Journal of Cleaner Production 286 (2021) 125468
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Journal of Cleaner Production
journal homepage: www.elsevier.com/locate/jclepro
Life cycle assessment of household biogas production in Egypt:
Influence of digester volume, biogas leakages, and digestate
valorization as biofertilizer
Lida Ioannou-Ttofa a, Spyros Foteinis a, Amira Seifelnasr Moustafa b, Essam Abdelsalam c,
Mohamed Samer b, Despo Fatta-Kassinos a, d, *
a
Nireas-International Water Research Center, University of Cyprus, P.O. Box 20537, CY 1678, Nicosia, Cyprus
Cairo University, Faculty of Agriculture, Department of Agricultural Engineering, El-Gammaa Street, 12613, Giza, Egypt
National Institute of Laser Enhanced Sciences (NILES), Cairo University, 12613, Giza, Egypt
d
Department of Civil and Environmental Engineering, University of Cyprus, P.O. Box 20537, CY1678, Nicosia, Cyprus
b
c
a r t i c l e i n f o
a b s t r a c t
Article history:
Received 5 May 2020
Received in revised form
16 November 2020
Accepted 7 December 2020
Available online 9 December 2020
Biogas production from animal manure can address many negative impacts of traditional energy generation and also improve living conditions in rural communities in Egypt, the case study herein, and
further afield. Even though techno-economical aspects of the household biogas digester technology have
been thoroughly studied, its environmental sustainability, especially under Egypt’s conditions, remains
largely unknown. To this end, life cycle inventory data were collected from typical fixed-dome digesters
operating in Egypt. Environmental modelling was based on the life cycle assessment methodology using
SimaPro. It was identified that the 100-year global warming potential for producing 1 m3 of biogas under
Egypt’s conditions amounts to 2.72 kg CO2eq., while its total environmental footprint was 160.1 mPt. The
main contributor was the operational phase (89.1%), while the construction phase had a much smaller
contribution (10.9%). The main environmental hotspots were identified as the manure required to drive
the process, closely followed by biogas leakages and intentional releases. By minimizing biogas losses,
the system’s environmental sustainability largely improves (~60% reduction) and could be on the same
level with the one of larger biogas units operating in developing countries. Furthermore, it was identified
that the digester volume plays an overall small role in the system’s environmental performance; however, oversized digesters grossly affect the environmental sustainability, due to the large amounts of
biogas intentional releases. Finally, the use of digestate as a biofertilizer appears to be environmentally
sustainable (~38% reduction of total environmental footprint). Taken together, the results obtained in this
study provide substantial information for policy- and decision-making on renewable energy development in rural Egypt and beyond.
© 2020 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND
license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
Handling editor: Bin Chen
Keywords:
Biogas
Fixed-dome digester
Life cycle assessment/analysis (LCA)
Environmental impact assessment
Biofertilizer
Sensitivity analysis
1. Introduction
Nowadays, about 1.6 billion people, i.e. one fourth of the world’s
population, do not have access to electricity, mostly in rural areas
(Garfi et al., 2016). Furthermore, worldwide, approximately 2.4
billion people still depend on traditional biomass, such as firewood,
agricultural residues and dried dung for their cooking and heating
needs (Kanagawa and Nakata, 2008). As a result, large quantities of
* Corresponding author. Nireas-International Water Research Center, University
of Cyprus, P.O. Box 20537, CY 1678, Nicosia, Cypru.
E-mail address: dfatta@ucy.ac.cy (D. Fatta-Kassinos).
carbon dioxide (CO2) are emitted from the traditional cooking
stoves used in rural areas, contributing to the global warming and
climate change. The incomplete combustion of biomass occurred in
these stoves also release toxic and hazardous emissions, such as
carbon monoxide (CO), nitrous oxide (N2O), methane (CH4), polycyclic aromatic hydrocarbons (PAHs), and other organic compounds
(Bhattacharya et al., 2000). These are responsible for serious impacts, both on the environment and human health (Miah et al.,
2009).
Therefore, biogas production has enjoyed support in developing
countries, by both local authorities and international organizations
(Thu et al., 2012), on account of being a renewable energy
https://doi.org/10.1016/j.jclepro.2020.125468
0959-6526/© 2020 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-nd/4.0/).
L. Ioannou-Ttofa, S. Foteinis, A. Seifelnasr Moustafa et al.
Journal of Cleaner Production 286 (2021) 125468
production in developing countries, including Ethiopia (Lansche
and Müller, 2017; Gabisa and Gheewala, 2019), Kenya (Nzila et al.,
2012), Brazil (Mendes et al., 2004), Bangladesh (Rahman et al.,
rez et al.,
2017), Colombia (Garfi et al., 2019), Latin America (Pe
2014), India (Bruun et al., 2014; Sfez et al., 2017), Vietnam (Vu
et al., 2015), Pakistan (Yasar et al., 2017), Mexico (Ramírez-Arpide
et al., 2018) and China (Han et al., 2010; Wang and Zhang, 2012;
Chen and Chen, 2013; Zhang et al., 2013; Hou et al., 2017; Wang
et al., 2018; Liu et al., 2018).
Specifically, Gabisa and Gheewala (2019) estimated that manure
biogas produced by 4500 household digesters operated in Ethiopia
can lead to huge reductions in total GHG and indoor pollutants (i.e.
the annual reduction was as follows: 1984 ton of CO2eq, 1516 ton of
CO, 108 tn of NMOC and 41 ton of PM10). Garfi et al. (2019) findings
shown that low-cost digesters can reduced up to 80% the environmental impacts associated with manure handling, as well as
fuel and fertilizer use in Colombian farms. The main environmental
hotspots of the household biogas plants operated in northwest
China were found to be the emissions through poor air tightness
and the lack of technical biogas purification, according to the LCA
study of Wang et al. (2018). Moreover, approximately 1.34 104 tons
of CO2eq can be saved annually by the operation of 8000 biogas
plants installed in Ethiopia (Lansche and Müller, 2017). According
to the studies of Hou et al. (2017) and Bruun et al. (2014), it was
found that small-scale biogas digesters can be a very useful manure
management tool, reducing significantly global warming impacts
when the digesters are used and managed appropriately. However,
poorly designed and not well managed biogas systems can increase
GHG emissions, having greater environmental impacts than the
impacts avoided by the replacement of fossil fuels used in developing countries (i.e. China and India) (Hou et al., 2017; Bruun et al.,
2014). Vu et al. (2015) found that biogas digesters in Vietnam
reduced CO2 emissions up to 27%. According to Zhang et al. (2013),
the annual CO2 emission reduction was found to be 1.25 tons by
using an 8 m3 household digester in rural China, while Wang and
Zhang (2012) found an emission reduction up to 2878.30 kg CO2.
However, most of them solely focused on GHG emissions, while
to the best of authors’ knowledge a comprehensive study dealing
with the Egyptian case-study scenario is missing from the literature. For this reason, actual life cycle inventory (LCI) data for typical
household biogas digesters were collected from systems already in
operation in rural areas in Egypt and their environmental sustainability was examined by means of the LCA methodology (SimaPro
9). Sensitivity analyses, which are grossly missing from the literature, were carried out to identify avenues to improve the environmental sustainability of this technology and provide decisionand policy-makers with tools for promoting sustainable development strategies in rural Egypt and beyond.
technology that can mitigate many of the impacts (e.g. climate
change, deforestation, and air pollution) of traditional energy
generation (Blenkinsopp et al., 2013). In remote rural communities,
where energy infrastructure may be weak or non-existing, household biogas digesters are considered a simple and effective technology to meet daily energy needs (cooking, lighting and heating),
thus substantially improving living conditions (Bond and
Templeton, 2011; Garfi et al., 2016; Hijazi et al., 2019). Furthermore, biogas technology could address handling problems of industrial, agricultural, and domestic wastes, and at the same time
produce renewable electricity, heat, and fertilizer (Essential
et al., 2014). More imporConsulting OregonDairy, 2009; Lijo
tantly, biomass can be considered as a relatively secure source of
domestically produced energy, which is not affected by fluctuations
in fuel prices (e.g. petroleum, natural gas) (Demirbas, 2008) and can
largely cover the energy needs of developing countries (Muench
and Guenther, 2013).
In rural areas, cellulosic biomasses, such as manure and agricultural residues, are abundant, which suggest their importance as
an energy source, particularly in the domestic sector of developing
countries (Sreekrishnan et al., 2004). As a result, in countries like
Egypt household biogas digesters are gaining popularity (Samer,
2012; Thu et al., 2012). However, even though biomass waste can
contribute ~151 PJ of primary energy in Egypt, it is not appropriately managed and practically energy is not produced (NREGA,
2002; Said et al., 2013). Therefore, biogas production can help
Egypt to sustainably manage biomass waste and produce renewable energy, in order to enact positive change both locally and
globally. To this end, household digesters can play an important role
and be a useful manure management tool, provided that they are
well-designed and operate appropriately (Hou et al., 2017). Biogas
plant designs largely vary, depending on cost, structure, substrate
availability and energy demand (Nzila et al., 2012). Worldwide, the
most popular design is the Chinese fixed-dome household digester,
followed by the Indian floating drum and the Taiwanese plastic
rez et al., 2014). However, emphasis should be
tubular type (Pe
placed to the design of household biogas systems, since poorly
designed systems could constitute a virtual climate bomb (Bruun
et al., 2014). Therefore, focus should be placed to the local conditions (e.g. insulated in cold climates, etc.) and farmers’ needs with
respect to manure management and local energy requirements
(Hou et al., 2017). In rural Egypt the Chinese type is popular, with
more than one thousand units already (co)funded by the United
Nations Development Programme (UNDP) and built by the Egyptian Ministry of Environment (MoE). However, their environmental
sustainability, along with their main environmental hotspots under
rural Egypt’s conditions remain largely unknown and hence were
comprehensively studied herein using the life cycle assessment
(LCA) methodology (Zhang et al., 2013).
LCA attributes environmental impacts/damages by quantifying
raw materials, energy use and emissions/wastes associated with a
process/system (Ioannou-Ttofa et al., 2016). It also identifies environmental hotspots, i.e. the by-processes that largely affect the
environmental impacts of the process, enabling thus the identification of more environmentally sustainable alternatives
(Evangelisti et al., 2014; Abdelsalam et al., 2019). Since the mid2000s, LCA has gained popularity as a tool for assessing the environmental sustainability of biogas production and use (Muench
and Guenther, 2013). However, in the existing literature most
works deal with biogas production systems based in Europe and
using different feedstocks, with focus given on GHG emissions and
et al., 2014; Vega et al.,
fossil fuel depletion impact categories (Lijo
2014; Evangelisti et al., 2014; Fuchsz and Kohlheb, 2015; Ertem
et al., 2016). In recent years, only few LCA studies have dealt with
animal and agricultural waste management towards biogas
2. Methodology
2.1. Goal and scope
In this study, the main goal was to examine, identify, and assess
the environmental performance and main environmental hotspots
of typical Chinese-type (fixed-dome) household digesters operating in rural Egypt. Specifically, in the framework of the UNDP’s
“Bioenergy for rural development” project, more than one thousand such units were constructed in rural communities, to promote
Egypt’s sustainable rural development, reduce environmental impacts associated with the use of fossil fuels, and improve the
environmentally unsound management practices of agricultural
and solid waste (Egyptian Ministry of Environment, 2013). To
further expand the project, in Egypt and beyond, quantitative data
on the environmental performance of these biogas household units
2
L. Ioannou-Ttofa, S. Foteinis, A. Seifelnasr Moustafa et al.
Journal of Cleaner Production 286 (2021) 125468
rez et al., 2014), since
The construction process is not simple (Pe
the digester should be gas-tight and waterproof (Rajendran et al.,
2012). Regarding their operation, household digesters are fed on a
daily basis, mainly with manure diluted with water (Garfi et al.,
2016). However, if the volume of the compensation tank is sufficient, substrate that correspond to many days of operation can be
inserted (Nzila et al., 2012). Also, the digestate (i.e. the anaerobic
digestion residue), should be appropriately managed and disposed
of, while it could also be reused in agriculture as a biofertilizer, i.e.
act as a process by/co-product rather than as a waste (Garfi et al.,
2011). Annual maintenance typically includes biogas leakages
checks and sludge removal, however, sludge can also be removed
by gravity force of input waste, which pushes out the sludge mixed
with the digestate, as is the case here. In general, fixed-dome digesters are characterized by relatively low construction-costs and
long lifespans, since no moving or rusting parts are used (Ocwieja,
2010).
should exist, which is carried out here. Finally, the attributional LCA
(aLCA) approach was followed, which, by definition, determines the
impact of the functional unit chosen to characterize a production
system, and in which allocations are based on average data and the
relative value of the products and co-products (Rehl et al., 2012;
Weidema et al., 2018).
2.2. Functional unit
The functional unit (FU) quantifies the performance of a product/
system, in this case the fixed-dome household digester, and provides
a reference to which all input and output LCI data are normalized
(Foteinis et al., 2018). Here, the FU is production of 1 m3 biogas,
which is typically used for cooking purposes and to a lesser extent for
heating and/or lighting. Cattle manure was the raw material for
biogas production, while the digester volume was 4 m3. As advised
during field investigations a useful lifetime of 25 years was considered (assuming it operates 360 days annually to account for short
stoppages from faults or during maintenance), which is in line with
the literature (Ocwieja, 2010; Nzila et al., 2012; Garfi et al., 2016).
2.4. System boundary
The system boundary defines the parts, associated processes,
and activities of the product/system life cycle that are included in or
excluded from the analysis (FAO, 2014). Herein, a typical, for the
Egyptian case-study scenario, Chinese-type biogas household unit
is examined, with all main inputs and outputs, land use, transportation, and the relevant emissions to soil, water and air, being
included in the analysis (Fig. 2 and Table 1). End of life impacts,
including unit demolition, waste processing and recycling, are also
inside the system boundary. Along with biogas, digestate is also
generated, which can be treated as a waste or as a system by/coproduct (biofertilizer), since it is nutrient-rich. Due to its highwater content, its transport to agricultural fields that are not in
close proximity with the biogas unit is costly and logistically difficult and therefore in these cases it is typically discharged to the
environment (directly into the aquatic environment or via lagoons)
(Vu et al., 2015). Here, digestate is treated as a residue (outside the
system boundaries), however through system expansion its use as
biofertilizer is examined in the sensitivity analyses.
To model the input and output data shown in Fig. 2, i.e. the unit
processes included in the system boundary, the ecoinvent database
2.3. Description of the fixed-dome digester unit
A typical Chinese-type digester consists of a closed cylindrical
chamber with an immovable gas space (gas holder), a feedstock
inlet, and a digestate outlet, which also serves as a compensation
tank (popularly known as displacement pit) (Nzila et al., 2012). In
Egypt, digesters are constructed by easily accessible and low-cost
materials (e.g. bricks and cement) and are fully buried underground (Fig. 1), which results in very low day/night fermentation
temperature fluctuations (±2 C) (Samer, 2010). The biogas is stored
at the upper- and the waste is decomposed at the lower-part of the
chamber (Rajendran et al., 2012). Gas pressure is created due to the
difference in the level between the slurry inside the digester and
the expansion chamber (Garfi et al., 2016). After biogas production
rez et al.,
begins, the slurry is moved to the digestate outlet (Pe
2014). The biogas can be directly used or shortly stored and
therefore the digester volume varies depending on local conditions,
the amount of organic waste available, and biogas requirements
(Rajendran et al., 2012).
Fig. 1. A typical fixed-dome digester unit, constructed and operating in the Research Station of the Faculty of Agriculture at Cairo University, Giza Governorate, Egypt.
3
L. Ioannou-Ttofa, S. Foteinis, A. Seifelnasr Moustafa et al.
Journal of Cleaner Production 286 (2021) 125468
Fig. 2. System boundary (with dash lines) of the typical household biogas unit under the Egyptian case study scenario. The use of digestate is examined in the sensitivity analyses
section.
was the prefer option, with the system model that best fit the goal
and scope of this LCA study being “Allocation at the point of substitution” (APOS), previously known as “Allocation, ecoinvent
default”. APOS is, in practice, an allocation approach that employs
system expansion to avoid allocating within treatment systems. It
finds application in recycled materials, where environmental burdens from their previous life cycles are proportionally attributed, at
the point where these are used, to selected processes. Finally, the
intended audience of this work comprise researchers and decisionand policy-makers alike. The temporal (time-related) coverage of
this study spans from 2010 to present (2020). The geographical
coverage is primarily Egypt and rural areas in Middle East and
North Africa (MENA), however, results provide insight in Europe,
and further afield.
assuming local transport and that electricity is provided by Egypt’s
power grid. Sand and gravel were assumed to be extracted locally,
while the water required for the unit construction and operation
was assumed to be pumped from nearby streams/rivers, typically
the Nile river, or wells/boreholes. Hence, for water pumping, only
the petrol consumed by a typical 4.1 kW (i.e. 5.5 hp) pump was
considered, since the infrastructure for water pumping in these
areas is already in place. Furthermore, the biogas unit has a very
small impact on land use (25 m2 for 25 years), while no land-use
change was considered. After the end of its lifespan and its
dismantling, the unit was assumed to be disposed of as inert waste
for landfilling, while the recyclable parts (i.e. PVC pipes), were
assumed to be recycled. Furthermore, the biogas unit is maintained
twice annually and for each visit a transportation distance of 55 km
was ascribed, by means of a diesel-powered small-passenger car.
Regarding the animal waste produced in Egypt, the total amount
of manure (including cows, buffaloes, horses, mules, camels, sheep,
goats, chickens, ducks and turkeys) is 13.58 million tons annually,
with cattle manure having the largest contribution (~10.5 million
tons/year) (Said et al., 2013). As such, here cattle manure was
assumed to be the sole feedstock, diluted with water in a ratio of
1:1. Specifically, in the digester under study (4 m3 volume) 35 kg of
liquid fresh dung are diluted with 35 L of water, producing 1.3 m3 of
biogas which is assume to consist of 60% CH4, 35% CO2 (assumed to
be of biogenic origin) and 5% other gases (taken as nitrogen here)
(Vu et al., 2015). Furthermore, manure was considered to be a
process co-product of the livestock supply chain and therefore
emissions and resources from the livestock industry were allocated
to manure production and storage. Manure could also be considered as a waste, i.e. emissions and resource use of manure production and storage would be solely allocated to the animal farm
and only its transport from the animal farm to the biogas unit
would be allocated to the biogas unit. However, this assumption
does not fit the goal and scope of this work, while, in general,
manure is not considered as a waste, since its main function, apart
from biogas production, is as a biofertilizer.
Finally, biogas losses are attributed to leakages from nooks and
cracks in the pipping and the digester unit or to intentional releases
2.5. Life cycle inventory analysis
In the LCI stage the resources, energy flows, and relevant
emissions associated with the life cycle of the product/system under study are quantitatively defined (Ertem et al., 2016). Here,
primary LCI data for the digester’s construction and operational
phase were collected from field investigations and by consulting
with MoE (Table 1). In cases where data were not readily available,
they were obtained from the literature. The software programme
SimaPro 9.0 was employed for the environmental modelling, with
Ecoinvent 3.5 being the preferred option. In cases where LCI data
were not identified in SimaPro’s databases, literature data were
used as proxy.
Specifically, for the mud-bricks required for the construction of
the unit, ecoinvent’s data for light clay bricks (i.e. clay, straw and
water are mixed and then dried using a natural gas heater) were
considered, assuming that they are produced in Egypt, i.e. electricity inputs were assumed to be solely covered by Egypt’s energy
mix (i.e. ~8.6% hydro, ~78.4% natural gas, ~12.2% oil and ~0.8% wind,
according to ecoinvent database). Local transportation (55 km by
means of a EURO 3 emissions standard small truck) was also
considered. This is also the case for cement, where ecoinvent’s data
for Portland cement (maximum of 5% other materials) was used,
4
L. Ioannou-Ttofa, S. Foteinis, A. Seifelnasr Moustafa et al.
Journal of Cleaner Production 286 (2021) 125468
Table 1
The life cycle inventory (LCI) of a typical 4 m3 fixed-dome biogas digester constructed and operating in rural Egypt.
Construction phase
LCI data reference*
Inputs
Burnt solid mud bricks (22 10 7 cm)
Cement
Water
Sand
Gravel
Metal (screws, nails, etc.)
High-density polyethylene (HDPE) pipes e 0.5 inch
Polyvinyl chloride (PVC) pipes e 15 cm diameter
Sealants
Local transportation (delivery of raw materials)
Land area required
Output
Biogas digester
1800 bricks/3420 kg
1250 kg
663 L
5 m3
2.5 m3
1.7 kg
5 m/0.35 kg
1.7 m/2.4 kg
0.8 kg
55 km
25 m2
[1]
[1]
[1]
[1]
[1]
[1]
[1]
[1]
[1]
[1]
[1]
1 unit
[1]
Operational phase
LCI data reference*
Inputs
Water
Feedstock/substrate (cattle manure)
Manure transportation
Hydraulic Retention Time (HRT)
Temperature
Transportation (maintenance)
Outputs
Total biogas
Available biogas after losses and fugitive emissions
Digestate
Airborne emissions
CO2 (fugitive emissions)
CO2 (intentional releases)
CH4 (fugitive emissions)
CH4 (intentional releases)
Other gasses (fugitive emissions)
Other gasses (intentional releases)
Waterborne emissions (digestate leakages)
Ammonium
Potassium
Phosphorus
Emissions to soil (digestate leakages)
Ammonium
Potassium
Phosphorus
*
35 L per HRT day
35 kg per HRT day
5 km
40 days
38.5 C
55 km twice annually
[1]
[1]
[1]
[1]
[1]
1.3 m3 per HRT day
1.105 m3 per HRT day
66.5 m3 per HRT day
[1]
[1]
[1]
0.039 m3
0.0195 m3
0.078 m3
0.039 m3
0.013 m3
0.0065 m3
[1],
[1],
[1],
[1],
[1],
[1],
3.81 mg
1.12 mg
0.52 mg
[1], [5]
[1], [5]
[1], [5]
2.54 mg
0.74 mg
0.35 mg
[1], [5]
[1], [5]
[1], [5]
[2],
[2],
[2],
[2],
[2],
[2],
[3],
[3],
[3],
[3],
[3],
[3],
[4]
[4]
[4]
[4]
[4]
[4]
LCI data references: [1]: Egypt’s Ministry of Environment (MoE) [2]: Vu et al. (2015) [3]: Garfí et al. (2019) [4]: Lansche and Müller (2017) [5]: Mukhuba et al. (2018).
categories, to help understand and interpret the environmental
impacts and damages and also identify possible avoided environmental impacts (ISO 14040, 2006; ISO 14044, 2017). Several LCIA
methods are available, each one having a different approach to
modelling environmental impacts/damages. Here, a robust,
harmonized, multi-issue LCIA method was used, i.e. ReCiPe 2008
(version 1.13), which comprises both midpoint and endpoint indicators. The midpoint, or problem-oriented, and the endpoint, or
damage-oriented, approach examine different stages of the causeeffect chain to calculate environmental impacts, with the first
examining the impact earlier and the latter considering the impact
at the end of the cause-effect chain, i.e. after midpoint is reached.
Specifically, at midpoint level, environmental impacts are
translated into environmental themes, such as climate change and
human toxicity, whereas at endpoint level impacts are translated
into issues of concern, such as damage on human health, natural
environment and natural resources. The midpoint approach was
used to provide a robust understanding of the environmental impacts of the biogas system. The endpoint approach is associated
with higher levels of statistical uncertainty, due to data gaps and
assumptions stacking up along the cause-effect chain, however, it
was also used since endpoint results are easier to communicate to
decision- and policy-makers and the public (Chatzisymeon et al.,
to the atmosphere, in cases where biogas production is greater than
consumption and biogas pressure builds up in the digester (Bruun
et al., 2014; Vu et al., 2015). Here, after consulting with MoE, biogas
leakages and intentional releases were considered at 10% and 5% of
the produced biogas respectively, which is in line with the literature (Vu et al., 2015; Lansche and Müller 2017; Garfi et al., 2019).
Apart from biogas, digestate is also produced, which can be used as
a biofertilizer or discharged to the environment. Here, digestate
was assumed to be a residue of the process, i.e. outside of the
system boundary, but its effect is examined in the sensitivity analyses section. Furthermore, 5% of the produced digestate is assumed
to be lost, due to leakages through cracks in the pipping or during
its transport from the compensation tank. Half of this loss (i.e. 2.5%
of the produced digestate) is treated as a waterborne emission and
the other half as emission to the soil. Following Mukhuba et al.
(2018), digestate’s ammonium, potassium and phosphorus content was assumed to be 1.91, 0.56, and 0.26 ppm, respectively,
assuming ~14% dry matter content.
2.6. Life cycle impact assessment
The LCIA is a vital stage in LCA studies since the collected LCI
data are modelled and transformed into selected impact/damage
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L. Ioannou-Ttofa, S. Foteinis, A. Seifelnasr Moustafa et al.
Journal of Cleaner Production 286 (2021) 125468
2.7.1. Assumptions and limitations
The main assumptions/hypotheses and limitations are summarized below:
2017; Foteinis et al., 2018). At midpoint level ReCiPe use 18 impact
categories, which at endpoint level, can be aggregated into 3
endpoint damage categories, namely ‘ecosystems’, ‘resources’ and
‘human health’. Finally, the Hierarchist (H) perspective was adopted in ReCiPe, since it provides a consensus model based on common policy principles and uses the medium time frame (i.e. a 100year timeframe GWP is used) (Chatzisymeon et al., 2017).
A 70% recycling of the plastic and metal materials was considered, with the remaining 30% assumed to be discarded as inert
waste to landfill.
Irrigation water was assumed to be used both in the construction and operational stage of the unit, assuming ~40% pumping
from wells and ~60% river water.
Manure is typically generated in the close proximity of the
biogas unit, and hence a transportation distance of 5 km was
ascribed to manure, by means of a EURO 3 diesel-powered light
commercial truck.
Manure storage at the biogas digester facilities are not included
in the analysis, and thus manure storage emissions were not
considered since it is assumed that the liquid manure is directly
fed into the digester.
2.7. Sensitivity analysis
Sensitivity analyses were carried out to evaluate the effect of: (i)
the digester volume, (ii) biogas leakages and intentional releases,
and (iii) using the produced digestate as a substitute for chemical
fertilizers in agriculture. Specifically, in Egypt the volume of
household digesters varies. In this study, a base scenario dealing
with a 4 m3 household digester volume was first considered.
Thereafter, three additional geometric scales, i.e. 2, 3, and 6 m3
were considered in the sensitivity analyses section, along with an
oversized system (i.e. biogas production is greater than consumption). Furthermore, a common problem in household digesters
operating in developing countries is oversizing, which can grossly
affect their environmental sustainability. Therefore, an additional
scenario was considered, where a 6 m3 volume has been constructed in an area in which a 4 m3 volume would suffice, and in
this case the additional biogas generated is intentionally released to
the atmosphere. The LCI data for each digester are presented in
Table 2.
Furthermore, in contrast to developed countries, where biogas
losses are ~1%, in developing countries they can be as high as 60% of
the total production (Lansche and Müller, 2017). Household biogas
digesters in the developing world are not well maintained, mainly
due to lack of skilled labour in rural regions, thus facing, among
others, persistent leakage and seepage problems from cracks in the
walls caused by temperature variations. Biogas intentional releases
could also be a problem, suggesting the importance of the energy
supply reliability Brunn et al. (2014). Flaring (i.e. biogas burning)
could be a promising strategy to reduce the environmental impact
of biogas intentional releases. This practice is relatively simple to
perform, but, for health and safety, it requires some training (Vu
et al., 2015). In the sensitivity analysis both biogas leakages and
intentional releases were considered, the latter including flaring. To
this end, data from the literature regarding the GHG gas emissions
of biogas combustion (flaring) were collected, i.e. the emissions per
MJ of biogas flared are 81.5 g CO2, 57 mg CH4, 0.11 g CO and 5.4 mg
N2O (Vu et al., 2015). Airborne emissions were assumed to take
place in rural areas, i.e. low-density population areas.
Finally, through system expansion the use the produced digestate as a biofertilizer was examined. Similarly to chemical fertilizers/manure, digestate field application leads to airborne
emissions (NH3 and N2O) and nitrogen and phosphorous leaching
to soil/water, among others. However, the amount of the airborne
emissions during storage, the digestate’s final fertilizing potential
and the airborne/waterborne emissions during field application
depend on various factors, such as the type (open/close tanks or
lagoon) and duration of storage. Moreover, field emissions from
fertilizer application would take place regardless of the type of
fertilizer used. Therefore, including the emissions of the specific
storage type and duration, as well as emissions from field application, are outside the scope of this study. Hence, only the substitution factors of synthetic fertilizers were considered, examining
two popular storage types in Egypt, i.e. lagoon and open container
storage. When the digestate is stored in open containers the mean
avoided mass of N, P, K fertilizer would amount to 0.88, 0.030 and
0.64 kg, while for lagoon storage is much lower, at 0.29, 0.015, and
0.32 kg per m3 of liquid digestate, respectively (Styles et al., 2018).
3. Results and discussion
3.1. ReCiPe at midpoint level
First, ReCiPe results at midpoint level are presented. To
accommodate an easier analysis/discussion, the system was divided
into its two main sub-systems: (a) the construction phase, which
also includes the unit’s disposal/recycling after the end of its useful
life, and (b) the operational phase, which also includes biogas/
digestate leakages. In Fig. 3, the influence of each of the two main
sub-systems into ReCiPe’s midpoint impact categories is shown
(Characterisation). The main contributor across categories is the
operational phase, while the construction phase has a much
smaller contribution. This was expected since: (i) in general the raw
materials required for the construction of the biogas digester are
not suspected to be carcinogenic, mutagenic or toxic to reproduction (EC, 2006; NTP, 2016), and (ii) the biogas unit exhibit an overall
high lifespan (25 years) and hence only an insignificant amount of
its total environmental footprint is allocated per FU. It should be
mentioned that the recycling of plastics and metals, after the end of
the unit’s life span, was also considered. However, due to their
overall low mass, it was identified that they had a very small
contribution across ReCiPe’s midpoint impact categories.
The total GWP for the construction phase was found to be
1945 kg CO2eq (4 m3 digester). This is lower than Rahman et al.
(2017) estimation (2838 kg CO2eq.) for a 3.2 m3 fixed-dome
biogas plant in Bangladesh, as well as Wang and Zhang (2012)
estimation for the construction of an 8 m3 biogas digester in
China (2357 kg CO2eq). On the other hand, it was higher than the
estimation of Hou et al. (2017), where the annualized GHG emissions from the construction of an 8 m3 typical Chinese rural
household biogas unit was found to be up to 305 kg CO2eq. However, different assumptions/limitations and LCIA methods were
used, while in our case study we assumed that the main raw materials (bricks, cements), were produced locally using Egypt’s energy mix which grossly depends on natural gas (~78.4%), having
thus significantly lower emissions compared to other fossil fuels,
such as coal or oil.
Regarding the operational phase, the much higher contribution
across midpoint impact categories, compared to the construction
phase, can be mainly attributed to: (i) the feedstock that is daily fed
to the digester (primarily to manure and to a much lesser extent to
water) and (ii) the airborne emissions, i.e. biogas leakages and
intentional releases. For example, the GWP of the operational phase
was found to be 2.52 kg CO2 eq. instead of 0.19 kg CO2 eq. produced
during the construction of the Egyptian household digesters.
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Table 2
LCI data for the 2, 3, and 6 m3 digesters, as well as for the oversized digester, used to identify the influence of the digester volume on the environmental sustainability of the
Egyptian household biogas production unit.
Construction phase
Digester volume
Inputs
Burnt solid mud bricks (22 10 7 cm)
Cement (kg)
Water (L)
Sand (m3)
Gravel (m3)
Metal (screws, nails, etc.) (kg)
HDPE pipes (0.5 inch) (m)
PVC pipes (15 cm diameter) (m)
Sealants (kg)
Local transportation (km)
Land area required (m2)
Output
Biogas digester
2 m3
3 m3
6 m3
6 m3 (oversized)a
1000 bricks
850
451
3
1
1.7
5
1.7
0.8
55
16
1200 bricks
950
504
4
1.5
1.7
5
1.7
0.8
55
20
2600 bricks
1750
928
8
3.5
1.7
5
1.7
0.8
55
35
2600 bricks
1750
928
8
3.5
1.7
5
1.7
0.8
55
35
1 unit
1 unit
1 unit
1 unit
17 L/HRT day
17 kg/HRT day
5
40
38.5
55 km twice annually
25 L/HRT day
25 kg/HRT day
5
40
38.5
55 km twice annually
50 L/HRT day
50 kg/HRT day
5
40
38.5
55 km twice annually
50 L/HRT day
50 kg/HRT day
5
40
38.5
55 km twice annually
0.7 m3/HRT day
0.595 m3/HRT day
32.3 L/HRT day
1 m3/HRT day
0.85 m3/HRT day
47.5 L/HRT day
2 m3/HRT day
1.7 m3/HRT day
95 L/HRT day
2 m3/HRT day
1.7 m3/HRT day
95 L/HRT day
0.021
0.0105
0.042
0.021
0.07
0.0035
0.03
0.015
0.06
0.03
0.01
0.005
0.06
0.03
0.12
0.06
0.02
0.01
0.06
0.21
0.12
0.42
0.02
0.07
61.693
18.088
8.398
90.725
26.6
12.35
181.45
53.2
24.7
181.45
53.2
24.7
1.85079
0.54264
0.25194
2.72175
0.798
0.3705
5.4435
1.596
0.741
5.4435
1.596
0.741
Operational phase
Inputs
Water
Feedstock/substrate (cattle manure)
Manure transportation (km)
Hydraulic Retention Time (HRT) (days)
Temperature (oC)
Transportation (maintenance)
Outputs
Total biogas production
Available biogas (after losses and fugitive emissions)
Digestate
Airborne emissions
CO2 (fugitive emissions) (m3)
CO2 (intentional releases) (m3)
CH4 (fugitive emissions) (m3)
CH4 (intentional releases) (m3)
Other gasses (fugitive emissions) (m3)
Other gasses (intentional releases) (m3)
Waterborne emissions (digestate leakages)
Ammonium (mg)
Potassium (mg)
Phosphorus (mg)
Emissions to soil (digestate leakages)
Ammonium (mg)
Potassium (mg)
Phosphorus (mg)
a
6 m3 digester installed in an area where a 4 m3 digester would be sufficient.
which biogas leakages were the main contributor. In the study of
Wang et al. (2018), the GWP of the 8 m3 household biogas digesters
was found to be 366.87 kg CO2eq. Finally, according to Gabisa and
Gheewala (2019), the operation of 4500 fixed-dome biogas digesters in Ethiopia was found to have a potential of emitting 1.5 Gg
CO2 per year, and accordingly each household unit annually emits
0.33 tonnes of CO2eq. As shown from the results above, and as
already highlighted, the comparison of the results of different LCA
studies cannot be direct, since the goal and scope, the functional
unit, the size and the characteristics of the digesters under study, as
well as the impact assessment methods, the assumptions, the energy mix and the geographical conditions could significantly differ.
Overall, this study focuses on the environmental sustainability of
the household biogas units operated in Egypt (Table 3) and is sufficient to draw recommendations to the decision makers from an
environmental point of view, while more research is required on
the socio-economic and policy perspective of the biogas sector.
In order to gain a deeper understanding of the relative magnitude of the scores shown in Table 2, results were normalized using
ReCiPe’s Hierarchist version and applying the normalization values
of the world. Even though normalization, where results are
ReCiPe’s 18 midpoint impact categories, along with the score for
the construction and operational phase and the total score, are
shown in Table 3. In terms of carbon emissions, it was found that
the GWP for producing 1 m3 of biogas under Egypt’s conditions
amounts to 2.72 kg CO2eq. The GWP of the Egyptian digester under
study is in agreement with the results of the study of Vu et al.
(2015), in which the GWP of the overall handling of pig manure
in small-scale Vietnamese digesters was 3.2 kg CO2eq. Almost twice
higher GWP values were reported by Garfi et al. (2019) for smallscale digesters (tubular plastic) located in Colombia. According to
rez et al. (2014), the GWP of the fixed-dome digesters (0.12 m3
Pe
biogas/m3 digester day) in Latin America was found to be 0.24 kg
CO2eq./m3 biogas. Significant lower (0.02 kg CO2/MJ or ~0.68
CO2eq./m3 biogas when assuming a conversion factor of 34 MJ/m3
biogas (IRENA, 2016)) were also the CO2 emissions of a family-size
Chinese-type digester in rural China (Zhang et al., 2013). Singh et al.
(2014) estimated that up to 0.0105 kg CO2eq./MJ or ~0.68 CO2eq./m3
biogas, are produced by a household fixed-dome digester in India
(2 m3). Hou et al. (2017) found that the overall GHG emissions from
the construction and operation of an 8 m3 Chinese household
biogas unit were in the range of 706e2794 kg CO2 per year, from
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Journal of Cleaner Production 286 (2021) 125468
Fig. 3. The influence of the construction and operational phase of the household biogas digester to ReCiPe’s 18 midpoint impact categories. FU: 1 m3 of biogas production.
Table 3
ReCiPe midpoint scores (Characterisation) per FU, i.e. production of 1 m3 biogas from a 4 m3 household digester under rural Egypt’s conditions.
Impact category
Unit
Total
Construction phase
Operational phase
Climate Change (CC)
Ozone Depletion (OD)
Terrestrial Acidification (TA)
Freshwater Eutrophication (FE)
Marine Eutrophication (ME)
Human Toxicity (HT)
Photochemical Oxidant Formation (POF)
Particulate Matter Formation (PMF)
Terrestrial Ecotoxicity (TE)
Freshwater Ecotoxicity (FEC)
Marine Ecotoxicity (MEC)
Ionising Radiation (IR)
Agricultural Land Occupation (ALO)
Urban Land Occupation (ULO)
Natural Land Transformation (NLT)
Water Depletion (WD)
Mineral Resource Depletion (MRD)
Fossil Depletion (FD)
kg CO2eq.
kg CFC-11 eq.
kg SO2 eq.
kg P eq.
kg N eq.
kg 14DCB eq.
kg NMVOC
kg PM10 eq.
kg 14 DCB eq.
kg 14DCB eq.
kg 14DCB eq.
kBq U235 eq.
m2a
m2a
m2
m3
kg Fe eq.
kg oil eq.
2.72 Eþ00
4.02E-08
5.54E-03
4.55E-04
5.44E-03
1.33E-01
2.95E-03
1.74E-03
4.42E-03
7.41E-03
6.05E-03
2.56E-02
6.29E-01
1.44E-02
1.37E-03
4.95E-02
3.08E-02
1.11E-01
1.96E-01
1.27E-08
6.58E-04
2.21E-05
5.15E-05
2.89E-02
6.88E-04
2.90E-04
3.43E-05
7.08E-04
6.73E-04
5.83E-03
1.23E-02
4.00E-03
4.30E-05
2.96E-03
8.22E-03
4.01E-02
2.52 Eþ00
2.76E-08
4.88E-03
4.33E-04
5.39E-03
1.04E-01
2.27E-03
1.45E-03
4.39E-03
6.71E-03
5.38E-03
1.98E-02
6.17E-01
1.04E-02
1.41E-03
4.65E-02
2.26E-02
7.09E-02
cement apart from requiring a relatively large energy input are also
responsible for air pollution (e.g. particulate matter emissions),
while cement production is an energy-intensive process by nature.
The human toxicity (HT) impact category is also affected by
airborne emissions from sand and cement mining and processing.
Furthermore, raw material transportation is achieved by dieselpowered trucks, thus consuming fossil fuels. Although the comparison among LCA studies cannot be direct, as mentioned before, it
is noted that the environmental impacts for the construction of the
household digesters in Egypt were found to be comparable than
those of others available in the scientific literature. Specifically, the
HT impact potential in the study of Lansche and Müller (2017) for
the construction of an 8 m3 fixed-dome biogas digester in Ethiopia,
was 0.9 kg FCB eq. and the terrestrial ecotoxicity (TEC) potential
was 0.15 g DCB eq. per MJ of cooking heat energy, which are
comparable, however quite higher than the results of our study.
Specifically, in this LCA study for the construction of a 4 m3 fixeddome digester in Egypt, HT was 0.0289 kg FCB eq. and TEC was
transformed by dividing impact categories with corresponding
reference values (Masindi et al., 2018), is not a compulsory step of
the LCIA it was employed to identify the impact categories that are
mainly affected and identify their relative magnitude. As shown in
Fig. 4 the midpoint impact category, MEC, closely followed by FEC,
FE, and then TE, ME, and HT are mainly affected. The remaining
categories are affected to a lesser extent, at least an order of
magnitude lower compared to MEC.
Regarding the construction phase, the contribution to the (eco)
toxicity (MEC, FEC and HT) and eutrophication (FE and to a much
lower degree ME) impact categories is mainly attributed to raw
material mining and processing. The material with the biggest
contribution is burnt solid bricks, closely followed by cement, and
to a lesser extent to sand and gravel mining. For brick production,
clay needs to be extracted and transported, which requires fossil
fuels, typically diesel, while the brick drying process is energy
intensive. Here, drying by heat was considered, which was achieved
by burning natural gas. Sand and gravel mining, and particularly
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Journal of Cleaner Production 286 (2021) 125468
Fig. 4. Normalized scores for the construction and operation phase of the biogas unit. Functional unit 1 m3 biogas, using ReCiPe’s Hierarchist version with the normalization values
of the world.
0.0343 g 1.4 DCB eq. per m3 of produced biogas, respectively.
Concerning the operational phase, the much larger contribution
on the (eco)toxicity impact categories (Fig. 4) is mainly attributed to
manure production and transportation. Manure is a co-product of
the livestock supply chain and as such its production requires animal feed (i.e. maize, soybean and grass), which requires pesticides
and fertilizers that lead to emissions directly affecting the (eco)
toxicity and eutrophication impact categories, respectively. Their
production as well as the feed cultivation process are also energyintensive, while the infrastructure used for manure production
requires energy and metal input (e.g. metals to build the housing
system for the cattle and electricity to run it). This is also the case
for the tractors used for the feed cultivation and the truck for
manure transportation, which both require large metals input for
their production and therefore exhibit an overall high contribution
to these categories. Furthermore, biogas losses and intentional releases directly affected the (eco)toxicity impact categories.
It should be noted that fossil fuel extraction, refining, transportation and burning release heavy metals, polycyclic aromatic
hydrocarbons and other toxic compounds to the environment
(Ioannou-Ttofa et al., 2017), affecting directly the (eco)toxicity
impact categories. The relatively high score on the eutrophication
impact categories can be traced back to fossil fuel mining activities,
where phosphorus and sulphate emissions, among others, cause
eutrophication (Masindi et al., 2018). Furthermore, fossil fuel
burning emits NOx, which directly affect ME, while phosphate
emissions from fossil fuel mining directly affect FE (Ioannou-Ttofa
et al., 2017). This is also the case with the digestate emissions to
soil and water, which directly affect FE and ME. Furthermore, during the fertilizing the crops used as the animal feed, excess phosphorus and nitrogen from can end up to receiving waterbodies,
thus polluting freshwater, including groundwater, and marine
ecosystems leading to eutrophication. The lower score of ME (nitrogen enrichment of seawater), compared to FE (phosphorus
enrichment of freshwater) impact category, can be attributed to the
fact that freshwater ecosystems are less resilient than marine
ecosystems to eutrophication stresses (Chatzisymeon et al., 2016).
Lastly, the scores of the less affected categories, such as CC, TA, NLT,
MRD, and FD, can be similarly traced back to the abovementioned
reasons, but yield a much smaller degree when using the normalization values of the world.
Regarding the overall environmental impacts of the construction and operation of a household biogas unit in developing
countries, it was shown that the Egyptian units under study have
lower, similar or higher impacts to other units examined in the
literature. Specifically, the marine eutrophication (ME) and freshwater eutrophication (FE) potential for the operation of the Vietnamese biogas digesters, was 0.82 kg N eq. and 0.551 kg P eq. per FU
(i.e. FU: treatment of 100 kg of solid pig manure and 1000 kg of
liquid pig manure), respectively (Vu et al., 2015). In our case study,
ME and FE potentials were equal to 5.44E-03 g N eq. and 4.55E04 kg P eq. per m3 biogas (i.e. per less than 35 kg of liquid manure),
respectively. However, the main contribution to the eutrophication
impact categories in the study of Vu et al. (2015) was the discharge
of the digestate. Furthermore, FE for fixed-dome digesters operating in Pakistan was 0.023 kg P eq. per FU (4 tonnes/day) (Yasar
et al., 2017), and for the 8 m3 household digesters operating in
China was 1.92 kg P eq. per FU (2136 tonne/yr of manure) (Wang
et al., 2018), while in our case study FE potential was 0.455
103 kg P eq per m3 of biogas. The human toxicity (HT) potential,
according to Wang et al. (2018) for the household digesters operated in China, was 0.06e0.15 kg 1.4-DCB per tonne manure, while in
this work was 0.133 kg 1.4-DCB eq. per FU. On the other hand, the
photochemical oxidation formation (POF) potential, was 1.16 kg
NMVOC for the digesters in China (Wang et al., 2018), while here it
was 2.95 103 kg NMVOC per FU, with VOC emissions largely
contributing to this score, followed by CH4 emitted during the unit
operation. The differences between LCA studies and the results of
this study (Table 3) were expected since different FUs, system
boundary, and LCIA methods are used. Furthermore, in our study
the electricity input required for raw material extraction/processing, such as for cement production, were solely covered by Egypt’s
energy mix, which is based on natural gas and not coal/oil that
significantly increase environmental impacts.
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Journal of Cleaner Production 286 (2021) 125468
traced back to bricks and cement. The underlying reason for this
contribution is raw material extraction and processing, which are
both energy-intensive processes and are responsible for airborne
emissions.
Finally, when comparing the biogas produced in Egypt with
biogas produced in Europe (Switzerland), large differences were
observed (Fig. 5), as was expected. Specifically, the process of biogas
production in Europe is included in ecoinvent database, having
time coverage 2009e2016, with LCI data collected from 18
centralized and decentralized agricultural biogas plants in
Switzerland and complemented with literature data. Mean values
are used for the European biogas plant, considering as input livestock manure (cattle and pig slurry and cattle manure), while the
treatment process includes manure storage and 10% of relevant
emissions. It was found that for the production of 1 m3 of biogas,
the total environmental footprint of the typical biogas plant operating in Europe is about 53% lower compared to biogas produced
from a household biogas digester under Egypt’s conditions (i.e. 77
mPt in Europe instead of 160.1 mPt in Egypt) (Fig. 5). Even though
the comparison cannot be direct, it provides context and gives a fair
estimate of the environmental sustainability of the system under
study compared to typical biogas plants in Europe and in the
developed world, in general. This can be of particular importance to
decision- and policy-makers, since it suggests the large possibilities
of this technology to provide green and renewable energy locally.
3.2. ReCiPe at endpoint level and comparison with biogas
production in Europe
In Fig. 5, the total environmental footprint for the production of
1 m3 of biogas under Egypt’s conditions (i.e. 160.1 mPt) is given,
along with a comparison with a typical large biogas plant in Europe
(i.e. Switzerland), which also uses cattle manure as feedstock and
has the same FU with the Egyptian scenario. It should be noted that
the comparison with a large biogas plant in Europe is only provided
for context, since the comparison cannot be direct; but it can provide insight to decision- and policy-makers.
From the 160.1 mPt per m3 biogas (Fig. 5), the main contributor
was the operational phase (~89.1%), while the construction phase
had a much smaller contribution (~10.9%). Regarding the operational phase 38.8% of this score is attributed to manure, 36.9% to
biogas leakages, 18.4% to biogas intentional releases (i.e. biogas
losses amount to 55.3% of the total environmental footprint of the
operational phase), 4.84% to the personnel transportation required
for the maintenance of the biogas unit, and 1.01% to irrigation water
used to dilute the manure. As regards the construction phase, 31.9%
of this score is attributed to the bricks, 43.5% to cement, 7.73% to
sand, 5.47% to gravel, while pipping, adhesives and metals has each
less than 1% contribution. Finally, 10% of the construction phase is
attributed to the disposal of the biogas unit, after the end of its
useful lifespan.
Concerning the large contribution of ‘human health’ damage
category (Fig. 5), this is mainly attributed to biogas losses and
intentional releases. To a lesser extent, emissions arising from
manure production and transportation (e.g. emissions arising from
diesel combustion in the tractor and truck) and emissions from
water pumping, also contribute to this damage category. Also, fossil
fuel burning lead to its depletion, thus directly affecting the damage
category ‘resources availability’. The category ‘damage to ecosystems’ is mainly affected by the midpoint impact categories of (eco)
toxicity and eutrophication, which can be traced back to manure
production (for the reasons described above), and fossil fuel
burning. Regarding the construction phase, its main contribution is
in ‘human health’ and ‘resources’ damage categories, which can be
3.3. Sensitivity analyses
First the effect of the digester volume on the system’s environmental sustainability was examined (as presented in Table 2).
Fig. 6 suggests that the digester volume has little effect on the
system’s environmental performance, provided that the digester
sizing is fit for purpose. The 2, 3 and 4 m3 digester volume have
almost identical environmental footprints (<2% differences), while
only when scaling up the process to a 6 m3 digester the effect of the
digester volume is somewhat noticeable (up to 6% reduction). On
the other hand, oversizing grossly affects the system’s environmental sustainability (~91 increase), since it leads to large biogas
Fig. 5. Comparison of the environmental performance of a household biogas unit in Egypt with a large biogas plant operating in Europe per functional unit (1 m3 biogas), when
using ReCiPe Hierarchist version with world normalization and average weighting set.
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Journal of Cleaner Production 286 (2021) 125468
Fig. 6. e Sensitivity analysis dealing with the biogas digester total volume.
Table 4
Best- and worst-case scenarios with regard to biogas leakages, intentional releases and flaring.
Scenario
Fugitive emissions
Intentional releases
Base scenario
Biogas leakages best-case scenario
Biogas leakages worst-case scenario
Biogas Intentional releases best-case scenario
Biogas Intentional releases worst-case scenario
Biogas Flaring base scenario
Biogas Flaring worst-case scenario
Biogas best-case scenario
10%
1%
60%
10%
10%
10%
10%
1%
5%
5%
5%
0%
60%
5%
60%
0%
Fig. 7. Sensitivity analyses dealing with: (a) biogas leakages; (b) intentional releases; (c) flaring; and (d) best scenario compared to the base scenario.
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Journal of Cleaner Production 286 (2021) 125468
intentional releases (Table 4). This suggest the importance of
properly sizing biogas units before installation.
For biogas leakages and intentional releases/flaring, the best- and
worst-case scenarios shown in Table 4 were examined. For biogas
leakages results suggest that the best-case scenario achieves ~36%
reduction and the worst-case scenario a six-fold increase (1030 mPt
instead of 160.1 mPt) in the total environmental footprint (Fig. 7a).
This very large increase is mainly attributed to two reasons: firstly, to
the large amounts of the biogas emitted directly to the atmosphere
(mainly to CH4 and to a smaller degree to CO airborne emissions),
and, secondly, to the fact that a much lower amount of biogas is
available in the worst-case scenario, compared to the base scenario,
since most of the produced biogas is lost to the atmosphere.
In the intentional releases best-case scenario (i.e. no intentional
releases) (Table 4) the system’s environmental footprint is reduced
by ~19% (132.8 mPt), while in the worst-case scenario (i.e. 60% of
biogas being directly released to the atmosphere) a seven-fold increase (~1120 mPt instead of 160.1 mPt) is observed (Fig. 7b). When
flaring is examined a 19% reduction in the base scenario (133.9 mPt
instead of 160.1 mPt in the base scenario) and a 74% reduction in the
worst-case intentional releases scenario is observed (Fig. 7c). The
results of our study are in line with those of the available literature,
where flaring, instead of intentionally releasing it, significantly
improves the environmental profile of the biogas systems (Bruun
et al., 2014; Vu et al., 2015).
As expected, the environmental sustainability is optimised
when the two best-case scenarios are combined (Table 4) (i.e. 1%
biogas leakages and no intentional releases), since the system’s
total environmental footprint reduces by ~60% compared to the
base scenario (Fig. 7d). Although this scenario is hypothetical, it
highlights the many possibilities of this technology, since results
suggest that the household biogas digester technology operating in
rural Egypt could be on the same environmental level with larger
units operating in Europe.
Overall, it was identified that biogas leakages and intentional
releases grossly affect the system’s environmental sustainability,
with leakages having a larger influence due to the potentially larger
volume of biogas escaping to the environment in the respective
worst-case scenario. However, it appears that flaring can largely
reduce the environmental impact of biogas intentional releases, by
more than three-fold in the worst-case scenario.
Finally, regarding the use of digestate as biofertilizer, it was
found that when applying lagoon stored digestate the system’s
environmental footprint would be reduced by ~14% (138 mPt
instead of 160.1 mPt on the base scenario). Due to its significantly
larger fertilizing potential, the open container stored digestate
leads to a larger reduction (~38%) (98 mPt). Even though more
research is required on the potential of digestate as a biofertilizer,
these preliminary results are suggestive of its potential to drastically reduce the environmental impacts of household biogas technology operating in rural Egypt.
the system’s environmental sustainability. At midpoint level, the
100-year GWP for producing 1 m3 of biogas under Egypt’s conditions was 2.72 kg CO2eq/m3, while at endpoint level its total
environmental footprint was 160.1 mPt/m3. As with all LCA studies,
our study also includes several limitations associated with the
geographical and the time related coverage, the technology under
consideration, the LCIA method, and the functional unit, among
others. Therefore, even though the comparison with other studies
cannot be direct, the identified environmental impacts of the
Egyptian household digesters were found comparable with the
results already included in the existing body of knowledge. The
total environmental footprint was also twice as high as that of
biogas produced by large biogas plants operating in Europe.
Through sensitivity analyses it was identified that the digester
volume plays a small role on the system’s environmental sustainability, however, the system was found to be very sensitive to
biogas leakages and intentional releases, the latter commonly
encountered in oversized digesters. This suggests the need to
properly sizing digesters according to local needs, while flaring the
excess biogas, instead of releasing it, can largely reduce environmental impacts and improve the system’s environmental profile.
Properly build digesters and important aspects, such as regular
cleaning and proper maintenance of the unit (i.e. check for cracks),
which are often neglected by the farmers, affect not only the digester’s productivity but also increasing its overall environmental
footprint. Finally, it appears that digestate use as biofertilizer could
be a promising strategy to improve the environmental sustainability of the system, but more research is needed to gain a better
understanding about the potential application of the digestate, as
well as of the co-digestion of manure with other agricultural waste
towards minimizing the system’s environmental sustainability and
also maximizing biogas production.
From the above, it is concluded that government support for the
deployment of appropriately sized and maintained household
biogas units, could significantly improve operating efficiency and
economic viability and also reduce the environmental impacts of
biogas units. Furthermore, widespread adoption of more precise
nutrient management planning would ensure that digestate is
utilized efficiently, contributing thus significantly to system’s
environmental sustainability. Finally, it is quite crucial that local
decision- and policy-makers should create incentives for the construction and installation of additional household digesters in the
rural Egypt and introduce legislation about their proper installation
and operation and maintenance.
CRediT authorship contribution statement
Lida Ioannou-Ttofa: Conceptualization, Methodology, Investigation, Validation, Writing - original draft, Project administration.
Spyros Foteinis: Conceptualization, Methodology, Software,
Investigation, Validation, Writing - original draft. Amira Seifelnasr
Moustafa: Investigation. Essam Abdelsalam: Investigation.
Mohamed Samer: Investigation, Supervision. Despo Fatta-Kassinos: Writing - review & editing, Supervision, Funding acquisition.
4. Conclusions and recommendations
Results show that the majority of the environmental impacts of
the household biogas technology operating under rural Egypt’s
conditions are mainly attributed to the system’s operational phase,
and to a lesser extent to its construction phase. The underlying
reason for the overall low contribution of the construction phase
lies in its high lifespan and relatively low environmental footprint.
The first original feature of this work is that the environmental
impacts of the operation of the Egyptian household digesters are
mainly attributed to the feedstock used (i.e. manure) and to the
system’s airborne emissions. Waterborne emissions and emissions
to soil, from digestate’s leakages, had an overall low contribution on
Declaration of competing interest
The authors declare that they have no known competing
financial interests or personal relationships that could have
appeared to influence the work reported in this paper.
Acknowledgments
This work was prepared in the framework of the BIOGASMENA
project (KOINA/ERANETMED/0316/01), financed by the Cyprus
12
L. Ioannou-Ttofa, S. Foteinis, A. Seifelnasr Moustafa et al.
Journal of Cleaner Production 286 (2021) 125468
Research and Innovation Foundation (DESMI 2009e2010). This
research was supported by EranetMed Biogasmena (Project ID
72e026), which aims to demonstrate dry fermentation and optimize biogas technology for rural communities in MENA (Middle
East and North Africa) region. The authors would like to acknowledge the Science and Technology Development Fund (STDF) of
Egypt for funding the survey, where this study was conducted in
the framework of the research project no. 30278.
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