A promising strategy for nutrient recovery using heterotrophic indigenous microflora from liquid biogas digestate

Science of the Total Environment 690 (2019) 492–501

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

A promising strategy for nutrient recovery using heterotrophic indigenous microflora from liquid biogas digestate Mostafa Sobhi a, Jianbin Guo a,⁎, Xian Cui a, Hui Sun a, Bowen Li a, Dominic Aboagye a, Ghulam Mustafa Shah b, Renjie Dong a,c a b c

College of Engineering (Key Laboratory for Clean Renewable Energy Utilization Technology, Ministry of Agriculture), China Agricultural University, Beijing 100083, PR China Department of Environmental Sciences, COMSATS University Islamabad, Vehari Campus, Vehari, 61100, Pakistan Yantai Institute, China Agricultural University, Yantai 264032, Shandong, PR China

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Heterotrophic indigenous microflora grew in undiluted liquid digestate successfully. • Phosphate & total nitrogen recovery were 68% & 19%, total nitrogen removal was 97%. • N6 g L−1 of dry biomass was simultaneously produced without pathogens. • Biomass features with 65% crude protein, 10.9% lipids and 19.6 MJ kg−1 gross energy • Integrated production of microbial biomass and ammonium salts is recommended.

a r t i c l e

i n f o

Article history: Received 31 May 2019 Received in revised form 24 June 2019 Accepted 28 June 2019 Available online 02 July 2019 Editor: Huu Hao Ngo Keywords: Nutrient recovery Microbial biomass Liquid digestate Indigenous microflora Aerobic heterotrophic mode

a b s t r a c t Nutrient overloading resulting from digestate (effluent of anaerobic digestion process) application has become a major bottleneck for the development of the biogas industry and raised environmental concerns in regions with intensive animal husbandry. Due to this, it is imperative to find low cost and effective alternative to export nutrient from digestate. Among the numerous applications, indigenous microflora has recently been utilized successfully as a biofloc technology in aquatic systems for controlling ammonia and subsequent reduction of feeding cost. Accordingly, performance of the indigenous microflora in undiluted liquid digestate of chicken manure was evaluated in this study to recover nutrients and produce high-value biomass under aerobic heterotrophic mode in batch shaking experiments. The results showed that 68% of phosphate was recovered and 97% of total nitrogen was removed from the liquid digestate. Additionally, N6 g L−1 of dry biomass was simultaneously produced and featured with up to 65% crude protein without pathogens, 10.9% lipids, 10.7% ash and 19.6 MJ kg−1 gross energy. Therefore, the produced biomass could be used either as an alternative sustainable source for animal or fish feeding or as a substrate for energy applications. © 2019 Elsevier B.V. All rights reserved.

Abbreviations: C/N, Carbon to nitrogen ratio; COD, Chemical oxygen demand; DM, Dry matter; DNA, Deoxyribonucleic acid; G, Glycerol; GCE, Gross carbon efficiency; HHV, Higher − heat value “gross energy content”; HTL, Hydrothermal liquefaction; LD, Liquid digestate; NO− 2 -N, Nitrite nitrogen; NO3 -N, Nitrate nitrogen; PAO, Polyphosphate-Accumulating Organism; PO3− 4 -P, Phosphate phosphorus; S, Sucrose; TAN, Total ammonia nitrogen; TKN, Total Kjeldahl nitrogen; TN, Total nitrogen; VFA, Volatile fatty acid. ⁎ Corresponding author at: College of Engineering, China Agricultural University, Beijing 100083, China. E-mail address: [email protected] (J. Guo).

https://doi.org/10.1016/j.scitotenv.2019.06.487 0048-9697/© 2019 Elsevier B.V. All rights reserved.

M. Sobhi et al. / Science of the Total Environment 690 (2019) 492–501

1. Introduction Biogas industry is one of the main bioenergy applications that is growing quickly for recovering energy and mitigating the environmental pollution from organic wastes. Globally, utilization of biogas as an energy resource has increased by 56% during the last six years from 0.84 EJ in 2010 to 1.31 EJ in 2016 (WBA, 2017). However, digestate utilization has become a major bottleneck for expansion of the biogas industry in future (Xia and Murphy, 2016). It contains nitrogen, phosphorus, potassium and other minerals present in the original substrate (Khagendra et al., 2017). Direct fertilizing of substrate on soils is considered as a main application (José et al., 2012) but it is usually limited by seasonal application and finite arable land (Yun et al., 2018). Otherwise, environmental risks may occur due to the nutrient overloading. Therefore, the European Nitrates Directive (91/676/EEC) regulates the application of organic materials as fertilizers for agricultural land according to nitrogen content, location, and crops demands (Drekeke et al., 2010). Also, phosphorus is a finite non-renewable element which is subjected to an open cycle with heavy losses to the environment (Marlous, 2019) and identification of ways to close this cycle has become very necessary (Marilys and Lynda, 2019). Thus, the challenges of digestate utilization as reported by Mohanakrishnan and Chettiyappan (2019) include the identification of ways of increasing its value, creating new markets for it, reducing the dependence on land applications, ensuring more secure and sustainable treatment methods, and potentially reducing the operating cost of the facility. To overcome these challenges, attempts have been carried out to find some alternatives. Recently, microalgae have been utilized due to their ability to recover nutrients efficiently from digestate (Xia and Murphy, 2016). In addition, microalgae are able to produce high added value and sustainable products based feed (Moheimani et al., 2018), energy (Ji et al., 2015) and biochemical (Meixner et al., 2016). Although liquid digestate contains nutrients that algae needs, its high ammonia concentrations and turbidity inhibit algae growth. Dilution of the digestates in high ratios (up to 20 times) will be enough to solve this problem (Ji et al., 2015; Xia and Murphy, 2016). However, this action for large scale microalgae production is not so practical due to the large water consumption, hence making microalgae less attractive for applying on large industrial scales (Xia and Murphy, 2016). Moreover, single strain cultivation of microalgae for different specific applications often requires pre-cultivated strains and pre-treated digestate by filtration, dilution and sterilization (Xia and Murphy, 2016), which increases the total cost and makes the implementation inapplicable on the large scale. On the other hand, the concept of cultivating indigenous microflora for nutrient recovery, wastes treatment, and biomass production has recently been proven as an effective technology in some applications by different researchers (Cerda et al., 2019). One of these applications is called biofloc technology. It is an aquaculture system, which can reduce ammonia and other nutrients from fish water in a form of microbial biomass by adding organic carbon to the water of aquaculture in the presence of aeration and mixing (Avnimelech and Kochba, 2009). Cultured fishes can feed on the produced biomass which was proved to be beneficial for fish health (Azim and Little, 2008; Carmen and Mejía, 2017). Another application of indigenous microflora was to produce high value-added products (hydrolytic enzymes, cellulases and proteases) from solid state fermentation of digestate (Cerda et al., 2019). Although small amounts of cellulase and protease could be extracted after the fermentation, the concept of using indigenous microflora has sounded interesting due to its relatively simple requirements and executable at large scales (Cerda et al., 2019). In the light of biofloc technology applied in aquatic systems, we hypothesized that high-value microbial biomass production and nutrients recovery could be achieved simultaneously by the indigenous microflora cultivated with undiluted liquid digestate (LD). The objectives of this study were to (i) evaluate the ability of the indigenous

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heterotrophic microorganisms to grow on high strength LD, (ii) determine the optimal conditions for cultivation indigenous heterotrophic microorganisms, (iii) determine the efficiencies for nutrients recovery and wastewater treatment, (iv) determine the characteristics of the generated microbial biomass, and (v) identify their microorganisms by deoxyribonucleic acid (DNA) identification analysis. The utilization of the produced biomass was also discussed in this study. 2. Material and methods 2.1. Digestate origin, preparation, and specificities The digestate used in this study was collected from a continuous stirring tank reactor pilot-scale anaerobic digester, which locates at college of engineering, China agricultural university. Digestion temperature was 37 °C treating chicken manure of 15% TS, which run stably at hydraulic retention time of 40 days. After collection from the digester, raw digestate was centrifuged at 10000 rpm (10,610 ×g) for 20 min to separate solid and liquid fractions. Characteristics of the collected LD are listed in Table 1. 2.2. Experimental set-up and treatments Fifty mL of LD were transferred into 250 mL conical flasks, incubated at 28 °C and 150 rpm for 72 h in a shaking incubator. Indigenous strains grew under heterotrophic conditions without addition of external microbes as same as biofloc technology (Avnimelech and Kochba, 2009) and other similar applications. Since digestate contained all macro and micro-nutrients needed by heterotrophic microorganisms except organic carbon (Xia and Murphy, 2016), only organic carbon was added to the mediums. All treatments were at least triplicated. Samples of 5 mL were collected at the beginning of the experiment, after 24 h, after 48 h, and after 72 h. Then, they were centrifuged as mentioned later after which the supernatant was used for water quality analyses while solid sediment was used for evaluation of growth criteria and biomass characteristics. Performance of indigenous microflora was studied in two laboratory experiments. In experiment 1, the ability of indigenous microflora for biomass production, nutrients recovery and treatment of the high strength LD with different amounts of organic carbon were evaluated. In this experiment, the simple source of sucrose (laboratory grade, Sinopharm Chemical Reagent Co, Ltd) was used as an organic carbon source which does not inhibit microorganisms and has also been used in other biofloc technology studies (Sakkaravarthi and Sankar, 2015; Zhao et al., 2012) conducted so far. The carbon to nitrogen (C/N) ratios of the material were set to 0.7:1, 1:1, 2:1, 4:1, and 6.07:1, for C0, C1, C2, C4, and C6 treatments, respectively. Moreover, C0 represented control for this experiment with

Table 1 Characteristics of the liquid fraction of the biogas digestate used in this study. Parameter

Experiment I

Experiment II

Unit

COD TAN PO3− 4 -P NO− 2 -N NO− 3 -N TKN TN pH Acetic acid Propionic acid Isobutyric acid Butyric acid Isovaleric acid Valeric acid Total VFAs

10.50 ± 0.3 3808.7 ± 22.1 360.0 ± 24.2 4.4 ± 1.0 9.5 ± 1.1 3979.7 ± 87.9 3993.6a 8.2 ± 0.0 4137.3 ± 512.3 2099.1 ± 58.5 299.5 ± 71.0 139.2 ± 84.1 322.9 ± 74.8 158.2 ± 48.9 7156.3a

11.65 ± 0.6 3457.7 ± 38.9 396.3 ± 20.6 4.9 ± 1.0 8.4 ± 1.9 3702.4 ± 19.4 3714.6a 8.2 ± 0.0 2885.3 ± 566.5 666.7 ± 204.0 78.1 ± 28.1 64.5 ± 25.5 663.4 ± 363.4 133.5 ± 66.5 4491.5a

g L−1 mg L−1 mg L−1 mg L−1 mg L−1 mg L−1 mg L−1 – mg L−1 mg L−1 mg L−1 mg L−1 mg L−1 mg L−1 mg L−1

a

Means calculation based.

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similar preparations but without addition of sucrose. The highest addition of organic carbon in C6 was based on stoichiometry of heterotrophic microorganisms for converting all inorganic nitrogen of LD into microbial biomass (Ebeling et al., 2006). Because volatile fatty acids (VFAs) in LD could work as available organic carbon sources for heterotrophic microorganisms (Ebeling et al., 2006), they were taken into account in all the treatment when calculating the added amount of organic carbon while other traces were neglected. In addition, because of the high ammonia concentration of the digestate and the alkaline conditions, a part of ammonia was expected to volatilize during the experiment period, which could result in the sucrose overdose in C6. Therefore, C4, C2, C1 were included in this study to determine the optimum condition according to the evaluation criteria of water quality, recovered nutrients, and biomass productivity. In experiment 2, the principal aim was to evaluate characteristics of the generated microbial biomass (include contents of gross energy, crude protein, lipids, and pathogens) and discuss its possible utilization. The secondary aim was to evaluate the validity of using the optimum C/ N ratio of experiment I for different organic carbon sources. Therefore, two different sources of organic carbon were experimented and their effects on the biomass and water quality were evaluated. Accordingly, sucrose (S) was used as control and glycerol (G) was experimented as an alternative organic carbon source according to the optimum conditions obtained from experiment I. The purity of the glycerol used was N99% and it was purchased from Sinopharm Chemical Reagent Co, Ltd. Glycerol is the main byproduct of biodiesel industry that is available widely and needs to create new markets (César et al., 2013). After incubating samples for three days, the biomass was collected from all treatments. Then the harvested biomass from each treatment was mixed for carrying out different analyzes and DNA identification. Common pathogens would be discussed according to DNA identification analysis. 2.3. Analytical methods 2.3.1. Water quality measurements Firstly, all samples for physico-chemical analysis of digestate were centrifuged at 6000 rpm (3700 ×g) for 4 min, thereafter, the supernatants were filtered using 0.22 μm glass microfiber filters (Whatman Inc., USA) and was used for water quality evaluations. Total ammonia nitrogen (TAN), total nitrogen (TN), phosphate (PO3− 4 -P), and chemical oxygen demand (COD) were measured according to the standard methods of the American Public Health Association (APHA, 1998). Ni− trite (NO− 2 -N) and nitrate (NO3 -N) were measured calorimetrically according to procedures as described by EPA (1993). Total Kjeldahl nitrogen (TKN) was measured by the Kjeldahl method while pH was measured by electrode based device (model: Mettler Toledo). VFAs were measured using the GC-2010 Plus (Shimadzu, Japan) which was equipped with a flame ionization detector, and a capillary column type Rtx-wax (30 m × 0.25 mm × 0.25 mm). The column had an initial temperature of 60 °C with 5 min hold time, ramp 10 °C min−1 to 140 °C with 2 min hold time, and ramp 20 °C min−1 to 230 °C with 2 min hold time. The temperatures of the injector and detector were 230 °C and 250 °C, respectively. Nutrient removal efficiencies were calculated according to Qin et al. (2016) as shown in
Ri ¼ Si0 −Sif =Si0  100

ð1Þ

where: Ri was the removal efficiency of substrate i (TN, TAN, PO3− 4 -P, COD), Si0 and Sif were the concentrations of substrate i at initial time and final time, respectively. The overall fate of nitrogen in the system was recovered by microorganisms as biomass, volatilized to atmosphere, or remained in the LD. Accordingly, nitrogen mass balance was described by TNi ¼ TN f þ Nm þ Nv

ð2Þ

where, TNi and TNf were initial and final TN amounts in LD, respectively, which were estimated by the measured initial and final concentrations of the LD. Nm was the recovered nitrogen by microorganisms, which was estimated by nitrogen content in the microbial biomass. Nv was the volatilized nitrogen amount, which was estimated by solving Eq. (2). 2.3.2. Microbial growth criteria A gravimetric method was followed to estimate the dry cell weight of biomass, which was used for calculating biomass production and productivity according to Qin et al. (2016). Briefly, a pre-weighted empty sample was filled with known volume and centrifuged at 6000 rpm (3700 ×g) for 4 min. Next, the supernatant was discharged and cell pellets were washed twice with distilled water and dried at 105 °C until a constant weight was reached. Finally, sample was weighed again to estimate the biomass production (g L−1) while productivity was calculated by r ¼μ X

ð3Þ

where r was the dry biomass productivity (g L−1 d−1), X was the dry biomass production (g L−1), μ was the specific growth rate (d−1) which determined by plotting the natural logarithm of dry biomass concentration versus time during the exponential phase of growth. For indicating carbon recovery efficiency, gross carbon efficiency (GCE) was calculated as a ratio of the dry generated microbial biomass to total available organic carbon sources for microorganisms in the medium (gcell/gorganic carbon substrate). Data of GCE were compared with data of the classical thermodynamic principles which obtained by balanced reactions for the aerobic biological conversion of sucrose and VFAs (acetate based) to microbial biomass using ammonia as nitrogen source (TAN represented up to 96% of TN in LD so it was considered the nitrogen source). Sucrose and VFAs were considered as electron donors while the electron acceptor was oxygen because of the aerobic condition of the experiment. Stoichiometry of biological balance reactions was established as described by McCarty (1971, 1972, 1975) and Rittmann and McCarty (2001). 2.3.3. Microbial biomass characteristics evaluation Lipid content was determined by a modified solvent extraction method according to Bigogno et al. (2002). Fifty milligrams of dried samples were used for measuring the content of carbon, nitrogen, hydrogen, and sulfur by using elemental analyzer model Vario EL Cube (Elementar, Germany). The obtained data were used for calculating the crude protein by using factor of 6.25 (Qin et al., 2018). Also, these data were used to calculate higher heating value (HHV) by using wellestablished correlations given by Friedl et al. (2005) as shown in HHV ¼ 3:55C 2 −232C−2230H þ 51:2C  H þ 131N þ 20600

ð4Þ

where HHV was the gross energy (kJ kg−1) while C, N, and H were the percentages of carbon, nitrogen, and hydrogen in dry mass, respectively. Phosphorus element in biomass was evaluated by ashing method according to Donald and Kirkpatrick (1971). To obtain a full understanding of bacterial community structure of the promoted indigenous microflora on LD, DNA from microorganisms was extracted and analyzed as mentioned by Deng et al. (2018). DNA concentrations and purity were quantified and the integrity of extracted genomic DNA was detected by 1% agarose gel electrophoresis. The 16S rRNA genes were amplified using the specific primer for 16S V3-V4: the forward primer 338F (5′-ACTCCTACGGGAGGCAGCAG-3′) and the reverse primer 806R (5′-GGACTACHVGGGTWTCTAAT-3′) to target the V3-V4 regions of 16S rRNA. For Illumina MiSeq gene library construction, DNA from both treatments were amplified in triplicate by PCR. The PCR mixture contained 4 μL of 5 × FastPfu buffer, 2 μL of 2.5 mM dNTPs, 0.8 μL of each primer, 0.4 μL of FastPfu polymerase, and 10 ng template DNA. All mixtures had ddH2O added to a volume of 20 μL

M. Sobhi et al. / Science of the Total Environment 690 (2019) 492–501

and PCR was performed on an ABI GeneAmp® 9700 PCR thermal cycler (Applied Biosystems, Foster City, CA, USA). Steps were as the following, first 95 °C for 3 min, then 27 cycles at 95 °C for 30 s, then 55 °C 30 s, followed by 72 °C for 45 s. Finally, an extension step for 10 min at 72 °C. Amplicons were extracted from 2% agarose gels and purified using the AxyPrep DNA Gel Extraction Kit (Axygen Biosciences, Union City, CA, USA), and then evaluated by QuantiFluor™-ST (Promega, Madison, WI, USA). Purified amplicons were pooled using standard protocols according to Deng et al. (2018). 2.4. Statistical analysis Data of water quality and biomass growth criteria were analyzed statistically using one-way analysis of variance with complete randomized design using SAS 9.1.3 software. Duncan test was applied for multiple average comparisons and determine any significant differences between variables. A significance level of p b 0.05 was used in the statistical analysis. 3. Results and discussion 3.1. Experiment I 3.1.1. Effect of carbon/nitrogen ratio on microbial biomass production As shown in Table 2, optimum microbial biomass production and productivity were 6.12 g L−1 and 3.69 g L−1 d−1, respectively when C/ N ratio was 2:1 in C2 while the lowest values were 1.9 g L−1 and 0.41 g L−1 d−1 in C0 (control treatment) which had C/N ratio of 0.7:1. Results revealed that the biomass production increased significantly (P b 0.05) with adding organic carbon until reaching the peak at the level of C2, while no significant change was observed in biomass production compared with the higher doses of organic carbon addition in C4 and C6. Further, the maximum gross carbon efficiency (GCE) was 0.33 gcell/gsubstrate in C2 while the minimum efficiency was 0.1 gcell/gsubstrate in C6 when C/N ratio reached as high as 6.07:1. The GCE started with 0.27 gcell/gsubstrate for the control and increased by adding sugar until C2 then decreased significantly with more sugar addition (Table 2). The balanced equations of stoichiometry of aerobic heterotrophic microflora which obtained by thermodynamic principles using acetate and sucrose were as the following, CH3COO + 0.236 NH4 + 0.824 O2 → 0.236 C5H7O2N + 0.764 H2O + 0.76 HCO3 + 0.056 CO2 (acetate based) (5) C12H22O11 + 1.73 NH4 + 3.34 O2 + 1.73 HCO3 → 1.73 C5H7O2N + 9.27 H2O + 5.07 CO2 (sucrose based) (6) where C5H7O2N presents the chemical formula for microbial biomass (Ebeling et al., 2006). Accordingly, the maximum theoretical values of GCEs for acetate and sucrose were 0.45 and 0.57 gcell/gsubstrate, respectively. These values represented the maximum possible conversion ratios of all available organic carbon substrates to biomass. Because treatments contained different ratios of sucrose and VFAs, the maximum theoretical GCEs were different and they stated as balanced weighted means in Table 2. Accordingly, all values of GCEs obtained in this experiment were lower than the theoretical efficiencies

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indicating that part of organic carbons remained in the medium. Although C2 gave the optimum recovery of organic carbon, the process still needs to be improved in future studies. In fact, C4 and C6 did not perform better than C2 due to a relative lack of inorganic nitrogen during cultivation, which happened as a result of ammonia volatilization (as reported later in 3.1.2). Thereafter, the available organic carbon became not compatible with the available inorganic nitrogen. Accordingly, C2 was preferred due to its relatively little demand of organic carbon and its high biomass production. To compare the results of this study and other related studies, an intensive review of literature was summarized in Table 3. It revealed that scientists have focused mainly on utilizing microalgae cultivated under mixotrophic mode, while relatively little researches focused on other microorganisms such as yeasts or bacteria for digestate treatment and biomass production. Further in most of the studies carried out so far, diluted digestates or low strength digestates were used in addition to apply other pre-processes such as sterilizations and strains additions, which increases the total operational costs. However, in this study organic carbons were added after the separation only thus representing an economically feasible option. Additionally, a short cultivation period of 3 days as compared with 6 to 14 days for other studies makes the process more attractive. The highest biomass production as reported by Cheng et al. (2015) was 4.81 g L−1 (Table 3) whereas up to 6.12 g L−1 dry matter base (DM) was achieved in this study. 3.1.2. Effect of carbon/nitrogen ratio on water quality Nutrient removal efficiencies are important indicators for evaluating the indigenous heterotrophic microorganisms' wastewater treatment system. The removed nutrients in the system may be assimilated and biosorbed by microorganisms, or released in a gas form due to aerobic condition and stir. Fig. 1 shows the concentrations and removal efficiencies of water quality parameters including TAN, TN, COD, and PO3− 4 -P, − and NO− 2 -N & NO3 -N. TAN and TN were removed successfully in all the treatments as illustrated in Fig. 1A and Fig. 1F. The removal efficiencies of TAN reached about 60% and 88% after 24 and 48 h, respectively while the final removal efficiencies were higher than 98% and 95% of TAN and TN, respectively, without any significant differences between treatments for the removal rates of TAN and TN (P N 0.05). TAN decreased from 3809 mg L−1 to a range of 21 mg L−1 to 77.6 mg L−1 in C2 and C0, respectively while TN decreased from 3994 mg L−1 to a range of 120 mg L−1 to 205 mg L−1 in C2 and C0, respectively. In fact, these removal efficiencies agreed with other studies which ranged from 90% to 100% and 86% to 100% of TAN and TN, respectively, as reported in Table 3. The fate of N would be explained in details in experiment II. According to Ayre et al. (2017), Chlorella sp and Scenedesmus sp. were cultivated with undiluted liquid digestate of ammonium concentration up to 800 mg L−1 for avoiding negative impacts. However, the indigenous microflora in this study could be cultivated successfully with a high initial TAN concentration of 3809 mg L−1. It may be because the indigenous microflora has already adapted with high ammonia concentration in digestate during the digestion process and at the same time, ammonia has been removed from LD in high rates during the process. Previously, biofloc technology proved that the indigenous microflora

Table 2 Biomass production, productivity, and carbon recovery efficiency with different C/N ratios. Treatment C0 C1 C2 C4 C6

Biomass production (g L−1)

Productivity (g L−1 d−1)

GCEs, Experimental values (gcell/gsubstrate)

GCEs, Theoretical Values (gcell/gsubstrate)

Phosphorus ratio in dry biomass, %

1.90 ± 0.26b 2.65 ± 0.20b 6.12 ± 0.42a 5.37 ± 0.27a 5.64 ± 0.79a

0.41 ± 0.11b 0.86 ± 0.14b 3.69 ± 0.39a 3.01 ± 0.24a 3.26 ± 0.72a

0.27 ± 0.03b 0.28 ± 0.08b 0.33 ± 0.02a 0.14 ± 0.06c 0.10 ± 0.01d

0.45 0.48 0.52 0.55 0.56

3.7 ± 0.56 3.2 ± 0.40 4.1 ± 0.23 3.9 ± 0.51 4.2 ± 0.37

Different Superscripts (a–d) within the same column indicate significant differences (p b 0.05).

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Table 3 Comparing the correlated studies with Sucrose (S) & Glycerol (G) treatments with C/N ratio of 2:1. Inoculum

Digestate origin

Digestate pretreatment

Operation

Biomass production & productivity

Chlorella vulgaris ATCC13422 “Autotrophic”

Food waste

Dilution 10 times autoclaving separation Dilution sterilization glycerol addition Filtration, sterilization dilution Centrifugation autoclave dilution

Batch (7 d)

1.5 g L−1

Y.lipolytica &Chlorella vulgaris Dairy wastewater “Heterotrophic&mixotrophic” Chlorella pyrenoidosa “mixotrophic”

Starch processing

Chlorella PY-ZU1 “mixotrophic”

Swine manure and sewage

Desmodesmus sp.

Pig manure

Chlorella sp. “mixotrophic”

Wastewater sludge

Chlorella sp. Scenedesmus sp. “Autotrophic”

Pig manure

Indigenous microflora “Heterotrophic”

Chicken manure

Indigenous microflora “Heterotrophic”

Chicken manure

Filtration dilution Mixed with wastewater dilution Separation filtration CO2 addition pH adjusting Non-dilution Separation sucrose addition

Separation glycerol addition

Batch (6 d)

Batch (9 d)

Batch (13 d)

Batch (14 d) Batch (until stationary phase)

continuous

TAN N 90% PO3− 4 -P 77% COD 87% −1 1.62 g L TAN 100% PO3− 0.21 g L−1 d−1 4 -P 100% (P in biomass was 2.2%) COD 72% 3.01 g L−1 TN 91.6% TP 90.7% 0.58 g L−1 d−1 COD 75.8% −1 4.81 g L TAN 73% −1 −1 TP 95% 0.60 g L d COD 79% 0.41 g L−1 TN 100% TP 100% 0.03 g L−1 d−1 2.11 g L−1 TN 83.7% TP 94.2% 0.45 g L−1 d−1 COD 86.3% −1 −1 19 mg L d Up to 5.1% d−1

Batch (3d)

6.9 ± 0.20 g L−1 4.9 ± 0.21 g L−1 d−1

Batch (3d)

4.6 ± 0.10 g L−1 2.7 ± 0.09 g L−1 d−1

could grow and recover ammonia from aquatic systems, which have low concentrations of TAN (b0.5 mg L−1) as reported by Qian et al. (2018). However, this study has proved their ability to grow and recover ammonia from very high strength wastewater of LD. The removal ratio of nitrite and nitrate was ranged between 60% to 70% for all treatments except the control (C0) where the reduction − was 22% as illustrated in Fig. 1.E. In fact, NO− 2 -N&NO3 -N presented 0.35% of TN of LD in this experiment, thus TN removal was mainly subject to TAN removal. COD is also a critical parameter for determining optimum treatment as it is considered as a very important index for water contamination. As shown in Fig. 1.C, the initial COD of LD was about 10.5 g L−1, however, the starting points of each treatment were different due to the different added amounts of sucrose. The optimum total COD removal efficiency was 67% in C2 while C6 recorded the least removal efficiency of 19%. COD concentrations at the end time of the experiment were less than that of original digestate in C0, C1 and C2 only with concentrations of 6.9, 9.3 and 7.4 g L −1 , respectively, while it reached 30 and 53 g L−1 in C4 and C6, respectively. As indicated in Fig. 1.C, no significant differences between C0 and C2 and they were optimum while other treatments differed significantly (p b 0.05). The optimum COD removal efficiency in this study was comparable to the results reported by Qin et al. (2018) who cultivated Chlorella vulgaris with organic carbon (glycerol) added medium where 72% of COD was removed. However, Praveen et al. (2018) reported a higher COD removal efficiency of 87% which was observed in the cultivation of Chlorella vulgaris without external organic carbon addition. Therefore, the COD removal efficiency is vulnerable to be affected by the external organic carbon addition which should be synergistically optimized with biomass production to avoid the secondary treatment of the liquid. PO3− 4 -P was effectively removed and recovered by microorganisms in this study. The removal efficiencies according to the analysis of PO3− 4 -P in LD were 23%, 29%, 69%, 66%, and 72% for C0, C1, C2, C4, and

Nutrient removal efficiency

TN 96.8% TAN N 99% PO3− 4 -P 68% COD 67% TN 96.5% TAN N 99% PO3− 4 -P 53% COD 57%

Ref.

Praveen et al. (2018)

Qin et al. (2018)

Yang et al. (2015)

Cheng et al. (2015)

Ji et al. (2014) Mortensen et al. (2014) Ayre et al. (2017)

This study EXP.II

This study EXP.II

C6 respectively. No significant differences existed between C0 and C1 or between C2, C4, and C6, however, both groups differed significantly (p b 0.05). On other hand, according to the phosphorus analysis in biomass, it could be recovered inside the cells by 3.2% to 4.2% DM as mentioned in Table 2 and it matched the decrease of PO3− 4 -P in the LD. In general, microorganisms can accumulate phosphorus by 1–2% DM inside cells because phosphorus is a necessary element for several cellular processes such as energy transfer and synthesis of DNA, RNA and other components of nucleic acids (Orhon, 1997; Rittmann and McCarty, 2001). However, some strains of heterotrophic microorganisms under specific conditions can accumulate phosphorus within range of 3–8% (Metcalf, 2003). This occurs through a process known as enhanced biological phosphorus removal technique, which considers a pre-process of storing wastewater in an anaerobic tank before applying an aeration stage. Then, a group of heterotrophic bacteria, called Polyphosphate-Accumulating Organisms (PAOs) accumulates larger amount of phosphorus inside their cells and enhances phosphorus recovery (Metcalf, 2003). Because LD has already been implemented under anaerobic conditions during the process of biogas production, high ratios of phosphorus could be recovered inside cells within Metcalf's range after growing the indigenous microflora aerobically. pH levels of all treatments as shown in Fig. 1.B were increased from 8.18 to about 9.20 after one day. After that, pH values remained constant approximately in C0 and C1 while it decreased in C2, C4, and C6 significantly (P b 0.05) and reached 8.61, 8.07 and 7.36, respectively. This increase during the first day was because of aerobic and mixing conditions, which reduced the high concentrations of dissolved CO2 in the digestate. However, the decreasing after the first day happened because of the conversion of sucrose to pyruvate acids through glycolysis process (Nelson and Cox, 2008). Accordingly, the cultivating of the indigenous heterotrophic microorganisms could successfully produce microbial biomass in high productions rates in addition to recover nutrients and treat wastewater.

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Fig. 1. Water quality parameters' concentrations and removal efficiencies in the treatments using different C/N ratios. Notation: Different letters (a–d) within the same figure indicate significant differences (p b 0.05). Digestate column in fig. (1.D, 1.E, and 1.F) refers to the original concentration of parameters in the utilized liquid digestate.

C2 treatment gave optimum results thus it would be applied in the second experiment. 3.2. Experiment II 3.2.1. Effect of carbon sources on microbial biomass production and water quality The biomass production was 6.9 g L−1 for S treatment while it was 4.6 g L−1 in G treatment when they applied the same C/N ratio of 2:1 (Table 3). Water quality was discussed in details in the first part, however, the effects of using glycerol compared with sucrose were presented and discussed in this context briefly. For TAN and TN removal, both treatments were equal approximately. As seen in Table 4, TAN final concentrations were b30 mg L−1 for both treatments with removal efficiencies higher than 99% while TN concentrations were 119.3 and 131 mg L−1 with removal efficiencies of 97% and 96% for S and G treatments, respectively and these results agreed with first experiment and other studies in Table 3. On other hand, COD removal was better in S treatment due to the higher generation of biomass which in effect demanded more nutrients accordingly. Final COD concentration

reached 7.96 g L−1 in S treatment while it was 11.13 g L−1 for G treatment. Phosphorus removal was higher in sucrose treatment due to the higher biomass production. PO3− 4 -P concentration of LD decreased from 396 mg L−1 to 128 mg L−1 and 204 mg L−1 for S and G treatments, respectively. PO3− 4 -P removal efficiencies reached 68% and 49% for S & G treatments, respectively. On other hand, the absolute recovered amount in this experiment according to phosphorus content in the generated biomass reached 281 mg L−1 (4.1% DM) and 163 mg L−1 (3.6% DM) for S and G treatments, respectively. In fact, phosphorus accumulation efficiencies inside cells agreed with Metcalf's range (Metcalf, 2003) and higher than the other studies as mentioned in Table 3. Clearly, using different organic carbon sources at same C/N ratio affected the results of biomass growth criteria and some parameters of water quality significantly (p b 0.05), therefore glycerol or other organic carbon sources could be used instead of sucrose however the optimum C/N ratio need to be determined first. In fact, this result has opened the door to experiment more organic carbon source-based byproducts and wastes for this application especially, a comparable microbial biomass production with other studies of Table 3 was achieved by using glycerol.

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Table 4 Water quality and Biomass characteristics of Sucrose (S) & Glycerol (G) treatments with C/ N ratio of 2:1. Sucrose

Glycerol

Water quality parameter TAN (mg L−1) −1 PO3− ) 4 -P (mg L COD (mg L−1) − − NO3 &NO2 (mg L−1) TN (mg L−1) pH

26.9 ± 2.9 128.2 ± 23.3 7.96 ± 0.3 5.11 ± 0.4 119.3 ± 9.6 8.63 ± 0.2

20.46 ± 0.5 203.9 ± 28.0 11.13 ± 0.2 4.38 ± 0.4 131.03 ± 10.4 8.85 ± 0.2

Biomass characteristics C % DM N % DM H % DM O % DM P % DM S % DM Ash % DM HHV MJ/kg (kJ L−1) Crude protein % (g L−1) Lipids % (g L−1)

45.9 10.1 6.9 26.4 4.1 0.74 10.0 19.6 (135.2)a 63.0 (4.4)a 10.8 (0.75)a

45.5 10.4 6.6 26.1 3.6 0.67 10.7 19.4 (89.2)a 65.1 (3.0)a 10.9 (0.50)a

a The number in bracket referred to the gross yield of crude protein or lipids while outer number referred to the content percentage in the biomass.

3.2.2. Nitrogen mass balance Based on the nitrogen content in the microbial biomass (mentioned in Table 4), initial and final TN concentrations of the LD, about 19% and 13% of TN were recovered by the cultivated indigenous microorganisms for S and G treatments, respectively, as illustrated in (Fig. 2). In addition, 3% and 4% of TN remained in LD for S and G treatments, respectively. Finally, about 78% and 83% of TN was lost through volatilization as ammonia gas from S and G treatments, respectively, due to the effects of shaking, alkaline, and aerobic conditions of the experiment. In fact, effects of classic denitrification could be neglected because it happens in oxygen absent (Lena et al., 2015), while aerobic conditions were applied − in this study, in addition to the limited presence of NO− 2 -N&NO3 -N by 0.36% of TN according to Table 1. In fact, the share recovered by microorganisms were comparable with the study of Ayre et al. (2017) who recovered 10% of total nitrogen from an undiluted digestate using a mix of Chlorella sp. and Scenedesmus sp. Accordingly, these results motivate establishing compact reactors, which merge the processes of producing microbial biomass and recovering the volatilized ammonia in one process. It may increase nitrogen recovery ratio, reduce the total

Fig. 2. Nitrogen fate according to mass balance analysis in the treatment using sucrose and glycerol as organic carbon sources.

operational costs and recover the volatilized ammonia as ammonium salts (byproduct) without adding external chemicals. 3.2.3. Microbial biomass characteristics Clearly, the weight based elemental compositions for dried microbial biomass of both treatments were similar approximately C5H8.9O2.2N0.9 and C5H8.6O2.2N for S and G treatments, respectively, which were very close to the chemical formula of microbial biomass C5H7O2N. As reported in the Table 4, crude protein contents were 63% and 65%, while the crude protein yields were 4.35 and 2.99 g L−1, for S and G treatments, respectively. HHV was about 19.5 MJ kg−1 for both treatments which was comparable to 20.1 MJ kg−1 for fishmeal (Shao et al., 2017), but the energy yield was higher in S treatment (135.24 kJ L−1) than G treatment (89.24 kJ L−1). Lipids contents of the biomass of this study were relatively low, about 11% for both treatments while lipids yields were 0.75 g L−1 and 0.5 g L−1 for S and G treatments, respectively. Finally, ash ratios were 10.0% and 10.7% for S and G treatments, respectively. Therefore, the harvested biomass has a high possibility for the following application. Crude protein is a critical factor affecting the biomass viability for feed production. This biomass may be used for extracting protein or for direct feeding applications in case of utilizing a pure and disinfected digestate (agricultural origin). This application will be similar to the experiment of Shao et al. (2017) in which, a biofloc meal was prepared from an aquaculture system and replaced fishmeal partially for fish feeding successfully. In fact, the microbial biomass of this study has a very close profile similar to fishmeal profile as illustrated in Fig. 3 hence, it may become an interesting alternative for fishmeal especially in conjunction with high fishmeal prices of 1500$ approximately per metric ton (GEM, 2018). In fact, chicken manure is utilized as feed for animal feeding directly (FAO, 2019). Therefore, the generated biomass from digestate of chicken manure may sound interesting in the future as an alternative protein source. The HHV is a major indicator of biomass quality for energy applications. This microbial biomass was suitable for energy applications such as hydrothermal liquefaction (HTL) - a process for producing bio-oil (Petroleum crude oil alternative) from different bio-sources. According to Vardon et al. (2011), 34% of Spirulina microalgae was successfully converted to bio-oil contained 33 MJ kg1 by HTL although Spirulina had low lipids content of 5% and high protein content of 64%. They mentioned that the microbial biomass that had high protein content - such as the biomass of this study - could be converted to bio-oil by HTL effectively. In fact, oleaginous microorganisms that accumulate N20% lipids, are preferred for biodiesel industry (Bautista and Vicente, 2012). However, the high production of biomass compensated the low lipids contents. Also, this biomass as a substrate for biodiesel industry needs more researches for technical and economic assessments. In additions, some strategies may improve its ability to accumulate lipids by changing the cultivation conditions such as increasing C/N ratio followed by increasing cultivation period and applying stress conditions such as nutrient limitation especially after the maturity of microorganisms in mediums (Bautista and Vicente, 2012). 3.2.4. Bacterial community According to Fig. 4, it is clear that the bacterial communities' composition were significantly different between S and G treatments. Different organic carbon substrates had an influence on the dominant genera of bacteria because substrates have different characteristics and each one needs specific microorganisms to be assimilated. In S treatment, genus of Thiopseudomonas dominated by 36% while the dominant genus in G treatment was Atopostipes by 29%. Both genera are classified as PAOs according to the reports by Kleemann (2015) and Tan et al. (2015). Also, there were many other PAOs genera dominated by N1% in both treatments such as Bacillus, Pseudomonas, and others. These results were compatible with the previous

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Fig. 3. Composition of fishmeal compared with microbial biomass in the treatment using sucrose and glycerol as organic carbon sources. Notations: HHV was illustrated as points (●) on the secondary axis. Fishmeal characteristics were reported according to Shao et al. (2017).

interpretation of PO 3− 4 -P recovery from the LD and supported the reason of high phosphorus accumulation ratios inside cells according to Metcalf's range (Metcalf, 2003). On the other hand, according to Moheimani et al. (2018), genera of Coliform, Staphylococcus, Clostridium, Salmonella, Shigella, Campylobacter, and Listeria are the most common pathogenic microorganisms affecting microbial products where none of them was observed in the biomass of both treatments of this experiment. In fact, anaerobic digestion can decline the relative abundance of

pathogenic microorganisms of organic substrates effectively (Qian and Yu, 2019) however some pathogenic species might remain in the digestate after anaerobic digestion (Atif et al., 2018). Therefore, it is necessary to use pure and dis-infective digestate (without pathogens or heavy metals) if the generated biomass will be used for feeding otherwise the generated biomass can be used as an energy substrate. On the other hand, the generated biomass could be used for feeding after processing using techniques such as pasteurization, drying and many others. Hence, no further concerning about

Fig. 4. Bacterial community analysis abundance on genus level in the treatment using sucrose and glycerol as organic carbon sources. Notations: All genera dominated by N1% was included, otherwise they were included within “other” part. Unclassified genus belonged to Microbacteriaceae family dominated in G treatment by 2.92% and it was included in “other” part of this treatment.

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potential pathogenicity of other genera and it will be more secure than direct manure feeding applications. However, evaluating the full nutritional profile for generated biomass and carrying out real feeding experiments are highly recommended. 4. Conclusion Nutrient recovery and biomass production were achieved simultaneously using heterotrophic indigenous microflora in liquid biogas digestate. As high as 68% of PO3−-P and 97% of TN could be removed from the undiluted liquid digestate. Additionally, N6 g L−1 of dry biomass was simultaneously produced and featured with up to 65% crude protein without pathogens and 10.9% lipids, which has a big potential to be used as either an alternative source of animal feed or as a substrate for energy applications. Results proved that the indigenous microflora could recover nutrients effectively from LD as a biofloc technology. Also, an integrated process of producing microbial biomass and recovering the volatilized ammonia would be motivated in the future. Acknowledgments This work was financially supported by the China National Key Research and Development Plan Project (Grant No. 2018YFD0800100) and “double first-class” program. Also, the first author would like to thank Prof. Aziz Nour (Faculty of Agriculture, Alexandria University, Egypt) and Ms. Han Tongtong (Key Laboratory for Clean Renewable Energy Utilization Technology, Ministry of Agriculture, College of Engineering, China Agricultural University) for their support. References APHA, 1998. Standard Methods for the Examination of Water and Wastewater. 20th edition. American Public Health Association, American Water Works Association and Water Environmental Federation, Washington DC. Atif, M., Shubiao, W., Jiaxin, L., Zeeshan, A., Hongzhen, L., Renjie, D., 2018. Nutrient recovery from anaerobically digested chicken slurry via struvite: performance optimization and interactions with heavy metals and pathogens. Sci. Total Environ. 635, 1–9. Avnimelech, Y., Kochba, M., 2009. Evaluation of nitrogen uptake and excretion by tilapia in biofloc tanks using 15N tracing. Aquaculture 287, 163–168. Ayre, J.M., Moheimani, N.R., Borowitzka, M.A., 2017. Growth of microalgae on undiluted anaerobic digestate of piggery effluent with high ammonium concentrations. Algal Res. 24, 218–226. Azim, M.E., Little, D.C., 2008. The biofloc technology (BFT) in indoor tanks: water quality, biofloc composition, and growth and welfare of Nile tilapia (Oreochromis niloticus). Aquaculture 283, 29–35. Bautista, L.F., Vicente, G., 2012. Biodiesel from microbial oil. In: Rafael, L., Juan, A.M. (Eds.), Advances in Biodiesel Production. Woodhead Publishing Limited, UK, pp. 179–203. Bigogno, C., Khozin-Goldberg, I., Boussiba, S., Vonshak, A., Cohen, Z., 2002. Lipid and fatty acid composition of the green oleaginous alga Parietochloris incisa, the richest plant source of arachidonic acid. Phytochemistry 60, 497–503. Carmen, M., Mejía, J.C., 2017. Probiotics used in Biofloc system for fish and crustacean culture: a review. J. Fish. Aquat. Sci. 5, 120–125. Cerda, A., Mejias, L., Rodríguez, P., Rodríguez, A., Artola, A., Font, X., Gea, T., Sánchez, A., 2019. Valorisation of digestate from biowaste through solid-state fermentation to obtain value added bioproducts: a first approach. Bioresour. Technol. 271, 409–416. Cheng, J., Xu, J., Huang, Y., Li, Y., Zhou, J., Cen, K., 2015. Growth optimisation of microalga mutant at high CO2 concentration to purify undiluted anaerobic digestion effluent of swine manure. Bioresour. Technol. 177, 240–246. Deng, M., Chen, J., Gou, J., Hou, J., Li, D., He, X., 2018. The effect of different carbon sources on water quality, microbial community and structure of biofloc systems. Aquaculture 482, 103–110. Drekeke Iyovo, G., Guocheng, D., Jian, C., 2010. Sustainable bioenergy bioprocessing: biomethane production, Digestate as biofertilizer and as supplemental feed in algae cultivation to promote algae biofuel commercialization. Journal of Microbial & Biochemical Technology 2, 100–106. Ebeling, J.M., Timmons, M.B., Bisogni, J.J., 2006. Engineering analysis of the stoichiometry of photoautotrophic, autotrophic, and heterotrophic removal of ammonia-nitrogen in aquaculture systems. Aquaculture 257, 346–358. EPA, 1993. Determination of Nitrate-Nitrite Nitrogen by Automated Colorimetry. Environmental Protection Agency, U.S. Method 353.2, revision 2.0. FAO, 2019. Feeding Poultry Waste to Sheep. In Chapter 2 Feeding Animal Wastes. Online article. http://www.fao.org/3/X6518E/X6518E03.htm. Friedl, A., Padouvas, E., Rotter, H., Varmuza, K., 2005. Prediction of heating values of biomass fuel from elemental composition. Anal. Chim. Acta 544, 191–198.

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