Journal of Cleaner Production xxx (2014) 1e10
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Journal of Cleaner Production
journal homepage: www.elsevier.com/locate/jclepro
Life cycle energy and greenhouse gas assessment of the co-production
of biosolids and biochar for land application
Leslie Miller-Robbie a, Bridget A. Ulrich a, Dotti F. Ramey a, Kathryn S. Spencer b,
Skuyler P. Herzog a, Tzahi Y. Cath a, Jennifer R. Stokes c, Christopher P. Higgins a, *
a
NSF Engineering Research Center ReNUWIt, Department of Civil and Environmental and Engineering, Colorado School of Mines, Golden CO 80401, USA
Clackamas County Water Environment Services, Oregon City, OR 97045, USA
NSF Engineering Research Center ReNUWIt, Department of Civil and Environmental and Engineering, University of California, Berkeley, Berkeley,
CA 94720, USA
b
c
a r t i c l e i n f o
a b s t r a c t
Article history:
Received 20 August 2014
Received in revised form
2 December 2014
Accepted 12 December 2014
Available online xxx
The co-production and co-application to land of biochar and biosolids may make conventional wastewater treatment more sustainable; therefore, this study provides a first life cycle evaluation of the colocation of a pyrolysis (biochar production) plant with a municipal wastewater treatment plant. Life
cycle energy use and greenhouse gas emissions are evaluated for the co-production of these two materials. Three US national scenarios are included: biosolids only produced and disposed of as is the
current common practice in the US (landfilled, incinerated, and land applied); biosolids only produced
but not land applied; and co-production of biosolids and biochar and their disposal as currently practiced. A case study is also presented with slightly different disposal practices. Large contributors to life
cycle greenhouse gas emissions are from the disposal of biosolids in landfills and incinerators, making
alternatives like land application attractive for mitigating greenhouse gas emissions. The addition of
biochar production adds little to the overall energy use but provides substantial (26%) reduction in
greenhouse gas emissions for the national case, largely due to the recalcitrant carbon storage in biochar.
Because biochar production is an energy positive process, incorporation of the energy co-products into
the plant is explored for producing high quality Class A biosolids. The addition of biochar to biosolids is
shown to be more sustainable from a co-production standpoint. This study contributes a benchmark
analysis in evaluating the co-production of biochar and biosolids for potential environmental, economic,
and health benefits and the results are useful for utility planning.
© 2014 Elsevier Ltd. All rights reserved.
Keywords:
Wastewater treatment
Energy recovery
Pyrolysis
Life cycle assessment
Contaminants of emerging concern
Class A biosolids
1. Introduction
The long-term sustainability of conventional wastewater treatment may be threatened by potential environmental and human
health risks associated with biosolids at their end-of-life. Land
application of biosolids is a widespread practice on a global scale
(Chang et al., 2002). Currently in the United States, 55% of biosolids
are land applied with the remaining 45% disposed through incineration and landfilling (North East Biosolids and Residuals
Association (NEBRA), 2007). Land application allows for the nutrients in biosolids to be recycled back into the environment, displacing energy-intensive synthetic fertilizers that unnaturally fix
* Corresponding author. Tel.: þ1 720 984 2116; fax: þ1 303 273 3413.
E-mail address: chiggins@mines.edu (C.P. Higgins).
high levels of nitrogen (N) (Vitousek et al., 1997) and increasing
long-term reserves of carbon in soils. However, chemicals in biosolids may contribute to contamination of various environments.
Domestic wastewater contains thousands of synthetic organic
and inorganic chemicals which remain in the biosolids following
wastewater treatment. Levels of metals and pathogens in biosolids
are regulated in the US (United States Environmental Protection
Agency, 1994), but other pollutants, including synthetic organic
chemicals, are not. As contaminants of emerging concern are being
studied (Kinney et al., 2006; Sepulvado et al., 2011), further restrictions on biosolids land application may be implemented to
protect surrounding water and soil quality from contamination
(Camp Dresser and McKee Inc., 2011). In some European countries,
agricultural land application of biosolids has been restricted due to
concerns about related health risks (European Commission, 2002;
Fytili and Zabaniotou, 2008). To reap the benefits of biosolids
http://dx.doi.org/10.1016/j.jclepro.2014.12.050
0959-6526/© 2014 Elsevier Ltd. All rights reserved.
Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050
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L. Miller-Robbie et al. / Journal of Cleaner Production xxx (2014) 1e10
land application in the future, one strategy is to co-apply a material
that sorbs contaminants to mitigate their potential adverse environmental impacts.
A potential sorption material for many contaminants of
emerging concern is biochar, a charcoal-like material made from
pyrolysis of agricultural residues (e.g., corn stover, wood wastes,
yard trimmings, etc.). Pyrolysis stabilizes the carbon taken up by
plants and stores it in biochar, providing a carbon-sink (Meyer et al.,
2011; Sparrevik et al., 2013; Woolf et al., 2010). Biochar has been
used throughout the world to increase soil health and fertility
(Glaser et al., 2001; Kimetu et al., 2008). It has been shown to inlanger
crease soil nutrient retention and nutrient availability (Be
et al., 2004; Major et al., 2010; Prendergast-Miller et al., 2014).
There is also growing evidence of the sorption capacity of biochar
for herbicides and metals (Park et al., 2011; Spokas et al., 2009;
Uchimiya et al., 2010). The production of biochar is a net energy
positive process, as syngas, bio-oil, and thermal heat are also produced during pyrolysis (Laird et al., 2009).
Although the life cycle environmental tradeoffs of biosolids and
biochar have been studied separately (Peters and Rowley, 2009;
Roberts et al., 2010), their combination has not been investigated
prior to this study. In addition to the life cycle assessment (LCA) that
includes three national-level scenarios and four scenarios applied
to a real-world case study utility, the integration of energy between
the pyrolysis plant and the wastewater treatment plant (WWTP) is
explored.
2. Materials and methods
Process-based LCA methods were used for this LCA model.
Material flow data were collected from both primary and secondary
sources. Only energy use (EU) and greenhouse gas (GHG) emissions
were included as impact categories. The impact assessment was
carried out using global warming potential factors from IPCC's 2007
Report (Intergovernmental Panel on Climate Change (IPCC), 2007).
Global warming potential (in terms of carbon dioxide equivalents
(CO2e)) was calculated for a 100-year time horizon. Carbon
sequestration potential was only evaluated for the recalcitrant
fractions of biochar and biosolids that have been shown to be stable
for 100 years or longer (Recycled Organics Unit, 2007; Woolf et al.,
2010). All infrastructure included in this study was assumed to have
a 50-year lifetime (Horvath and Stokes, 2011). While all analysis is
compliant with ISO 14040 standards (International Organization
for Standardization, 2006), an extensive review by two LCA contributors was carried out in lieu of a review by an independent
practitioner. Tables S1 and S2 in the supplementary information
detail the data sources and emission factors employed throughout
this study.
2.1. Goal and scope
This study uses LCA as a preliminary step to evaluate the EU and
GHG emissions of a new approach in which yard trimmings are
delivered to a biochar plant co-located with a WWTP. Upon
completion, the two products are mixed and transported for land
application in the model. This scenario is compared with other
conventional and non-conventional scenarios. The study boundaries included all materials and energy inputs and outputs, as well
as the energy to produce and transport all inputs and outputs,
starting with waste biochar feedstock (domestic yard trimmings)
and wastewater solids or biosolids input (Fig. 1). The WWTP prior
to wastewater solids or biosolids input (including the anaerobic
digester itself) was outside the study boundary. Building materials
associated with the WWTP were not included because they are
Fig. 1. Process Map of scenarios in this study for the national scenarios. Yellow signifies processes included in all scenarios; blue signifies processes omitted for scenarios without
land application (N2); red signifies processes omitted for scenarios without biochar production (N1 and N2); green signifies a process omitted for scenarios with biochar production
(N3). The only major differences for the case study scenarios are the addition of a waste-to-energy facility and the omission of incineration of solids. The symbol
denotes
emissions as an output from the adjacent process and
denotes avoided emissions. (For interpretation of the references to colour in this figure legend, the reader is referred to the
web version of this article.)
Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050
L. Miller-Robbie et al. / Journal of Cleaner Production xxx (2014) 1e10
necessary regardless of the addition of a pyrolysis plant. However,
all construction materials associated with the pyrolysis plant were
included. The pyrolysis plant was allocated solely to biochar and
not to co-products because the potential uses for syngas and bio-oil
were outside the scope of the LCA. This assumption is conservative
because the inclusion of plant construction in the biochar analysis
increases EU and GHG impacts. In reality, when co-products are
included, the scenarios that include biochar will be even more
beneficial. As shown in Fig. 1, utilizing waste heat is within the
scope because it is used to dry feedstock; however, it cannot be
transported to offset energy use in other systems, and related GHG
emissions are assumed to be negligible. Energy to collect, transport,
and shred biochar feedstock was included, but not the energy to
cultivate the yard material because feedstock production is inevitable. Yard trimmings not used for biochar production in the national scenarios are either composted commercially (~60%) or
landfilled (~40%) (United States Environmental Protection Agency,
2010). Yard trimmings in the case study are 100% composted
commercially. Associated EU and GHG emissions were calculated
for operations at each modeled facility. For land application, landfilling and waste-to-energy disposal practices, biosolids were produced using anaerobic digestion and dewatered prior to
transportation. For incineration, wastewater solids from primary
and secondary treatment were dewatered prior to incineration.
Avoided processes as a result of this new approach were also inside
the boundary, including avoided fertilizer, carbon dioxide (CO2)
emissions sequestered as carbon in biosolids-amended soils, and
carbon sequestered in biochar.
2.2. Study scenarios and case study
Three national-level scenarios for US biosolids handling operations and four case study scenarios for the Tri-City WWTP in
Clackamas County, Oregon are evaluated. The national-level scenarios are:
N1) Current US biosolids disposal practice (55% land applied, 30%
landfill, and 15% incinerated) (North East Biosolids and
Residuals Association (NEBRA), 2007);
N2) Biosolids land application is discontinued and all biosolids
are landfilled or incinerated at rates consistent with current
disposal practices (67% and 33%, respectively);
N3) Biochar (yard trimmings feedstock) and biosolids are coproduced, mixed at an approximate ratio of 1:3.5 biochar
to biosolids (by weight) and land applied (only 55% of the
total biosolids are used), and the remaining biosolids landfilled (30%) and incinerated (15%).
For the national-level study, the functional unit was one dry
metric ton of biosolids generated at a WWTP, and the reference
flow was 8.25 dry metric tons biosolids per day. A city of 300,000
people is expected to generate this reference flow (0.01 dry metric
tons biosolids per person per year for 2009) (Metcalf and Eddy,
2003; Peters and Rowley, 2009), which is used to size the theoretical biochar and wastewater treatment plants. The amount of
yard trimmings available for the theoretical city was conservatively estimated from global availability in urban areas at 0.005
dry metric tons yard trimmings per person per year (2009e2010
basis) (United States Environmental Protection Agency, 2010;
Woolf et al., 2010). Biochar production from this feedstock was
estimated at 1.3 dry metric tons of biochar per day. Unless specified otherwise, all units herein referred to as tons are dry metric
tons.
To account for the specific disposal practices and local conditions for the case study WWTP, specific local scenarios (similar to
3
the national scenarios) were included for the Tri-City WWTP case
study. For example, the biosolids land application rate for this
WWTP (84%) is higher than the national average (55%). This is
expected as the rate for beneficial reuse of biosolids is high in the
state of Oregon, estimated at approximately 98% of those produced
(Oregon Association of Clean Water Agencies (2012)). Also, the
availability of yard waste was higher than global averages, which
allowed for a higher ratio of biochar to biosolids produced. The four
case study scenarios include:
C1) Current Tri-City WWTP biosolids disposal practice (83.7%
land applied, 14.6% landfill, and 1.6% combusted in a wasteto-energy facility);
C2) Biosolids land application is discontinued with all biosolids
landfilled or combusted at ratios consistent with current
disposal practices (90% and 10%, respectively);
C3a) Biochar (yard trimmings feedstock) and biosolids are coproduced at rates consistent with material availability and
mixed at an approximate ratio of 1:1 biochar to biosolids (by
weight) and land applied (only 83.7% of the total biosolids
are used). The remaining biosolids are landfilled (14.6%) or
combusted (1.6%);
C3b) Same as C3a, but biochar production is based on the amount
of energy from excess heat generated during the pyrolysis
process required to produce Class A biosolids and therefore
mixed at an approximate ratio of 1:3.8 biochar to biosolids
(by weight) and land applied.
Due to the large amount of yard trimmings available in the case
study, slightly more biochar than biosolids can be produced for land
application, and therefore, a large amount of excess energy is
available for the model. Scenario C3b is included to limit biochar
production based on the amount of energy that is needed to produce higher quality, Class A biosolids.
For the Tri-City WWTP case study (2013 average flow of 9
million gallons per day (MGD)), the functional unit remained one
ton of biosolids, and the reference flow was 3 tons biosolids per day,
generated by the 64,000 people that this plant serves (0.02 tons
biosolids per person for the year of 2013). The amount of yard
trimmings recovered and commercially composted from these
64,000 residents was estimated using data collected by the Oregon
Department of Environmental Quality (personal communication
with Peter Spendelow) for the Portland metropolitan area and was
10.8 tons per day (0.06 tons of yard trimmings per person for the
year 2012; 55% moisture content assumed (Cooger et al., 2002)).
From this feedstock, 2.6 tons of biochar per day could be produced
at the pyrolysis plant.
2.3. Wastewater solids and biosolids systems
The system boundaries included the dewatering process of the
wastewater solids or biosolids with the addition of polymer and
high-speed centrifugation, followed by transportation to final
deposition, and end-of-life of the biosolids (Fig. 1). An emulsion
polymer is used (national case: 27 kg neat polymer per ton biosolids and 13 kg neat polymer per ton wastewater solids (United
States Environmental Protection Agency, 2000); case study:
44 kg neat polymer per ton biosolids; assumed 41% active).
Methods considered for managing wastewater biosolids at endof-life include beneficial reuse by land application of Class B
biosolids, disposal of dewatered biosolids in a landfill, and
disposal of incinerated solids ash. The case study included combustion of biosolids at a waste-to-energy facility but not
incineration.
Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050
4
L. Miller-Robbie et al. / Journal of Cleaner Production xxx (2014) 1e10
2.3.1. Land application
The modeled process for biosolids land application included
dewatering biosolids from an anaerobic digester, transporting to
the land application site, and spreading the biosolids across agricultural fields. Dewatering was modeled with the addition of
polymer to the biosolids prior to centrifugation to achieve 78%
moisture content (Metcalf and Eddy, 2003; National Biosolids
Partnership, 2005; Peters and Rowley, 2009).
For both the national and case study scenarios, dewatered biosolids (78% moisture content) were transported to the land application site (412 km round trip). Biosolids were transported twice
per week with a model GMC c8500 Dump Truck (cargo weight 12
tons per load assumed) (Horvath and Stokes, 2011) and were
transported three times per week by a 5-axle semitrailer combination truck (cargo weight limit 24 tons per load assumed) (Bureau
of Transportation Statistics, 2012; Federal Highway Administration,
2000; Parker Ag Services LLC).
Biosolids or biosolids/biochar mix were modeled as being
spread onto land using a tractor and a manure spreader (personal
communication with Mike Scharp of Parker Ag Services) fueled by
off-road diesel (Argonne National Laboratory, 2010). The total nitrogen content of the biosolids was used to calculate the amount of
urea N fertilizer avoided, equating to 41.8 kJ energy avoided
(Sawyer et al., 2010) and 4000 kg CO2e avoided per ton N in biosolids (Sylvis Environmental, 2011). An estimated first-year, plantavailable nitrogen factor for anaerobically digested biosolids of 35%
was included to calculate avoided fertilizer (Pacific Northwest
Extension, 2007). Emission factors for CO2, methane (CH4), and
nitrous oxide (N2O) associated with land application were taken
from the BEAM 1.1 model (Sylvis Environmental, 2009, 2011)
(Table 1).
2.3.2. Landfilling
To dispose of biosolids at a landfill, it was assumed that stabilization would be required to reduce vector attraction. Anaerobic
digestion produced biosolids destined for the landfill, and dewatering was modeled in the same manner as for land application
(United States Environmental Protection Agency, 1999). Nationally,
about 4% of biosolids are disposed of in dedicated surface disposal
units, or monofills (North East Biosolids and Residuals Association
(NEBRA), 2007). In this study, it was assumed that processes
associated with monofills are similar to landfills.
For both the national and case study scenarios, dewatered biosolids were transported to a landfill (226 km round-trip) using a
dump truck and a combination truck (described above), each for
half of the trips and stockpiled at the landfill. Standard landfill/
earth moving equipment (Caterpillar D8N Dozer) was used to
Table 1
Direct emissions due to biosolids disposal when land applied and landfilled. Emission factors were calculated using the Biosolids Emissions Assessment Model
(BEAM) 1.1 model (Sylvis Environmental, 2009, 2011).
National
Total nitrogen ¼ 3.3%a
Organic carbon ¼ 31%b
Case study
Total nitrogen ¼ 6.7%c
Organic carbon ¼ 31%b
a
b
c
d
Emissions from land
applying biosolids
(kg/dry metric ton)
Emissions from
landfilling biosolids
(kg/dry metric ton)
CO2 ¼ 250
CH4 ¼ 0.04d
N2O ¼ 0.7
CO2 ¼ 250
CH4 ¼ 0.04d
N2O ¼ 1.5
CO2 ¼ 224
CH4 ¼ 111
N2O ¼ 0.8
CO2 ¼ 224
CH4 ¼ 111
N2O ¼ 1.6
(Metcalf and Eddy, 2003).
(Brown et al., 2012).
Data from Tri-City WWTP.
Based on a one-day average storage time prior to land application.
spread and compact the biosolids (Horvath and Stokes, 2011).
Table 1 provides emission factors attributed to landfilling biosolids.
2.3.3. Incineration
The third solids-management method considered was incineration of wastewater solids and landfilling of the ash residue. In this
study, 1 ton of biosolids (digested solids) were produced for every
1.7 tons of wastewater solids (undigested solids) due to the loss of
volatile solids during anaerobic digestion, for a total of 13.75 tons
wastewater solids per day (Metcalf and Eddy, 2003). Kim et al.
states that the minimum gross caloric content of solids required to
sustain autothermal combustion is 3.35 MJ/kg (Kim et al., 2005).
Using higher heat values for primary and secondary solids (25.6 MJ/
kg and 18.6 MJ/kg, respectively) (Metcalf and Eddy, 2003), the
minimum solids content required to sustain autothermal combustion is 15% solids, which is achieved by centrifugation.
For the national scenarios only, incineration was modeled to
take place on site so transportation of wastewater solids would not
be required. The incinerator was specified to be a continuously
operated fluidized bed incinerator with electrostatic precipitation
and wet scrubbing for the air emissions control equipment. Direct
emissions were based on CH4 and N2O released during hightemperature incineration (48.5 g-CH4 and 4304 g-N2O per ton
undigested solids) (Sylvis Environmental, 2011), and the energy use
was based on the net consumption of electricity for operation and
effluent gas treatment (1753 MJ per ton undigested solids) (Suh and
Rousseaux, 2002).
The modeling assumed that ash and solid waste from the electrostatic precipitator would require transport to a landfill (226 km
round-trip). The ash content of wastewater solids was assumed to
be 25% (Metcalf and Eddy, 2003) and combustion was assumed to
be complete, generating 3.4 tons of ash per day with a density of
500 kg/m3 (Kosior-Kazberuk, 2011). It was assumed that the solid
waste generated was not hazardous based on the industrial pretreatment regulations that require wastewater utilities to develop
and implement a program to protect plant operations (including
solids management).
2.3.4. Combustion at waste-to-energy facility
For the case study only, a small amount of biosolids (2%) are sent
to a mass-burn waterwall resource recovery facility which uses dry
flue gas scrubbers and fabric filter baghouses to control emissions
of acid gases and particulates. The net energy production for these
biosolids is estimated at 15 MJ per ton biosolids based on a dewatered biosolids energy content of 82 MJ per ton biosolids and a
plant energy production of 19% net/total. Direct emissions are
based on CH4 and N2O released during high-temperature incineration (48.5 g-CH4 and 4304 g-N2O per ton digested solids) (Sylvis
Environmental, 2011). Transport modes to the facility are
modeled the same as for landfill (above) and the distance assumed
is 115 km per load. Ash transport attributed to biosolids is assumed
to be negligible and is not modeled in this study.
2.4. Biochar system
Biochar is produced by thermal decomposition of biomass under oxygen-limited conditions. While several technologies exist for
biochar production, the main body of literature focuses on pyrolysis
and gasification (Meyer et al., 2011). Pyrolysis occurs under oxygenfree conditions, while gasification occurs under low-oxygen conditions. Gasification processes are typically optimized for syngas
production, having lower biochar yields than pyrolysis processes.
Moreover, research suggests that slow pyrolysis (longer residence
times and lower temperatures than fast pyrolysis) has a greater
potential for carbon sequestration, retaining up to 50% of the
Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050
L. Miller-Robbie et al. / Journal of Cleaner Production xxx (2014) 1e10
feedstock carbon as stable carbon in the biochar (Gaunt and
Lehmann, 2008; Ibarrola et al., 2012). Therefore, a slow pyrolysis
process with high biochar yields was modeled for this study to
evaluate sequestration benefits.
For all appropriate scenarios, the biochar production system was
based on an improved-kiln slow pyrolysis reactor operated by
Biochar Solutions Inc. (Carbondale, CO) (Biochar Solutions, 2013).
While the materials requirements and production capacity specific
to this system were considered, the material and energy balance
parameters (i.e., biochar and co-product yield, biochar composition) were based on a broader compilation of data from the literature to make these results more broadly applicable. However, the
mass and energy balance parameters considered in this analysis
were generally within ~30% of those communicated by Biochar
Solutions (Biochar Solutions, 2013).
This system included four phases: pre-processing, start-up,
pyrolysis reaction, and post-processing. During pre-processing, the
feedstock is homogenized in an electric hopper and then dried.
Once the feedstock is fed to the reactor, a small amount of natural
gas is used for start-up, or ignition of the feedstock. During slow
pyrolysis, the feedstock is thermally degraded under oxygen-free
conditions at peak temperatures of 300e600 C and residence
, 2012). Longer residence times
times of 1 h to several days (Manya
are required for lower temperature processes, and vice versa. The
peak process temperature presents a trade-off in char properties e
lower temperatures mean higher biochar yields but higher temperatures mean more of the carbon in the biochar is recalcitrant
(i.e., fixed carbon that will not enter the atmosphere as CO2).
Therefore, a moderate peak temperature of 450e500 C and residence time of 4e6 h was chosen to balance these properties (personal communication with Jonah Levine of Biochar Solutions, Inc.).
Once the pyrolysis reaction is complete (during post-processing),
the char is air-dried, sieved, and packaged in plastic drums.
The three products of this process are biochar (primary product), syngas, and bio-oil (co-products). Biochar yields for slow pyrolysis are generally between 20% and 50% by dry weight for a wide
range of feedstocks, including nut shells, corncobs, husks, stalks,
, 2012); a biochar yield of 30%
and various types of wood (Manya
was assumed here to be representative of the chosen pyrolysis
conditions (Mayhead, 2010). Furthermore, it was assumed that the
process uses yard trimmings as the input to achieve high recalcitrant carbon contents (e.g., 80% of carbon in biochar is fixed)
(Roberts et al., 2010; personal communication with Jonah Levine of
Biochar Solutions, Inc.). Biochar produced from feedstocks with
high nutrient contents (e.g., food waste and manure) tend to have
less recalcitrant carbon and more amorphous labile carbon (no
defined structure and unstable) than those produced from plant
feedstocks (e.g., wood, yard trimmings). Therefore, only plantbased feedstocks were considered for this study to increase the
carbon sequestration potential of the biochar (Harvey et al., 2012;
Singh et al., 2012).
Syngas and bio-oil are generally produced in similar quantities
during slow pyrolysis, thus it was assumed that the yield for each
co-product is 35% by weight (Mayhead, 2010). The syngas is primarily composed of carbon monoxide, hydrogen gas, and CH4,
while bio-oil is composed of larger oxygen-containing organic
molecules. Bio-oil has a relatively high energy content; when
compared to the heating value of petroleum fuels, bio-oil from fast
pyrolysis has a relative heating value of approximately 40% (on a
mass basis), while syngas has a relative heating value of approximately 15% (Czernik and Bridgwater, 2004; Laird et al., 2009). Both
co-products may be used in conventional boilers and consequently
in on-site applications. Additionally, it was assumed that 90% of the
heat lost through the reactor walls can be recovered and used for
heating applications such as drying feedstock (Roberts et al., 2010).
5
In the model used in this study, ~35% of the heat from pyrolysis was
used for drying feedstock and the remaining was used elsewhere.
For this study, CO2 emissions resulting from the combustion or
decay of syngas, bio-oil, and the labile portion of biochar were not
counted. The major reasons are: 1) the carbon in the plant tissue
was captured from the atmosphere during the plant's lifetime
(biogenic), 2) the biomass source is a waste byproduct of other
human activities, and 3) no land use change occurs to produce this
biomass (Searchinger et al., 2009). Likewise, in scenarios where the
yard trimmings are not used to make biochar, they are composted
and landfilled, and resulting biogenic CO2 emissions are not
counted though CH4 and N2O emissions are included as appropriate. However, the recalcitrant fraction of carbon in biochar has
been shown to have a half-life of greater than 100 years, even
millennia (Woolf et al., 2010). Therefore, this carbon was assumed
to be sequestered in biochar and credits were allocated herein.
2.5. Plant energy integration
Co-location of a pyrolysis plant with a WWTP facilitates integration of the excess energy products from biochar production into
the WWTP. The syngas and bio-oil produced by pyrolysis can be
combusted to power a boiler. The heat energy from the reactor can
be recovered directly by pumping a heat transfer fluid through a
jacket around the reactor. Therefore, the co-products of biochar
could be used for electrical generation, to heat buildings, and to
heat biological processes in cold weather seasons. The excess heat
from the reactor may be best used within the WWTP to produce a
higher quality biosolids product for land application as it can be
used for drying processes without consuming additional fossil
fuels.
Most health concerns associated with biosolids application can
be reduced by meeting the strict requirements for producing Class
A biosolids. Class A biosolids are considered of exceptional quality
and can be used for broader agricultural applications when
compared to Class B biosolids; according to the US EPA's Part 503
Rule, Class A biosolids “can generally be applied as freely as any
other fertilizer or soil amendment to any type of land” (United
States Environmental Protection Agency, 1994). Generating Class
A biosolids increases the amount of biosolids that can be landapplied and decreases transport distances associated with biosolids application because more options are available. Unfortunately, Class A criteria do not account for potential health risks
associated with contaminants of emerging concern in biosolids, but
the co-application of biochar and Class A biosolids could alleviate
some of these concerns as discussed in Section 3.4.
To achieve Class A biosolids, US EPA regulations must be met for
vector attraction, pathogen content, and metals. To sufficiently
reduce vector attraction and pathogen content to required levels,
biosolids must be dried to 10% water content at 80 C or higher for
20 min or longer (United States Environmental Protection Agency,
1994). Wastewater influent levels for metals were assumed to be
low enough that the biosolids meet the metals content requirements following treatment at the WWTP, which is often the
case with domestic wastewater. The amount of energy needed to
produce Class A biosolids is equivalent to that required to dewater
biosolids to 10% content, as shown in Equation (1):
ED ¼ ðMWC MWD Þ*EWR
(1)
where ED is the energy required to dry biosolids (MJ/day), MWC is
the mass of water in centrifuged solids (tons water/day) [tons
biosolids/day * % water in centrifuged solids], MWD is the mass of
water in dried solids (tons water/day) [tons biosolids/day * % water
in target solids], and EWR is the dryer energy requirements (MJ/tons
Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050
6
L. Miller-Robbie et al. / Journal of Cleaner Production xxx (2014) 1e10
water removed). This energy requirement was compared to the
amount of available energy from biochar production in Section 3.3.
The production of Class A biosolids using a Komline-Sanderson
paddle dryer to achieve this drying requirement is examined in
the results for scenario C3b.
3. Results and discussion
The life cycle EU and GHG emissions results are presented followed by an analysis of energy integration between the WWTP and
pyrolysis plant and a discussion of biochar, biosolids, and contaminants of emerging concern for co-application.
3.1. Energy use and GHG emissions from co-production of biosolids
and biochar
For all scenarios, EU and GHG emissions by life cycle stage (i.e.,
inputs to biochar production as well as inputs to biosolids processing, transportation, and end-of-life) are summarized in Figs. 2
and 3. Results show that biosolids processing is the most energyintensive life cycle stage for all scenarios, largely due to electricity
use for dewatering biosolids in a centrifuge and offsite production
of the polymer used for biosolids dewatering (Fig. 2). The higher
energy use for transportation in scenarios N3 and C3a can be
attributed to the transport of materials for biochar production
(primarily the transport of natural gas needed for start-up of the
pyrolysis process). Energy use at end-of-life is dependent on the
type of disposal and how much it is used for each scenario. Higher
energy use at end-of-life is due to electricity use for incineration,
and smaller amounts are due to fuel for spreading and compacting
biosolids into a landfill and for spreading biosolids at land application sites.
In Fig. 3, the GHG emissions do not follow the pattern of energy
use (Fig. 2) because not all GHGs are energy related. For example, in
the national scenarios (Fig. 3A), the end-of-life stage contributes
the highest GHG emissions primarily due to direct emissions during
landfilling and incineration. For the case study scenarios (Fig. 3B),
impacts from large amounts of yard trimmings and biochar are
reflected in the high emissions from composting and landfilling
yard trimmings for scenarios C1 and C2, and a high amount of
negative emissions due to recalcitrant carbon in the biochar at endof-life.
Overall, for the national case (Fig 2A), the addition of biochar
production adds 5% to EU between scenarios N1 and N3 but results
in a 26% reduction in GHG emission (Fig. 3A). When compared to
discontinued land application (scenario N2), the addition of biochar
production is even more sustainable, with the N3 scenario yielding
a 58% reduction in GHG emissions. Co-application of biochar with
biosolids may also allow for continued land application and
nutrient recycling.
For the case study (Fig. 2B), overall EU is increased slightly with
the addition of the pyrolysis plant (when compared to C1, a 9%
increase for C2 and a 3% increase for C3a). Large reductions in GHG
emissions for C3a and C3b result from fewer emissions from the
biochar pyrolysis process when compared to alternative uses for
yard trimmings with even greater reductions from the recalcitrant
fraction of carbon in the biochar itself. Overall, the results show that
small investments in materials and energy to add biochar production are effective in decreasing total system life cycle GHG
emissions at various scales. Importantly, the results displayed in
Figs. 2 and 3 do not account for energy generated from the pyrolysis
plant.
A further breakdown of the largest GHG contributors to each life
cycle stage for co-production (scenarios N3 and C3a) is illustrated in
Fig. 4. For the national scenario (Fig. 4A), the largest GHG emissions
contributor is direct biosolids disposal (landfill and incineration).
There are also high energy-related GHG emissions due to the
electricity use in dewatering biosolids and polymer production. For
the national case, biochar production itself adds very little to
overall GHG emissions (less than 1%).
For the case study (Fig. 4B), polymer production is a large
contributor to GHG emissions due to a higher amount of polymer
added to dewater biosolids when compared to the national case. A
greater amount of biochar produced for this scenario also results in
higher GHG emissions in biochar production and transport stages.
A larger proportion of biosolids are land applied in scenario C3a and
therefore disposal impacts are lower when compared to scenario
N3, but it remains a significant contributor.
3.2. End-of-life life cycle stages for biosolids and biochar
A direct comparison of the end-of-life options is illustrated in
Fig. 5 to better understand the GHG emissions per ton of biosolids
(i.e., the percentages of biosolids destined for each option are not
taken into account here). Incineration produces the highest total
GHG emissions in CO2 equivalents (Nakakubo et al., 2012; Sylvis
Environmental, 2009), followed by waste-to-energy (based on the
same direct emissions from incineration) and landfilling for this
model (Peters and Rowley, 2009). Land application has less impact
due to lower direct emissions on site and offsets from avoided
Fig. 2. Energy use for all A) national and B) case study scenarios, displayed by life cycle stage. LA ¼ land application. For scenarios without biochar production (N1, N2, C1, and C2),
energy use attributed to inputs for biochar production are the impacts from composting and landfilling of yard trimmings that would occur in lieu of biochar production.
Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050
L. Miller-Robbie et al. / Journal of Cleaner Production xxx (2014) 1e10
7
Fig. 3. Greenhouse gas emissions for all A) national and B) case study scenarios, displayed by life cycle stage. LA ¼ land application. For scenarios without biochar production (N1,
N2, C1, and C2), GHG emissions attributed to inputs for biochar production are the impacts from composting and landfilling of yard trimmings that would occur in lieu of biochar
production.
fertilizer use and carbon sequestration observed in biosolidsamended soils. Therefore, increasing the amount of land-applied
biosolids provides GHG emissions benefits. These results are
consistent with earlier work (Murray et al., 2008), which estimate
that biosolids disposal (landfilling and combustion) are more
Fig. 4. Greenhouse gas emissions for co-production of biosolids and biochar for both
the national scenario (N3, panel A) and case study scenario (C3a, panel B) by life cycle
stage. Units are kg CO2e per dry metric ton biosolids. Avoided/negative impacts and
carbon stabilized in biochar are not shown.
problematic than beneficial reuse (land application) in terms of EU
and GHG emissions.
Because biochar production is an energy positive process, the
resulting energy co-products (syngas, bio-oil, and thermal heat
from pyrolysis) were also quantified. This excess energy is not
accounted for in the overall EU comparisons above, but is shown for
co-production of biosolids and biochar in Fig. 6A and B. The EU
values of the co-products are on the right. The results show that
when co-products are included in the LCA, approximately 69% of
the total energy use is produced in the co-products (Fig. 6A and C).
Utilizing the energy-rich co-products could offset other energy
sources and associated GHG emissions within the WWTP (see
Section 3.3); however, such avoided GHG emissions are not
considered in this model.
Fig. 6C shows the extent to which recalcitrant carbon in biochar
contributes to reductions in system GHG emissions for the national
case. For the case study (Fig. 6D), the large amount of biochar
Fig. 5. GHG emissions for biosolids end-of-life options. Current percentages for
disposal practices are not applied and data are shown per ton of biosolids disposed for
each option. Equipment fuel combustion for spreading and compacting at the landfill is
included, but is 0.4% of the total for GHG emissions at the landfill. Likewise, electricity
generated at the waste-to-energy facility is 0.4% of the waste-to-energy total and
equipment fuel combustion for spreading biosolids on land is 3% of the landfill total.
For the avoided fertilizer shown in this figure, 3.3% total nitrogen biosolids content
(dry weight) is assumed (Metcalf and Eddy, 2003). Dotted lines represent the total for
each option after including any negative emissions.
Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050
8
L. Miller-Robbie et al. / Journal of Cleaner Production xxx (2014) 1e10
Fig. 6. Energy use for scenarios A) N3 and B) C3a and GHG emissions for scenarios C) N3 and D) C3a by life cycle stage including associated energy credits from biochar co-products.
All credits are shown in light gray.
produced for scenario C3a contributes to large reductions in GHG
emissions from recalcitrant carbon in biochar, but also at the land
application site due to soil carbon sequestration and avoided fertilizer needs. GHG emissions from the labile carbon in biochar and
any subsequent combustion-related emissions from biochar coproducts (syngas and bio-oil) are not included due to their
biogenic nature (see Section 2.4).
3.3. Energy integration at the WWTP
Co-production of biochar and biosolids is attractive from an
energy standpoint because the energy-rich co-products generated
from the pyrolysis process can offset energy use in the WWTP.
Based on the model for biochar production at the Tri-City WWTP (9
MGD), there is an estimated 14,801 kWh per day available from
excess heat, syngas, and bio-oil (scenario C3a). According to a
report by the Electric Power Research Institute, electricity use for a
10 MGD advanced WWTP with nitrification is 17,912 kWh per day
total of which 800 kWh per day is used for lighting and buildings
(Burton, 1996).
Regarding the production of Class A biosolids, Scenario C3b for
the case study models the system impacts with a sufficient (but not
excess) amount of feedstock needed to produce enough excess
useable heat from the pyrolysis reactor alone to generate Class A
biosolids: 2.8 tons yard trimmings per day resulting in 0.7 tons
biochar per day and 4670 MJ in excess heat per day after using a
portion to dry feedstock (1297 kWh per day). The excess heat can
be used to dry 2.5 tons of land-applied biosolids per day to 10%
water content to achieve Class A biosolids quality.
For the national case, a total of 8533 MJ per day (2370 kWh per
day) is needed to dry the 4.5 tons of land-applied biosolids per day.
The excess heat amounts to 9065 MJ/day (2518 kWh/day). Therefore, based on the modeled national average, more than 100% of the
energy needs for producing Class A biosolids are available and can
be used in a dryer to dry biosolids.
3.4. Biochar, biosolids, and contaminants of emerging concern
Despite the sustainability benefits of biosolids use for land
application, the practice involves the risk of releasing contaminants
of emerging concern into the environment. Current regulations for
biosolids focus on heavy metals only (United States Environmental
Protection Agency, 1994). Other contaminants of concern include
nutrients, biocides, flame retardants, steroids, hormones, and
pharmaceuticals (United States Environmental Protection Agency,
2009). Some studies have shown that the use of biochar can sorb
metals (Mohan et al., 2007; Uchimiya et al., 2010), decrease nitrate
leaching from biosolids-amended soils (Knowles et al., 2011),
decrease metal leaching and plant uptake from contaminated soils
(Park et al., 2011), and reduce N2O and ammonia (NH3) emissions
without the reduction of plant-available nitrogen (TaghizadehToosi et al., 2011). Another study found that biochar amendment
can either enhance or suppress GHG emissions relative to control
soils depending on the soil and biochar properties (Spokas and
Reicosky, 2009). The addition of biochar to biosolids may sorb
contaminants of emerging concern (Koelmans et al., 2006); however, more experimental studies are needed to investigate potential
reduction in contaminant transport. Literature gaps remain and
include a general lack of studies evaluating impacts associated with
the co-application of biochar and biosolids to agricultural lands
(e.g., long-term carbon sequestration, soil fertility, crop yields, etc.),
discrepancies in the application rates of investigated biosolids and
Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050
L. Miller-Robbie et al. / Journal of Cleaner Production xxx (2014) 1e10
biochar (Park et al., 2011; Spokas and Reicosky, 2009), variability in
the properties of investigated biochar and biosolids (Mohan et al.,
2007), and differences in soil types and climates between investigated sites. Due to the complexity of the biochar-biosolids-field
system, future studies must be rigorous in isolating variables during experiments.
4. Conclusions
Biochar provides many benefits to biosolids management in
terms of energy integration and reduction of GHG emissions at the
plant level. This study shows that the small amount of energy and
related GHG emissions invested in a pyrolysis plant can provide a
large amount of usable energy and life cycle GHG reductions e
scenario N3 GHG emissions are 26% less than scenario N1. Increased
land application and decreased disposal of biosolids in landfills and
incinerators can significantly reduce GHG emissions, as disposal
contributes high GHG impacts. The case study results show that
there may be an optimal point where both EU and GHG impacts are
reduced and Class A biosolids are produced.
For all co-production scenarios (N3, C3a, and C3b), 100% of the
energy needs to produce Class A biosolids are available directly
from the pyrolysis plant in the form of waste heat. Class A biosolids
are superior product that can be applied to various types of land.
With increasing concern for continued release of contaminants of
emerging concern to the environment through biosolids land
application, solutions that simultaneously address global concerns
as well as local human and ecological health are needed. This study
contributes a first analysis in evaluating the co-production of biochar and biosolids for potential environmental, economic, and
health benefits. The results can be used by WWTP utilities in
planning for more sustainable approaches to energy use within the
plant and land application of biosolids.
Acknowledgments
The authors would like to thank the National Science Foundation Engineering Research Center Program under Cooperative
Agreement EEC-1028968 (ReNUWIt) and the team at the Tri-City
WWTP in Clackamas County, Oregon including Michael Trent and
Mona LaPierre for partial support of this study. The authors also
thank Jonah Levine (Biochar Solutions), Peter Spendelow (Oregon
Department of Environmental Quality), Mike Scharp (Parker Ag
Services), Warren Fuller (Komline Sanderson), and Covanta Energy
Corporation.
Appendix A. Supplementary data
Supplementary data related to this article can be found at http://
dx.doi.org/10.1016/j.jclepro.2014.12.050.
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Please cite this article in press as: Miller-Robbie, L., et al., Life cycle energy and greenhouse gas assessment of the co-production of biosolids and
biochar for land application, Journal of Cleaner Production (2014), http://dx.doi.org/10.1016/j.jclepro.2014.12.050