–M5), a rapid initial methane pro-duction was observed. As presented in Table 2 based on Gompertzmodel (Eq. 1), the lag phases ranged from 0.8 to 1.3 day. The lagphase of BSA only (M1) was higher than the lag phase of starchonly (M5), 1.3 d versus 0.8 d. That was expected as the hydrolysisof proteins is slower than the hydrolysis of carbohydrates (Pavlo-stathis and Giraldo-Gomez, 1991).

The highest cumulative methane production of 288 mL wasachieved for M4 (20% BSA plus 80% starch) compared to 197 mLfor M1 (BSA only) and 252 ml for M5 (starch only). It should bestressed here that co-digestion of BSA and starch resulted in higher

Table 2Gompertz model, methane yield, and maximum methane production rate (MMPR).

Mixture Gompertz model Methane yield MMPR

P Rm k R2

mL mL/d d mL CH4/g CODadded mL/d

M1 197 35 1.3 0.99 246 41M2 219 37 1.1 0.99 273 40M3 239 42 0.9 0.99 298 44M4 288 49 0.8 0.99 360 62M5 252 46 0.8 0.99 315 44

(a)

(b)

Fig. 2. Proteins concentrations for the five substrate mixtures; (a) particulateprotein, (b) soluble protein.

E. Elbeshbishy, G. Nakhla / Bioresource Technology 116 (2012) 170–178 173

methane production compared to BSA only (46% higher) or starchonly (11% higher). Based on the BSA and starch fractions in themixtures and the methane produced from M1 (197 mL) and M5(252 mL), the calculated methane production of M2, M3, and M4were expected to be 209, 226, 243 mL. The experimental cumula-tive methane production from M2, M3, and M4 were 219, 239,and 288 mL, respectively, which corresponded to 5%, 6%, and 18%higher than those calculated methane (based on the methane pro-duction from M1and from M5), respectively. Thus, it is obviousthat co-digestion of BSA and starch had a positive impact on themethane production as the methane production was higher thanthe sum of the methane production from each fraction.

The methane yield was normalized per substrate mass CODadded (mL CH4/g CODadded). As depicted in Table 2, the highestmethane yield of 360 mL CH4/g CODadded was achieved for M4.The lowest methane yield of 246 mL CH4/g CODadded was observedfor BSA only (M1). The methane yields presented in Table 2 empha-sized the positive effect of adding starch to BSA. The methaneyields (from M1 to M4) increased with increasing the starchpercentage, the maximum increase of 36% was achieved (M4 com-pare to M1). On the other hand, a methane yield of 315 mL CH4/gCODadded was observed for M5 (starch only). Adding BSA to thestarch had a positive effect only at M4 (20% BSA plus 80% starch),after which a negative effect was observed and reflected by a de-crease in the methane yield by 5% for M3 (50% BSA plus 50% starch)and 13% for M2 (80% BSA plus 20% starch) compared to the meth-ane yield of the starch only (M5).

Fig. 1b shows the methane production rates (MPRs) of the fivesubstrate mixtures (M1–M5). The MPRs increased sharply duringthe first two days, then fluctuated (from day 2 to day 6), prior todeclining sharply (from day 6 to day 10), and subsequently declin-ing gradually before stabilization at close to zero after about18 days. The average methane production rates during days 2 to6 (fluctuation period) were 29, 30, 34, 40, and 40 mL CH4/d for mix-tures M1 to M5, respectively. The highest maximum methane pro-duction rate of 62 mL CH4/d was obtained for M4, while themaximum methane production rates for the other mixtures (M1,M2, M3, and M5) ranged from 40 to 44 mL CH4/d, Table 2.

As apparent from Fig. 1c, the average methane contents of thebiogas produced were calculated based on the last 16 days (exclud-ing the data in the first 5 days) of the experiments. The highestaverage methane of 57% was observed for M1, while the averagemethane contents of the other mixtures (M2–M5) were contentwere ranged between 50% and 52%. The methane content rangedfrom 50% to 57% observed in this study is in agreement with therange of 50–60% which is normally obtained from conventionalanaerobic digestion of organic waste (Samani et al., 2001).

The initial methane production rates (the initial slope of theexponential phase of the cumulative methane production withtime) obtained from the Gompertz model are shown in Table 2.The data obtained from the Gompertz model emphasize that M4had the highest methane production rate of 49 mL/d. The initialmethane production rate of M4 was 32% higher than that of M1

(35 mL/d) and 7% higher than that of M5 (46 mL/d). On the otherhand, although the maximum methane production rate of M5was 7% higher than that of M1, the initial methane production ratewas 31% higher than that of M1.

4.2. Substrate degradation

Fig. 2a and b show the particulate proteins degradation and thesoluble proteins concentrations for the different substrate mix-tures (M1–M5). As evident from Fig. 2a, for all substrate mixtures,a sharp decline in particulate proteins was observed from thebeginning. Most of the particulate proteins (more than 90%) weredegraded in the first 3 days with final particulate proteins concen-trations of about 330 mg/L in M2, M3, and M4 and about 470 mg/Lin M1. The particulate proteins content of the seed biomass at thebeginning of the experiments was 0.16 mg particulate proteins/mgVSS. The final VSS concentrations at the end of the batches rangedfrom 1920 to 2280 mg/L. The particulate proteins to biomass ratioat the end of the experiments were 0.22, 0.15, 0.15, 0.14, and 0.16for M1–M5, respectively, thus emphasizing that all the particulateproteins were degraded in M2–M5 and about 96% of the particu-late proteins were degraded in M1.

As depicted in Fig. 2b, the soluble proteins profiles started withan exponential increase in the first 2 days followed by a sharp de-crease till leveling off at about 10% of the maximum values after9 days. Soluble proteins concentrations are impacted by two rates,production rate from particulate proteins and utilization rate tovolatile fatty acids and ammonia. At the beginning, the productionrate of soluble proteins was higher than the utilization rate whichis reflected by the exponential phase at the beginning of the exper-iments. The peak soluble proteins concentrations of 2200, 1940,1440, and 870 mg/L were observed for M1–M4, respectively. Theratios of maximum soluble proteins to initial particulate proteinsranged between 0.54 and 0.56. This very close range emphasizedthe direct relation between the initial particulate proteins and

(a)

(b)

Fig. 3. Carbohydrate concentrations for the five substrate mixtures; (a) particulatecarbohydrate, (b) soluble carbohydrate.

174 E. Elbeshbishy, G. Nakhla / Bioresource Technology 116 (2012) 170–178

the maximum soluble proteins concentrations. Moreover, thedecline phase of soluble proteins started when the solubleproteins concentration reached about 55% of the initial particulateproteins concentration regardless of the initial particulate proteinsconcentration.

Fig. 3a and b show the particulate and dissolved carbohydratesprofiles. As shown in Fig. 3a, particulate carbohydrates degradationfollowed the same trend of particulate proteins (sharp decreasefrom the beginning and reached almost a plateau after 3 days. Ittook 2 days to degrade more than 95% of the particulate carbohy-drates for M4 and M5 and 3 days for M2 and M3. Furthermoreand given that the final particulate carbohydrate to final biomassratios were less than 12% (which were approximately the same ra-tios in the seed only), thus all the particulate carbohydrates weredegraded. Fig. 3b showed the dissolved carbohydrates profiles forthe five mixtures. As illustrated, the dissolved carbohydrates deg-radation followed the same trend of soluble proteins. The maxi-mum dissolved carbohydrates concentrations of 500, 1320, 2100,and 3100 mg/L were observed for M2–M5 respectively, which cor-respond to 52%, 66%, 67%, and 78% of the initial particulatecarbohydrates.

Hydrolysis coefficients of the particulate proteins and particu-late carbohydrates are presented in Table 3. As shown in the Table,the hydrolysis coefficients (calculated in accordance with Eq. 3) ofparticulate proteins (BSA) in the four mixtures, M1–M4, rangedfrom 0.65 d�1 (M1) to 1.01 d�1 (M4), while the hydrolysis coeffi-cients of the particulate carbohydrate (starch) in the four mixtures,M2–M5, ranged from 0.53 d�1 (M2) to 1.06 d�1 (M4). The hydroly-sis coefficient of starch only (M5) of 0.78 d�1 was 20% higher thanthat of BSA only (M1). The hydrolysis coefficients of the starch andthe BSA obtained in this study were within the ranges reported byGarcia-Heras (2003) of 0.5–2 and 0.25–0.8 d�1 for carbohydrateand proteins, respectively. Moreover, the hydrolysis coefficient ofBSA of 0.65 d�1 was very close to the 0.6 d�1, obtained by Gavala

et al. (2003) for the experimental data of Nagase and Matsuo(1982) on gelatin. It is obvious from Table 3 that adding starch toBSA enhanced the hydrolysis coefficient of the BSA at all mixing ra-tios, peaking at 1.01 d�1 for M4 (20% BSA plus 80% starch). Basedon the specific surface areas of starch and BSA of 0.152 and0.00677 m2/g, respectively and the masses used in the batches,the estimated surface areas for M1–M5 were 0.005, 0.024, 0.053,0.082, and 0.102 m2. It is thus evident that since hydrolysis is a sur-face area phenomenon, the observed hydrolysis coefficient in-creased with the carbohydrate content. On the other hand,adding BSA to starch had a positive effect on the hydrolysis coeffi-cient of starch only when the BSA was 25% of the starch otherwiseit decreased the hydrolysis coefficient. The maximum hydrolysiscoefficient of starch of 1.06 d�1 was achieved at M4. Moreover,mixtures containing BSA equal or greater than starch had a nega-tive effect on the hydrolysis coefficient of the starch, reducing itfrom 0.78 d�1 for M5 to 0.53 and 0.55 d�1 for M2 (80% BSA plus20% starch) and M3 (50% BSA plus 50% starch), respectively.

Methanisation (%M) and biodegradability (%BD) calculatedusing Eqs. 2 and 3 are presented in Table 3. The difference betweenmethanisation and biodegradability is the COD of the final TVFAs.The highest methanisation and the highest biodegradability of91% and 93%, respectively, were observed for M4. The very closevalues of methanisation and biodegradability of M4 emphasizedthat there was no inhibition of methanogenesis. Moreover, the fi-nal TVFAs concentrations of 60 mg COD/L also confirmed that.The lowest methanisation and biodegradability of 62% and 76%,respectively, were observed for M1 (BSA only). Methanisationand biodegradability of 80% and 84% were achieved for M5 (starchonly). The differences between biodegradability and methanisationof 21% and 13% for M1 and M2, respectively, emphasized methano-genesis inhibition in those two mixtures. On the other hand, forM3–M5, the differences between biodegradability and methanisa-tion were less than 5% proving that there was no inhibition.

The TCOD destructions were calculated based on the initialTCOD, final TCOD, and the TCOD destructions of 43%, 48%, 52%,59%, and 53% were obtained for M1–M5, respectively. The CODmass balances for all the batches computed considering the initialand final TCOD, and the equivalent COD of methane (0.395 L CH4/gCOD), indicated a closure of 91–93%, thus emphasizing datareliability.

4.3. Ammonia released and free ammonia

The main products of the biodegradation of proteins in anaero-bic conditions are ammonia and various amino acid compounds.However, the ammonia produced may be toxic for methanogenicbacteria (Soubes et al., 1994). It is generally believed that ammoniaconcentrations below 200 mg/L are beneficial to anaerobic processsince nitrogen is an essential nutrient for anaerobic microorgan-isms (Liu and Sung, 2002). A wide range of inhibiting ammoniaconcentrations has been reported in the literature, with the inhib-itory ammonia concentration that caused a 50% reduction in meth-ane production ranging from 1.7 to 14 g/L (Chen et al., 2008). Freeammonia (FA) has been suggested as the main cause of inhibitionsince it is freely membrane-permeable (De Baere et al., 1984).Among the four types of anaerobic microorganisms, the methano-gens are the least tolerant and the most likely to cease growth dueto ammonia inhibition (Kayhanian, 1994). The concentrations of FAfrom 100 to 140 mg/L inhibit mesophilic treatment (De Baere et al.,1984), while according to Omil et al. (1995) concentrations of FAfrom 25 to 140 mg/L inhibit mesophilic treatment.

Fig. 4a and b show the ammonia concentrations (NH4–N) andcalculated free ammonia (NH3–N) for the five mixtures. As de-picted from Fig. 4a, the highest ammonia concentration of860 mg/L was observed for M1 (BSA only). The lowest ammonia

Table 3Methanisation, biodegradability, kh of BSA and kh starch, ammonia yield, and C:N ratio.

Mixture Methanisation (% M) Biodegradability (% BD) kh of BSAa kh of starchb Ammonia yield C:N ratio

% % d�1 d�1 mg NH4/mg total proteinsc mg C/mg N

M1 62 76 0.65 – 0.21 3.5M2 69 78 0.86 0.53 0.18 4.2M3 75 79 0.84 0.55 0.20 6.3M4 91 93 1.01 1.06 0.20 12.8M5 80 84 – 0.78 0.37 48.5

a R2 of the linear fit used to derive hydrolysis coefficients of BSA ranged from 0.96 to 0.99.b R2 of the linear fit used to derive hydrolysis coefficients of starch ranged from 0.86 to 0.92.c Total proteins = particulate proteins + soluble proteins.

(a)

(b)

(c)

Fig. 4. Total ammonia, free ammonia, and pH for the five substrate mixtures.

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176 E. Elbeshbishy, G. Nakhla / Bioresource Technology 116 (2012) 170–178

concentration of 180 mg/L was observed in M5 (starch only). Theammonia concentration of M1 was about 23 folds of M4. Ammoniayield was calculated as mg ammonia produced (highest ammoniaconcentration minus initial ammonia concentration) per mg initialtotal protein (initial particulate proteins plus initial soluble pro-teins), Table 3. The ammonia yields for M1–M4 were very close,ranging from 0.18 to 0.21 mg NH4/mg total proteins emphasizingthe direct relation between the initial amount of proteins and theproduced ammonia. On the other hand, a decrease in ammoniaconcentration was observed at the end of the experiments forM1, M2, and M3. This decrease might be due to precipitation ofammonia as struvite (MgNH4PO4�6H2O) (Mamals et al., 1994).Fig. 4b showing FA concentrations, which were computed usingEq. (6), indicates that the highest concentrations of FA of 20 mg/Lin M1 was below the methanogenic bacteria inhibition level of25–100 mg/L reported above. The maximum free ammonia con-centrations of 16, 13, 5, and 2 mg/L were observed for M2–M5,respectively.

Fig. 4c shows the change in the pH during the experiments. Theinitial pH ranged between 7.2 and 7.3 while the final pH rangedfrom 6.85 to 7.33. Anderson and Yang (1992) reported a pH rangeof 6.4–7.6 in a normal functioning digester, beyond which a state ofinhibition may occur resulting from toxic effects of the hydrogenions which are believed to be closely related to the accumulationof VFAs. For all mixtures, a decrease in the pH was observed duringthe first 3 days followed by an increase to day 10, after which,there was no significant change in the pH. The initial decrease in

Fig. 5. TVFAs and individual VFA

the pH could be due to the rapid VFAs production at the beginning(as shown later), while the increase in the pH from day 3 to day 10could be due to generation of NH4–N during protein degradation,as ammonia which is a base combines with carbon dioxide andwater to form ammonium bicarbonate, a natural pH buffer (Zhanget al., 2005).

4.4. N ratio

Optimum values of C:N and COD:N ratios of 20 and 70, respec-tively, have been suggested for the stable performance of anaerobicdigestion (Chen et al., 2008). However, lower values of C:N ratiosbetween 6 and 9 have been reported as suitable for the anaerobicdigestion of nitrogen-rich waste (Mshandete et al., 2004). Further-more, Itodo and Awulu (1999) reported successful anaerobic batchdigestion for poultry, cattle, and piggery wastes slurries with C:Nratios of 6:1 and 9:1. The C:N ratios were calculated based onthe carbon and nitrogen content of the BSA, starch, and the seed.The chemical formulae of (C4H6.1O1.2N), (C6H12O5), and (C5H7O2N)were used for BSA, starch, and seed, respectively. The calculatedinitial C:N ratios were 3.5, 4.2, 6.3, 12.8 and 48.5 for M1–M5,respectively, Table 3. As the minimum suitable C:N ratio reportedin the literature was 6, that might be one of the factors affectingthe partial inhibition of the methanogens in M1, M2, M3. On theother hand, the high value of C:N of 48.5 seems to have negativeeffect on the methane production. It must be noted that M5

profiles of the five mixtures.

E. Elbeshbishy, G. Nakhla / Bioresource Technology 116 (2012) 170–178 177

contained only starch and thus the nitrogen utilized for biomassgrowth dissociated with substrate utilization originated only fromthe seed material and hence it is plausible that nitrogen availabilitymay have hindered methane production. The highest methaneyield in this study was achieved at C:N ratio of 12.8 at M4. It mustbe emphasized that in all five mixtures, the ammonia nitrogen con-centration throughout the experiment ranged from 80 to 860 mg/Lwell below the potential ionized ammonia inhibitory level of about1700 mg/L (Chen et al., 2008).

4.5. Volatile fatty acids (VFAs)

Carbohydrates are known to be easily and rapidly converted viahydrolysis to simple sugars and subsequently fermented to VFAs.Protein is hydrolyzed to amino acids and further degraded to VFAs(McInerney, 1998). If a high concentration of VFAs is formed, pHwill decrease potentially to levels which inhibit and eventually killthe methanogens (Mshandete et al., 2004). Moreover, VFAs canaccumulate in the biomass boundary layer and induce local pH val-ues that inhibit the process (Vavilin et al., 2008). Fig. 5a–f showsthe TVFAs profile as well as the acetate, propionate, butyrate, andvalerate of the five mixtures. As shown in Fig. 5a, the TVFAs con-centrations of the five mixtures increased almost linearly initially(first 2 days) reaching peak concentrations between day 2 andday 3, after which a considerable decline was observed prior to sta-bilization in the last 10 days of the experiments. During the peakperiod (day 2 to day 3), the highest cumulative TVFAs of1250 mg COD/L was observed for M5 followed by 1150 mg COD/L for M1, while the lowest cumulative TVFAs of 780 mg COD/Lwas observed for M3. The final TVFAs concentrations in all exper-iments ranged from 60 to 530 mg COD/L. The lowest residualTVFAs of 60 mg COD/L was achieved for M4 and the highest resid-ual TFVAs of 530 mg COD/L was observed for M1 (BSA only). Thefinal TVFAs obtained correlated linearly (not shown, R2 of 0.961)with the initial concentration of the proteins. For M1–M4, the finalTVFAs increased with increasing the initial protein concentration.That is might be due to the accumulation of the propionic acid pro-duced via fermentation of protein for M1 to M3 and accumulationof propionic and butyric acids in M4. As shown in Fig. 5b–f, therewas accumulation of propionic acid in all batches except in M4and M5. Moreover, the fraction of propionic acid in the final TVFAsfor M1, M2, and M3 ranged from 73% to 86%, due to slower meth-anogenic degradation of propionate compared with the acetate andbutyrate. This observation agrees with the study by Zhang et al.(2005) who reported an accumulation of propionic acid in batchanaerobic treatment of kitchen wastes.

The profiles of acetate, propionate, butyrate, and valerate areshown in Fig. 5b–f. Similar trends to TVFAs (initial increase fol-lowed by decline and stabilization at low concentrations of lessthan 120 mg COD/L) were observed in all mixtures. As illustratedin Fig. 5b, for M1 (BSA only), the main VFA was propionic acid(ranging from 39% to 73% on a COD basis) followed by acetic acid(ranging from 14% to 40%). The butyric and valeric acids were lessthan 15% of the TVFAs concentration. On the other hand as de-picted in Fig. 5f, for M5 (starch only), the main VFA was butyricacid (ranging from 55% to 75%) followed by acetic acid (rangedfrom 11% to 46%). The propionic and valeric acids were below20% and 10% of the TVFAs concentration, respectively. As shownin Fig. 5b–f, a relatively high residual acetic acid (80–100 mgCOD/L) in batches for M1, M2, and M5 compared to <35 mg COD/L for M3 and M4.

5. Conclusions

The outcome of this study emphatically revealed the benefits ofmixing carbohydrate-rich with protein-rich wastes. The results of

this study revealed that mixing ratio of 20% BSA and 80% starchwas superior to all other mixtures with respect to methane pro-duction, methane yield, and methane production rate. The highestmethane yield in this study was achieved at C:N ratio of 12.8 forM4. The Hydrolysis first-order rate coefficients of 0.78 and0.65 d�1 observed for starch only and BSA only, respectively, andincreased to 1.06 and 1.01 d�1, respectively, when 20% BSA mixedwith 80% starch.

Acknowledgements

Authors want to thank Trojan Technologies and the EgyptianMinistry of Higher Education for their financial contributionstowards this study.

References

Alatriste-Mondragon, F., Samar, P., Cox, H.H., Ahring, B.K., Iranpour, R., 2006.Anaerobic codigestion of municipal, farm, and industrial organic wastes: asurvey of recent literature. Water Environ. Res. 78, 607–636.

Anderson, G.K., Yang, G., 1992. Determination of bicarbonate and total volatile acidconcentration in anaerobic digesters using a simple titration. Water Environ.Res. 64, 53–59.

Angelidaki, I., Alves, M., Bolzonella, D., Borzacconi, L., Campos, J.L., Guwy, A.J., 2009.Defining the biomethane potential (BMP) of solid organic wastes and energycrops: a proposed protocol for batch assays. Water Sci. Technol. 59 (5), 927–934.

APHA, 1998. Standard Methods for the Examination of Water and Wastewater, (20ed). American Public Health Association, Washington DC, USA.

Batstone, D.J., Keller, J., Blackall, L.L., 2004. The influence of substrate kinetics on themicrobial community structure in granular anaerobic biomass. Water Res. 38,1390–1404.

Breure, A.M., Mooijman, K.A., van Andel, J.G., 1986. Protein degradation in anaerobicdigestion: influence of volatile fatty acids and carbohydrates on hydrolysis andacidogenic fermentation of gelatine. Appl. Microbiol. Biotechnol. 24, 426–431.

Chen, Y., Cheng, J.J., Creamer, K.S., 2008. Inhibition of anaerobic digestion process: areview. Bioresour. Technol. 99, 4044–4064.

Converti, A., Drago, F., Ghiazza, G., del Borghi, M., Macchiavello, A., 1997. Co-digestion of municipal sewage sludges and pre-hydrolysed woody agriculturalwastes. J. Chem. Tech. Biotech. 69, 231–239.

De Baere, L., 2006. Will anaerobic digestion of solid waste survive in the future?Water Sci. Technol. 53, 187–194.

De Baere, L.A., Devocht, M., Van Assche, P., Verstraete, W., 1984. Influence of highNaCl and NH4Cl salt levels on methanogenic associations. Water Res. 18, 543–548.

Dubois, M., Gilles, K.A., Hamilton, J.K., Rebers, P.A., Smith, F., 1956. Colorimetricmethod for determination of sugars and related substances. Anal. Chem. 28,350–356.

Fang, H.H.H., Yu, H.Q., 2002. Mesophilic acidification of gelatinaceous wastewater. J.Biotechnol. 93, 99–108.

Garcia-Heras, J.L., 2003. Reactor sizing, process kinetics and modelling of anaerobicdigestion of complex wastes. In: Mata-Alvarez, J. (Ed.), Biomethanization of theOrganic Fraction of Municipal Solid Wastes. IWA Publishing, TJ InternationalLtd., Padstow, Cornwall, UK, pp. 21–62.

Gavala, H.N., Angelidaki, I., Ahring, B.K., 2003. Kinetics and modelling of anaerobicdigestion process. Adv. Biochem. Eng. Biotechnol. 81, 57–93.

Hartmann, H., Angelidaki, I., Arhing, B.K., 2003. Co-digestion of the organic fractionof municipal waste with other waste types. In: Mata-Alvarez, J. (Ed.),Biomethanization of the Organic Fraction of Municipal Solid Wastes. IWAPublishing, TJ International Ltd., Padstow, Cornwall, UK, pp. 181–200.

Itodo, I., Awulu, J.O., 1999. Effects of total solids concentrations of poultry, cattle andpiggery waste slurries on biogas yield. Trans ASAE 42, 1853–1855.

Kayhanian, M., 1994. Performance of a high-solids anaerobic digestion processunder various ammonia concentrations. J. Chem. Tech. Biotechnol. 59, 349–352.

Kuang Y., 2002. Enhancing Anaerobic Degradation of Lipids in Wastewater byAddition of Co-substrate. Ph.D. Thesis, School of Environmental Science,Murdoch University, Australia.

Lay, J.J., Lee, Y.J., Noike, T., 1999. Feasibility of biological hydrogen production fromorganic fraction of municipal solid waste. Water Res. 33, 2579–2586.

Liu, T., Sung, S., 2002. Ammonia inhibition on thermophilic aceticlasticmethanogens. Water Sci. Technol. 45, 113–120.

Lowry, O.H., Rosebrough, N.J., Farr, A.L., Randall, R.J., 1951. Protein measurementswith the folin phenol reagent. J. Biol. Chem. 193, 265–275.

Mamals, D., Pitt, P.A., Cheng, Y.W., Loianco, J., Jenkins, D., 1994. Determination offerric chloride dose to control struvite precipitation in anaerobic sludgedigesters. Water Environ. Res. 66, 912–918.

Mata-Alvarez, J., Mace, S., Llabres, P., 2000. Anaerobic digestion of organic wastes.An overview of research achievements and perspectives. Bioresour. Technol. 74,3–16.

178 E. Elbeshbishy, G. Nakhla / Bioresource Technology 116 (2012) 170–178

McInerney, M.J., 1998. Anaerobic hydrolysis of fats and proteins. In: Zehnder, A.J.B.(Ed.), Biology of Anaerobic Micro-Organisms. John Wiley and Sons, New York,pp. 373–415.

Mshandete, A., Kivaisi, A., Rubindamayugi, M., Mattiasson, B., 2004. Anaerobic batchco-digestion of sisal pulp and fish wastes. Bioresour. Technol. 95, 19–24.

Nagase, M., Matsuo, T., 1982. Interactions between amino-acids-degrading bacteriaand methanogenic bacteria in anaerobic digestion. Biotechnol. Bioeng. 24,2227–2239.

Omil, F., Mendez, R., Lema, J.M., 1995. Anaerobic treatment of saline wastewatersunder high sulphide and ammonia content. Bioresour. Technol. 54, 269–278.

Owen, W.F., Stuckey, D.C., Healy, J.B., Young, L.Y., McCarty, P.L., 1979. Bioassay formonitoring biochemical methane potential and anaerobic toxicity. Water Res.13, 485–492.

Pavlostathis, S.G., Giraldo-Gomez, E., 1991. Kinetics of Anaerobic Treatment. WaterSci. Technol. 24, 35–59.

Samani Z., Yu W.H., Hanson A., 2001. Energy production from segregated municipalwaste and agricultural waste using bi-phasic anaerobic digestion. In:

VanVelsen, A.F.M., Verstraete W.H., (Eds.), Anaerobic Conversion forSustainability. Proc. 9th World Congress Anaerobic Digestion, Antwerpen,Belgium, pp. 195–197.

Soubes, M., Muxõ, L., Fernandez, A., Tarlera, S., Quirolo, M., 1994. Inhibition ofmethanogenesis from acetate by Cr+3 and ammonia. Biotechnol. Lett. 16, 195–200.

Ten Brummeler, E., 2000. Full scale experience with the Biocel process. Water Sci.Technol. 41, 299–304.

Tommaso, G., Ribeiro, R., Varesche, M.B.A., Zaiat, M., Foresti, E., 2003. Influence ofmultiple substrates on anaerobic protein degradation in a packed-bedbioreactor. Water Sci. Technol. 48, 23–31.

Vavilin, V.A., Fernandez, B., Palatsi, J., Flotats, X., 2008. Hydrolysis kinetics inanaerobic degradation of particulate organic material: an overview. WasteManagement 28, 939–951.

Zhang, B., Zhang, L.L., Zhang, S.C., Shi, H.Z., Cai, W.M., 2005. The influence of pH onhydrolysis and acidogenesis of kitchen wastes in two-phase anaerobicdigestion. Environ. Technol. 26, 329–340.