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The price of protein: Review of land use and carbon footprints from life cycle assessments of animal food products and their substitutes

Authors:
The price of protein: Review of land use and carbon footprints from life cycle
assessments of animal food products and their substitutes
Durk Nijdam
, Trudy Rood, Henk Westhoek
PBL Netherlands Environmental Assessment Agency, Bilthoven/The Hague, The Netherlands
article info
Article history:
Received 13 January 2012
Received in revised form 13 July 2012
Accepted 7 August 2012
Keywords:
Protein
Meat
Dairy
Fish
Meat substitutes
Environmental impact
Land use
Carbon footprint
Life cycle assessment (LCA)
abstract
Animal husbandry, aquaculture and fishery have major impacts on the environment. In order to identify
the range of impacts and the most important factors thereof, as well as to identify what are the main
causes of the differences between products, we analysed 52 life cycle assessment studies (LCAs) of animal
and vegetal sources of protein. Our analysis was focused only on land requirement and carbon footprints.
In a general conclusion it can be said that the carbon footprint of the most climate-friendly protein
sources is up to 100 times smaller than those of the most climate-unfriendly. The differences between
footprints of the various products were found mainly to be due to differences in production systems.
The outcomes for pork and poultry show much more homogeneity than for beef and seafood. This is lar-
gely because both beef and seafood production show a wide variety of production systems.
Land use (occupation), comprising both arable land and grasslands, also varies strongly, ranging from
negligible for seafood to up to 2100 m
2
ykg
1
of protein from extensive cattle farming. From farm to fork
the feed production and animal husbandry are by far the most important contributors to the environ-
mental impacts.
Ó2012 Elsevier Ltd. All rights reserved.
Introduction
The role of animal husbandry in climate change and loss of bio-
diversity has been highlighted in several studies in the past decade
(e.g. Kramer, 2000; Steinfeld et al., 2006; Garnett, 2008; FAO,
2009). These publications provide the larger picture of the impacts
of livestock production on a global scale. More focus and detail can
be found in environmental life cycle assessment (LCA) studies of
animal food products, many of which were also published the past
few years.
In an LCA the environmental impacts of a product is quantified
as much as possible in a consistent and standardized way. De Vries
and De Boer (2010) have reviewed a selection of LCA studies on
animal products. Other meta-publications on LCAs of food products
include Yan et al. (2011),Roy et al. (2009), Flachowsky and
Hachenberg (2009) and González et al. (2011). These publications
mainly focus on greenhouse gases and carbon labelling, or include
a limited group of products or a limited number of studies. The
present review presents a broader view, based on the analysis of
52 LCA studies on meat, milk, seafood and other sources of protein.
The goals were to:
– Identify the ranges in land requirements and carbon footprints
of different sources of protein.
Identify the most important inputs and processes in the life
cycles.
– Identify what are the main causes of differences.
We focused on land use (occupation) and greenhouse gas emis-
sions because these aspects are very relevant to damage to ecosys-
tems and consequential global loss of biodiversity (Alkemade et al.,
2009; Rockstrom et al., 2009).
Methods
A method commonly used to analyse the environmental im-
pacts of products is the environmental life cycle assessment
(LCA). It is an internationally recognized method, and the ISO stan-
dards (ISO 14040 and 14044) provide guidelines for conducting
LCAs. They can be used to identify the most important contributors
in a production chain (gravity analysis or contribution analysis), or
to make a systematic comparison of different products or produc-
tion methods. Many different environmental impact categories can
be quantified in LCAs.
When they are only aimed at quantifying greenhouse gases the
method is often referred to as carbon footprinting. For this type of
LCA specific guidelines have been written (BSI, 2008).
0306-9192/$ - see front matter Ó2012 Elsevier Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.foodpol.2012.08.002
Corresponding author. Address: PBL Netherlands Environmental Assessment
Agency, PO Box 303, NL-3720 AH Bilthoven, The Netherlands. Tel.: +31 302742223.
E-mail address: durk.nijdam@pbl.nl (D. Nijdam).
Food Policy 37 (2012) 760–770
Contents lists available at SciVerse ScienceDirect
Food Policy
journal homepage: www.elsevier.com/locate/foodpol
In recent years, a large number of LCAs of agricultural products
have been published. There are specific methodological issues in
LCAs for agricultural and fishery products, such as: (Ayer et al.,
2007; Pelletier et al., 2007; Andersson, 2000; Guinée, 2002; Thrane,
2006).
– There are physical limits to production, mainly caused by the
pace of photosynthesis and availability of fertile land. Produc-
tion is not very flexible.
– Boundaries between natural and economic systems cannot
always be clearly defined. Are crop fields part of the natural
environment, or are they production sites? LCA methodology
requires clear system boundaries.
– Environmental effects are often local and difficult to quantify,
such as soil degradation, groundwater depletion and natural
habitat fragmentation.
– Some effects have to be discounted over a certain period of
time, such as those on biodiversity and changes in soil organic
carbon.
– In many cases agriculture or fishery produces more than one
product (co-production). The environmental pressure has to
be allocated between the different products.
Land is often included as a resource. Land use as a resource indi-
cator has to be interpreted with care, as there are many differ-
ent types and intensities of land use, all with different impacts
on the present and surrounding ecosystem. Moreover, land can
be made more fertile by increased inputs. Certain studies differ-
entiate between several types of land use. In the impact assess-
ment phase of LCA, where emissions and used resources are
aggregated and converted of into environmental impact catego-
ries, land use (and sometimes also land use change) is often
regarded as an indicator for loss of biodiversity, taking into
account quality (intensity) and quantity.
Life cycle assessments mostly are based on averages, represent-
ing a given production system. Such studies are also known as
‘attributional LCAs’ (ALCAs) (Ekvall and Weidema, 2004). Averages
may be taken from only a few farms or from national statistics.
They can represent actual farms or modeled farms. Another type
of LCA describes marginal rather than the average effects, which
are the result of new, additional production. The marginal pressure
often deviates from the average environmental pressure because
the most fertile soils already have been cultivated, and additional
production usually takes place on ‘new’ less productive land. Also
this additional production may displace other products from the
market. Such change-oriented LCAs are known as ‘consequential
LCAs’ (CLCAs) (Ekvall and Weidema, 2004). These two methods
may differ strongly in terms of application area, aim, scope, time-
scale, uncertainty and outcome (Brander et al., 2009; Thomassen
et al., 2008a).
Selection of LCA studies
For the purposes of this study 52 LCA studies were examined,
which are listed in Table 1. More details are presented in the an-
nex. We wanted to include as many studies as possible in order
to have a robust range and to be able to examine the differences.
For this reason some grey literature was also included. CLCAs
and other LCAs that use system expansion were excluded because
of the different scope, making them less comparable with regular
LCAs. Studies that did not describe the production system and
method in detail were excluded as well.
Most of the included LCAs have been published in peer re-
viewed journals. Some were published in reports, (e.g. Blonk
et al., 2008, 2009; Ponsioen et al., 2010; Williams et al., 2006;
Cederberg et al., 2009a; Hirschfeld et al., 2008). We only selected
reports which were executed or commissioned by non-commercial
scientific institutes (universities or governments).
All of the studies quantified the emission of greenhouse gasses,
17 of them also reported eutrophication and 18 also reported land
use. Several studies compared conventional production methods
with alternative methods, such as free range production systems.
Most studies covered one product or type of product, others cov-
ered many different types of products (e.g. Blonk et al., 2008; Wil-
liams et al., 2006).
Data used for the LCAs sometimes were based on a single farm
and in other cases on complete national industries. Often a typical
farm or production system was modeled based on national statis-
tics. Most of the LCAs were focused on European or North American
production processes, covering a variety of production systems.
The studies were published between 1998 and 2011. Most of the
data stem from the late 1990s up to around 2005. All studies de-
scribe the situation in a certain year, therefore no time series are
presented. Only Cederberg et al. (2009b) compares Swedish 1990
data with 2005 data, using several LCA studies.
Methodological issues
By-products and allocation
There are many co-production processes in agriculture and
much recycling takes place, as a result of which the environmental
pressures often have to be divided over several products or life cy-
cles, also known as allocation. Rapeseed, soybeans and sunflowers,
for example, supply both oil and fodder. At the end of the produc-
tion chain animals provide a wide range of products, such as
various cuts of meat, fat, hide and bones. A technique often used
Table 1
Overview of LCA studies reviewed.
Beef Pork
Blonk et al. (2008) Zhu and van Ierland (2004)
Casey and Holden (2006) Basset-Mens and van der Werf (2005)
Cederberg et al. (2009a,b) Williams et al. (2006)
Edward-Jones et al. (2009) Cederberg and Flysjö (2004a)
FAO (2010) Blonk et al. (2008)
Flachowsky and Hachenberg (2009) Eriksson et al. (2005)
Hirschfeld et al. (2008) Kool et al. (2009)
Nguyen et al. (2010) Hirschfeld et al. (2008)
Ogino et al. (2007) Poultry products
Pelletier et al. (2010) Blonk et al. (2008)
Peters et al., 2009 Katajajuuri (2007)
Phetteplace et al., 2001 Mollenhorst et al. (2006)
Ponsioen et al., 2010 Vergé et al. (2009)
Vergé et al., 2008 Williams et al. (2006)
Williams et al., 2006 Seafood (incl. freshwater fish)
Sheepmeat Aubin et al. (2009)
Edward-Jones et al. (2009) Blonk et al. (2009)
Peters et al. (2009) Ellingsen et al. (2009)
Williams et al. (2006) Gronroos et al. (2006)
Blonk et al. (2008) Iribarren et al. (2010a)
Milk and cheese Iribarren et al. (2010b)
Berlin (2002) Pelletier et al. (2009)
Blonk et al. (2008) Ramos et al. (2011)
Casey and Holden (2005) Silvenius and Grönroos (2003)
Cederberg and Flysjo (2004b) Svanes et al. (2011a,b)
FAO (2010) Vazquez-Rowe et al. (2010)
Haas et al. (2001) Vázquez-Rowe et al. (2011)
Hirschfeld et al. (2008) Vázquez-Rowe et al. (2012)
Sheane et al. (2011) Ziegler and Valentinsson (2008)
Thomassen et al. (2008b) Ziegler et al. (2003)
Vergé et al. (2007) Ziegler et al. (2011)
Weiske et al. (2006) Meat substitutes
Williams et al. (2006) Blonk et al. (2008)
Pulses
Blonk et al. (2008)
Nemecek et al. (2005)
Sheenan et al. (1998)
D. Nijdam et al. / Food Policy 37 (2012) 760–770 761
in co-production is economic allocation (Suh et al., 2010): dividing
the impacts over various products on the basis of their economic
value. However, as prices fluctuate in time due to changes in de-
mand, allocated environmental pressures may also change. A tech-
nique that avoids allocation is that of system expansion. Here it is
assumed that a co-product displaces another product on the mar-
ket (i.e. the avoided product) for which the environmental impacts
are known. These impacts are subsequently deducted from the im-
pacts of the original system. The often subjective nature of the
choice of avoided product and its production system can affect
the robustness of the results. The effect of using system expansion
instead of allocation in the life cycle of milk was analyzed in two
studies (Thomassen et al., 2008a; Flysjö et al., 2011a,b). In the cal-
culation with system expansion, Thomassen et al. chose regular
beef and pork as the avoided product from dairy farms, assuming
the beef of culled dairy cows is equivalent to it. As a result, the car-
bon footprint of the milk decreased considerably (Thomassen et al.,
2008a). Cederberg and Stadig and Flysjö et al. also found that the
outcome showed large differences, depending on the methodology
(Cederberg and Stadig, 2003; Flysjö et al., 2011a,b).
In the case of open loop recycling, such as the application of
manure, also an allocation of the burdens of ‘primary production’
should be made. However, this was not done or not reported in
the LCAs. When mentioned, all burdens of manure management
and application were fully attributed to meat or milk production.
Emissions of greenhouse gases
All the studies, except some LCAs of seafood, took into account
the N
2
O emissions from feed production, that is, from fertilization
of arable land, both direct (from fertilization and mostly also from
biological N-fixation) and indirect (from deposition, leaching and
run-off). Methane emissions from enteric fermentation were in-
cluded in all studies on pork, milk and ruminant meat. Emissions
from manure management (CH
4
and N
2
O) were included in all LCAs
on meat and milk. For these greenhouse gas emissions usually the
emissions factors from the IPCC guidelines (IPCC, 2006) were used.
Most important are the direct N
2
O emissions from fertilization (1–
1.25% of the input of nitrogen from fertilizer and 2% from manure
on grazed grasslands). However, in many cases, reference is made
to IPCC publications, without providing actual data used; for exam-
ple, only three studies provide data on the relatively important
emission of methane from rumen fermentation (Edwards-Jones
et al., 2009; Cederberg et al., 2009b; Ponsioen et al., 2010). These
emissions from cattle depend on diet characteristics and age, with
values ranging from 49 to 64 kg of methane per cow, per year.
Changes in soil organic carbon and the consequential emissions
of CO
2
due to cultivation were included only in studies by Blonk
et al. (2008) and Nguyen et al. (2010). Results were presented with
and without these emissions. We used the results without soil car-
bon emissions. Ten of the studies did not mention soil organic car-
bon, so it was unclear whether these emissions were included or
not. We assumed that changes in soil organic carbon were not in-
cluded. All other studies explicitly excluded soil carbon emissions
or assumed no change in soil organic carbon.
Blocking of the natural CO
2
sink (preventing the build-up of soil
organic carbon) through the arable use of land was also taken into
account by Blonk et al. (2008) and Nguyen et al. (2010). Results
were presented with and without these emissions. We used the re-
sults without soil carbon emissions.
One study mentioned methane emissions from soil (Williams
et al., 2006). As these appeared to have a very small effect on the
outcomes, we did not correct for these emissions. Production of
capital goods was excluded in about 40% of the LCA studies. Espe-
cially in LCAs of fish the production of vessels was often excluded,
because of the insignificant contribution, and the PAS 2050 guide-
lines for carbon footprinting (BSI, 2008), which explicitly require
the exclusion of capital goods production.
Emissions from the use of fossil energy were included in all
studies. For electricity production mostly national mixes were
used, which vary significantly between countries. For example
the Dutch mix, as used in studies by Zhu and Van Ierland (2004)
and Kool et al. (2009) resulted in 755 g CO
2
kW h
1
, whereas the
Swedish mix, as used by Cederberg and Flysjö (2004a) resulted in
only 45 g CO
2
kW h
1
, due to the high proportion of hydropower
and nucleair power. For background processes, such as transport,
tillage and other combustion processes, often general LCA dat-
abases were used, such as the Ecoinvent database (Ecoinvent-
centre, 2009).
In the impact assessment phase of LCAs greenhouse gases are
aggregated to CO
2
equivalents by GWP (Global Warming Potential)
factors. About half of the studies used the ‘old’ IPCC factors (21 for
CH
4
and 310 for N
2
O) (IPCC, 2006), and the other half used the
more recent equivalence factors (25 for CH
4
and 298 for N
2
O) (Sol-
omon et al., 2007). Since not all studies were explicit about the
GWP-factors used, no corrections were made. Recalculating the re-
sults from some LCAs of pork would have resulted in a 1–2% differ-
ence in the outcome. For beef, where methane is much more
important, the difference is larger. The carbon footprint of the
extensive Brazilian beef from (Cederberg, 2009a), for example, is
about 10% lower when recalculated with the ‘old’ factors.
In total the studies resulted in 104 carbon footprints and 43 val-
ues for land use. The land use is given in plain square metres, occu-
pied during a year. No weighing was performed. One study
differentiated between types of land used (Williams et al., 2006),
and one specified the continent on which the land use occurred
(Blonk et al., 2008). Another study used the ‘ecological footprint’
method, and presents aggregated global hectares (Pelletier et al.,
2010). These footprints were not included here.
Based on the selection of studies it is assumed that, methodo-
logical differences have limited impact on the results, and that it
is justified to compare results. On the whole similar choices and
assumptions were used. Standardized calculation procedures were
applied, and often performed with special LCA software. However,
the large amount of choices and assumptions did hamper compar-
isons. System descriptions and methodologies used were far from
uniform and often also far from complete. Several recommenda-
tions on this matter were made by Roy et al. (2009) after analyzing
13 milk LCAs. For our study we tried to use as much as possible
comparable results, i.e. with economic allocation and without out
soil carbon emissions. We did not correct for differences due to
old or new IPCC guidelines or differences in energy mixes.
Adjustment of functional unit and system boundaries
Although a full life cycle assessment should cover ‘cradle to
grave’, most of the studies cover only the chain from ‘cradle to farm
gate’. The most commonly used functional unit for meat is either
kilogram of carcass weight or live weight. To be able to compare
the results, the scores were converted to kilograms of boneless
Table 2
Yield factors in the meat chains. Source:Williams et al. (2006), Blonk et al. (2008, 2009) and Nguyen et al. (2010).
Beef (%) Pork (%) Mutton (%) Poultry (%) Fish (%)
Killing out factor (carcass weight of live weight) 53 75 46 70 40
Edible meat yield (retail meat of carcass) 70 75 75 80 100
762 D. Nijdam et al. / Food Policy 37 (2012) 760–770
retail meat. An emission of 0.2 kg CO
2
kg
1
meat was assumed for
the slaughterhouse and 0.1 kg kg
1
for the transport of meat (Blonk
et al., 2008). From live weight to carcass weight to retail weight,
average yield factors (killing out percentage and carcass saleable
meat percentage) were taken from Blonk et al.(2008), Nguyen
et al. (2010) and Williams et al. (2006)(Table 2). For milk, a farm
to retail emission of 0.12 kg CO
2
-eq kg
1
was assumed (Sevenster
and DeJong, 2008), excluding product loss. For seafood, four
values were found for the post landing phase, ranging from
0.1 kg CO
2
-eq kg
1
for canned mussels (Iribarren et al., 2010a)to
0.9 kg CO
2
-eq kg
1
for fresh lobster (Ziegler and Valentinsson,
2008). For frozen cod fillet Ziegler et al. (2003) found 0.6 kg
CO
2
-eq kg
1
, and Blonk et al. (2009) found 0.5 kg CO
2
-eq kg
1
.
We used a value of 0.5 kg CO
2
-eq kg
1
fillet as a default. We did
not correct for packaging, as this is of minor importance in the life
cycles (Blonk et al., 2008).
Animal food products, from a nutritional point of view, are
mainly consumed for their protein. As the products contain differ-
ent percentages of protein (e.g. milk has a much lower protein con-
tent than meat or fish), we recalculated the outcomes per unit of
protein. The protein content was taken from the Dutch food data-
base (NEVO, 2010). Apart from protein animal and vegetal prod-
ucts also supply other important nutrients. Oily fish are
important sources of vitamin D and long chain omega 3 fatty acids.
Pulses are a large source of calcium and carbohydrates. Dairy is an-
other large source of calcium. Red meat is a large source of iron. All
environmental impacts however, were allocated to the protein
content only. Moreover proteins are not fully comparable, vegetal
proteins have an amino acid composition that is less easy to digest
for humans than are animal proteins. As the modern western diet
contains much more protein than needed (Westhoek et al., 2011),
we did not take the difference in protein quality into account here.
Results
The next two sections present the land use and carbon foot-
prints of the various food products, according to the LCA studies.
Tables and graphs only present results for conventional produc-
tion, with the exception of the results for eggs from free range pro-
duction from Mollenhorst et al. (2006), as this represents a large
part of the consumer market in Europe.
Results per kilogram product
LCA outcomes provide a range in environmental impacts of the
different products (Table 3). On the whole, ruminant meat has the
largest impact, both in terms of greenhouse gas and land use. Poul-
try products and vegetal products have the smallest impacts per kg
product. The range of the carbon footprint is especially large for
beef products and seafood. Ruminant meat from extensive produc-
tion systems and seafood from energy-intensive fisheries show by
far the largest carbon footprints per kg edible product. On the low-
er end are mussels, milk, poultry products and vegetal products.
Land use ranges from zero for wild-caught fish to over 400 m
2
-
ykg
1
for beef from extensive production systems.
The results indicate that there are large differences between the
products. Apart from the high values for pork (Zhu and Van Ierland,
2004) and poultry (Williams et al., 2006), the outcomes for pork
and poultry show much more homogeneity than for beef. This is
largely because of the very wide variety in beef production sys-
tems, ranging from very intensive to very extensive. A wide range
can also be seen for seafood. On the carbon efficient side there are
some pelagic fisheries and aquaculture systems, and on by far the
least efficient side there is lobster trawling, which requires eight li-
tres of diesel for each kilogram of lobster catch, delivering only
300 g of edible meat (Ziegler and Valentinsson, 2008). The largest
differences are found in beef farming. The 15 LCA studies on beef
cover a variety of cattle farming systems; from intensive fattening
calf production (both dairy and beef calves) to highly extensive
pastoral systems. Production of 1 kg of extensively farmed beef re-
sults in roughly three to four times as many greenhouse gas emis-
sions as the equivalent amount of intensively farmed beef.
Differences are mainly caused by differences in farming system.
In intensive systems the nutrients in the feed are relatively effi-
ciently ‘transformed’ into meat and dairy because the animals do
not have to (or cannot) walk much about to find their food.
The process of fermentation in the rumen of ruminants pro-
duces the greenhouse gas methane. This is the main reason why
beef and lamb score relatively high in terms of greenhouse gas
emissions. Land use related to ruminant meat is relatively great,
particularly for meat from extensive grassland farming, since these
grasslands are less productive than arable lands, and because cattle
have a slow reproductive cycle and a relatively low feed
conversion.
Table 3
Carbon footprint and land use of protein rich products per kilogram of product, from several LCA studies (cradle to retail, n= number of analyzed products, NB for land use the
number may be lower).
Product Carbon footprint (kg CO
2
-eq kg
1
) Land use (m
2
ykg
1
) Of which grassland (m
2
ykg
1
)
Beef (15 studies, n= 26) 9–129 7–420 2–420
Industrial systems (n= 11) 9–42 15–29 2–26
Meadows, suckler herds (n= 8) 23–52 33–158 25–140
Extensive pastoral systems (n= 4) 12–129 286–420 250–420
Culled dairy cows (n= 3) 9–12 7 ca 5
Pork (eight studies, n= 11) 4–11 8–15
Poultry (four studies, n= 5) 2–6 5–8
Eggs (four studies, n= 5) 2–6 4–7
Mutton and Lamb (four studies, n= 5) 10–150 20–33 ca 18–30
Milk (12 studies, n= 14) 1–2 1–2 ca 1
Cheese
*
6–22 6–17 ca 7
Seafood from fisheries (nine studies, n= 18) 1–86
**
Seafood from aquaculture
***
(seven studies, n= 11) 3–15 2–6
***
Meat substitutes containing egg or milk protein (one study, n= 2) 3–6 1–3 0–2
Meat substitutes, 100% vegetal (one study, n= 4) 1–2 2–3
Pulses, dry (two studies, n= 3) 1–2 3–8
*
Range based on milk range and results from the study by Berlin (2002). For cheese, 6–7 kg of milk is required (Blonk et al., 2008).
**
Land use: bottom trawling may have an effect on large areas of the seabed (Davies et al., 2009; Ellingsen and Aanondsen, 2006; Vázquez-Rowe et al., 2011; Ziegler and
Valentinsson, 2008).
***
Land use: only land used for vegetal feed component.
D. Nijdam et al. / Food Policy 37 (2012) 760–770 763
The environmental impact of the beef from culled dairy cows is
low compared to that from beef cattle. This is mainly due to the
relative efficient co-production of meat and milk in intensive sys-
tems. Because dairy cows need to be milked regularly, distances
to the milking parlour are usually short. This means intensive graz-
ing takes place nearby the farm, or grazers are kept indoors perma-
nently. Therefore, livestock management systems of dairy farms
generally do not vary greatly, with values between 1 and 1.5 kg
CO
2
-eq kg
1
milk (12 studies). Weiske et al. (2006) give an average
of 1.4 kg CO
2
-eq kg
1
for milk for the EU-15. In a study by the FAO
(2010), an average of 1.3 kg CO
2
-eq kg
1
is calculated for western
Europe. The differences can be traced back to soil condition and
consequent N
2
O emissions (De Vries and De Boer, 2010), feed com-
position and race (related to yield) (Vergé et al., 2007), intensity of
farming (mainly related to yield and diet) and manure manage-
ment (Haas et al., 2001; Phetteplace et al., 2001; Weiske et al.,
2006). The emission of methane from cows is the main contributor
in the dairy carbon footprint. The processing of the milk to dairy
products is of less importance (Berlin, 2002; Sheane et al., 2011;
Blonk et al., 2008).
Pork shows a medium carbon footprint. Most of the eight stud-
ies reported values of around 5 kg CO
2
-eq kg
1
meat, which are
reportedly mostly due to the N
2
O emissions from feed production.
Considerably higher values were presented only by Williams et al.
(2006), with 8.7 kg CO
2
-eq kg
1
, and Zhu and Van Ierland (2004),
with 10.6 kg CO
2
-eq kg
1
.Zhu and Van Ierland (2004) state that
the high value is caused by a very high energy related CO
2
emis-
sion, which is about eight times higher than the value reported
by Kool et al.(2009), who also describe Dutch production. When
this high energy related CO
2
emission (ca 4 kg CO
2
kg
1
) is re-
placed by the value from Kool et al. (2009) (ca 0.5 kg CO
2
kg
1
),
then the outcome from Zhu and van Ierland (2004) is much more
in line with the others studies. Williams et al. (2006) also reported
a relatively high CO
2
emission (2.5 CO
2
-eq kg
1
), but it is unclear
whether this is purely energy related.
Basset-Mens and Van der Werf (2005) compare pork in conven-
tional systems to free range pork (Label rouge) which has about a
50% higher carbon footprint. According to Kool et al. (2009), the
carbon footprint of organic pork is 20% higher than that of conven-
tional pork. According to Williams et al. (2006) the carbon foot-
print of organic pork is 12% lower.
Poultry meat (four studies) has a small environmental impact
compared with other types of meat, in terms of both greenhouse
gas emissions and land use. Three of the four studies found
approximately 3 kg CO
2
-eq kg
1
chicken meat. Only Williams
et al. (2006), presented a much higher outcome (6 kg CO
2
-eq).
The carbon footprints of eggs (four studies) is of the same order
of magnitude. The smallest carbon footprint is found by (Vergé
et al., 2009) (1.7 kg CO
2
-eq kg
1
), and the largest by Williams
et al. (2006) (5.5 kg CO
2
-eq kg
1
). Free range poultry has a mod-
estly larger carbon footprint (Kool et al., 2009). According to Mol-
lenhorst et al. (2006) free range eggs have a 10% larger carbon
footprint. This probably also applies to free range poultry meat.
In the category of seafood (marine and freshwater products, 16
studies, n= 29) there are considerable differences, ranging from
about 1 kg CO
2
-eq per edible kilogram for Spanish mussels
Fig. 1. Carbon footprints per kilogram of protein.
Fig. 2. Land use per kilogram of protein.
764 D. Nijdam et al. / Food Policy 37 (2012) 760–770
(Iribarren et al., 2010a), North–East Atlantic Mackerel (Ramos
et al., 2011), Baltic herring (Silvenius and Grönroos, 2003)to
86 kg CO
2
-eq kg
1
for trawled Norwegian lobster (Ziegler and Val-
entinsson, 2008). Other studies also showed the relative ineffi-
ciency of bottom trawling (Vázquez-Rowe et al., 2011; Svanes
et al., 2011a,b; Ramos et al., 2011).
A large variance was found by Ramos et al. (2011) for North–
East-Atlantic Mackerel, both in time as depending on fishing fleet,
ranging from less than 1 kg CO
2
-eq kg
1
fillet from the Basque
purse seine fleet to over 6 kg CO
2
-eq kg
1
fillet from the Galician
bottom trawl fleet. Alaskan pollack also has a relatively small car-
bon footprint (about 3 kg CO
2
-eq kg
1
fillet) (Blonk et al., 2009).
Cod shows ranges from 3 to 7 kg CO
2
-eq kg
1
fillet (Svanes et al.,
2011a,b; Blonk et al., 2009; Ziegler et al., 2003).
Fishing methods using purse seines or gillnets, and midwater
trawling generally take much less energy than bottom trawling
and longline fishing (Tyedmers, 2004). Target species of bottom
trawling are redfish, flatfish, cod, hake and crustaceans, but these
types of fish can however also originate from demersal trawling
(just above the seabed), traps (crustaceans), longlines (cod) or
aquaculture (flatfish). In addition to fishing method, also the den-
sity of stocks and the distance to the fishing grounds is of impor-
tance to total energy consumption (Harman et al., 2008).
Aquacultured seafood also shows a large variance. Mussels are
grown from wild caught seed on artificial substrate or on reserved
sections of the seabed. They do not require feeding and can be har-
vested with little energy requirements. Land based aquaculture of
carnivorous species, like turbot, trout or shrimp are relatively en-
ergy-intensive and require high protein feed. Cages at sea (salmon,
sea-trout and sea bass) are less energy-intensive.
Farmed salmon (four studies) has a carbon footprint ranging
from 3 to 8 kg CO
2
-eq kg
1
fillet, and involves a certain amount
of land use for the vegetal component of the feed. Farmed panga-
sius, an omnivorous fish with a predominantly vegetarian diet,
shows a modest carbon footprint (3 kg CO
2
-eq kg
1
fillet) and has
a land use of 5 m
2
ykg
1
fillet (Blonk et al., 2009). Land-farmed
trout (two studies) has a carbon footprint of about 7 kg
CO
2
-eq kg
1
fillet.
Vegetarian products (meat substitutes, 1 study, n= 3) have car-
bon footprints ranging from less than 1 kg CO
2
-eq kg
1
for a vege-
tal meat substitute to 6 CO
2
-eq kg
1
for a meat substitute enriched
with milk protein (Blonk et al., 2008).
Pulses (three studies, n= 3) have carbon footprints ranging
from less than 1 CO
2
-eq kg
1
for peas and soya (Sheenan et al.,
1998; Nemecek et al., 2005) to 2 kg CO
2
-eq kg
1
for European
common beans (Blonk et al., 2008).
Results per kilogram protein
When comparing the environmental pressure per kilogram pro-
tein, the differences between products are smaller (Table 4 and
Figs. 1 and 2). The carbon footprint per kilogram protein ranges
from about 4 kg CO
2
-eq for vegetal meat substitutes, pulses, mus-
sels and herring to over 600 for highland ruminants. So the ‘best
case’ sources of protein have carbon footprints that are about
150 times smaller than the ‘worst case’ sources. The range in land
use is even larger, from less than 10 m
2
ykg
1
protein for vegetal
products and meat substitute with egg protein to over 2000 m
2
-
ykg
1
for meat from extensively farmed ruminants. The range
for sheep meat seems much smaller than for beef. This could how-
ever be due to the fact that only two (European) studies quantified
the land use for sheep meat. In the case of extensive production,
such as in Australia, land use may be much greater.
Contribution analysis and improvements within production chains
In addition to being able to compare products, LCAs also may be
used to identify the most important processes, from an environ-
mental point of view, in complex production chains.
Table 4
Carbon footprint and land use related to protein rich products per kilogram of protein, according to several LCA studies (cradle to
retail, protein content of products is given between brackets).
Product (%protein) GHG kg CO
2
-eq kg
1
protein Land use m
2
ykg
1
protein
Beef (20%) 45–640 37–2100
Industrial systems 45–210 75–143
Meadow systems, suckler herds 114–250 164–788
Extensive pastoral systems 58–643 1430–2100
Culled dairy cows 45–62 37
Pork (20%) 20–55 40–75
Poultry (20%) 10–30 23–40
Eggs (13%) 15–42 29–52
Mutton and lamb (20%) 51–750 100–165
Milk (3.5%) 28–43 26–54
Cheese (25%) 28–68 26–54
Seafood from fisheries (16–20%) 4–540
Seafood from aquaculture (17–20%) 4–75 13–30
Meat substitutes containing egg- or milk protein (15–20%) 17–34 8–17
Meat substitutes, 100% vegetal (8–20%) 6–17 4–25
Pulses, dry (20–36%) 4–10 10–43
Table 5
Distribution of the carbon footprint of pork (kg CO
2
-eq kg
1
retail meat and percentages).
Kool et al. (2009) Basset-mens and Van der Werf (2005) Eriksson et al. (2005)
kg CO
2
-eq % kg CO
2
-eq % kg CO
2
-eq %
Feed production 3.0 62 3.2 73 3.0 66
Husbandry 1.9 38 1.2 27 1.6 34
Enteric fermentation 4
Manure management 24
Energy use 10
Total 4.9 100 4.4 100 4.6 100
D. Nijdam et al. / Food Policy 37 (2012) 760–770 765
From the so-called contribution analysis in the LCAs we found
that, for meat products, feed production and animal husbandry
are the most important with respect to land use and greenhouse
gas emissions. For pork, the field emissions of nitrous oxide from
feed production, and for beef, the methane emission from enteric
fermentation are by far the most important emissions (Tables 5
and 6).
Since feed production is a very important factor in the environ-
mental impact of pork production, increasing feed conversion effi-
ciency is an important way to reduce these impacts. According to
Haxsen (2008) the standardized conversion ratio in seven Euro-
pean Countries ranges from 2.7 to 3 kg feed kg
1
weight gain. In
the Netherlands and Denmark, feed conversion is the most effi-
cient, while in Belgium and the United Kingdom it is the least effi-
cient. Feed conversion in Germany, France and Ireland is of
medium efficiency.
Beef
In Europe about half of the beef comes from culled dairy cows,
and the other half is produced by beef systems (Weidema et al.,
2008). The relatively intensive European dairy systems are charac-
terized by little or no meadow grazing, large portions of concen-
trates in the feed and housing in a confined space. Beef
production from fattening calves is characterized by intensive or
semi-extensive (meadow) production systems, with grazing in
summer, and indoor feeding in winter. In both systems, methane
from enteric fermentation and emissions from manure are by far
the most important contributors to the carbon footprint. Table 6
provides an overview of the shares in different systems, according
to various studies.
Some of the other studies also contained a contribution analy-
sis, but classifications of life cycle phases differed strongly, so they
could not be included in Table 6. According to Casey and Holden
(2006) 60% of the carbon footprint of Irish beef (35 kg CO
2
-eq kg
1
retail meat) can be attributed to enteric fermentation, 18% to fertil-
izer production and use, around 10% to manure management,
around 8% to concentrate production, and 4% to the use of diesel
and electricity. For organic beef, the share of enteric fermentation
is somewhat higher, as the share of fertilizer production and use is
zero, and that of concentrate production is much lower than for
other beef. Ogino et al. (2007) found similar shares in the case of
Japanese beef (36 kg CO
2
-eq kg
1
retail meat): 61% for enteric fer-
mentation, 27% for feed production and 12% for manure
management.
For poultry, nitrous oxide emissions from the fertilization of
arable land and manure management are by far the most impor-
tant contributors to its carbon footprint (Blonk et al., 2008; Mollen-
horst et al., 2006).
For dairy products methane from enteric fermentation, manure
management and field emissions from feed production are the
three major contributors to the carbon footprint (Blonk et al.,
2008; Sheane et al., 2011).
An optimum dosage of fertilizers, balanced animal feed and
accommodation, and good manure management may reduce these
emissions. According to Weidema et al. (2008), a reduction of 25%
in greenhouse gas emissions from animal husbandry in the EU
could be achieved by implementing several existing improvement
options. According to Weiske et al. (2006), for the dairy sector this
percentage could possibly be even higher: for example through
optimized lifetime efficiency and covered storage and anaerobic
digestion of manure.
In aquaculture, feed production also plays a major role. In land-
based aquacultural systems the energy use of water pumps is very
relevant. In fisheries the fuel consumption of the vessels is by far
the most important factor. It can vary between 0.1 to over 3 l/kg
landed fish in industrial fisheries (Tyedmers, 2004).
In the life cycle of vegetal meat substitutes, crop production and
energy use in food processing are the most important contributors
(Blonk et al., 2008). Development of energy-saving techniques may
lead to improvements here.
Food miles, storage and packaging
Throughout the analysed product life cycles transportation
takes place, both of raw materials and the processed (intermediate)
products. These ‘food miles’ can be very high and may contribute to
a product’s carbon footprint. Weber and Matthews calculated an
average transportation distance of over 8000 km in the life cycle
chain of food products (Weber and Matthews, 2008). Despite this
impressive distance, transportation on average accounts for only
11% of the carbon footprint of food. Apart from the distance, the
means of transportation is also very important. For large ships,
the energy use per tonne-kilometre is very low. According to the
Ecoinvent database (Spielman et al., 2007) bulk carriers (55–
250 kt) emit about 25–250 times less greenhouse gasses per
tonne-kilometre than trucks. Therefore, road transport across Eur-
ope can have a larger effect on the climate than transatlantic ship-
ping. Air freight, with a greenhouse gas emission level of 1–
2kgCO
2
-eq ton-km
1
, emits about five times more greenhouse
gases than trucks (Spielman et al., 2007). The additional impact
from product refrigeration during transport and storage may be
large, but the same applies here: the larger the volumes, the lower
the additional impact (Carlsson-Kanyama and Faist, 2000).
Several LCA studies took transport into account. In (Blonk et al.,
2008), the share of transportation of animal feed in the total carbon
footprint was about 10% for pork and poultry and 12% for salmon.
Post-farm transportation accounted for less than 10%, except for
pork and poultry, where it was 12%. In the life cycle of salmon, pork
and poultry total transport makes up about one fifth of the carbon
footprint, and therefore is not completely negligible. For Scottish
fresh milk, a share of 2% was calculated for product transport
(Sheane et al., 2011). For seafood, several products were analyzed
by Harman et al. (2008), who found a share of 15% to 55% for
transports and refrigeration in the total carbon footprint of
Table 6
Distribution of the carbon footprint for beef (kg CO
2
-eq per kg retail meat and percentages) of various production systems.
Suckler calves extensive
production Brazil
a
Suckler calves semi-extensive
production Ireland
a
Mixed calves intensive
production NL
a
Culled-dairy cows intensive
production NL
b
kg CO
2
%kgCO
2
%kgCO
2
%kgCO
2
%
Enteric fermentation 44.3 75 19.0 50 6.1 38 4.0 40
CH
4
from manure 1.2 2 3.8 10 2.2 14 1.1 11
N
2
O manure and fertilizer 11.8 20 10.6 28 4.3 27 1.4 14
Feed production 2.3 6 1.1 7 2.6 26
Others 1.8 3 2.3 6 2.2 14 0.9 9
Total 59 100 38 100 16 100 10 100
a
Blonk et al. (2008).
b
Ponsioen et al. (2010).
766 D. Nijdam et al. / Food Policy 37 (2012) 760–770
economically important fish products on the UK market. For fresh
products, the share is generally much higher than for frozen prod-
ucts. For flown in fresh tuna from the Maldives, the share of (air)
transport was as high as 95%. Vázquez-Rowe et al. (2012) found
that shipping frozen atlantic octopus to Tokio raised the carbon
foootprint by 2%.
Some studies also included packaging. Milk cartons contribute
about 5% of the carbon footprint of milk (Blonk et al. 2008; Sheane
et al., 2011), plastic packaging contributes about 4% of the carbon
footprint of pork (Blonk et al., 2008).
Discussion
Trade-offs and rebounds
This review covers the carbon footprints and land use of the ma-
jor protein-rich products in the western diet. Cereals are also a sig-
nificant supplier of protein, but are not included. On the basis of
differences in environmental impact of the various products, we
conclude that there is a large potential for reductions in the envi-
ronmental impact of food consumption by choosing low-impact
sources of protein. However, large scale shifts may have rebounds
or trade-offs. The impact of such shifts or improvements could be
analyzed using global agro-economic equilibrium models. In the
case of large-scale changes the dynamic approach will give more
accurate scenarios than the straightforward extrapolation of LCA
results as these models are dynamic and take into account changes
in price due to changing demand and supply, limited production
factors and rebound effects (Stehfest et al., 2012). An example of
this is the assessed option to increase livestock productivity (i.e.
the quantity of feed needed per unit of product) by 15%. The eco-
nomic models predict as a results of price feedbacks, that con-
sumption of livestock products would increase by about 3%, thus
reducing the potential environmental gains.
Other impacts, intensification or extensification?
In addition to greenhouse gasses and land use, there are other
issues that may also be very relevant, such as animal welfare,
eutrophication, emissions of pesticides, use of hormones and
depletion of resources.
Animal welfare is an important issue since the emergence of
industrial husbandry. One of the major aspects is the limited living
space for animals. Improvement of this aspect means the animal
can move about more freely, and other breeds would have to be
used. Results from such improvements would mean lower feed
conversions, however, and more time and feed would be needed
to bring animals up to slaughter weight. This is a dilemma, but
the added environmental pressure would only be small compared
to the large ranges between, but also within product categories.
Optimized free range production could have a lower impact than
regular production systems. The question is to find an optimum
balance. Implementing better welfare options would result in a
modest increase in the environmental impacts of poultry, pork,
and fish from aquaculture. However, the overall conclusions of this
review would still be valid. For cattle rearing in extensive systems,
a further intensification could affect semi-natural ecosystems. Gi-
ven the rising global demand for meat, the increase in yield how-
ever could alleviate the pressure on the remaining natural areas,
especially in South America. Beef production in Brazil is currently
relatively extensive, but is becoming much more intensive. Be-
tween 1970 and 2007, beef yield per hectare quadrupled and the
carbon footprint per kilogram of beef has more than halved (Pons-
ioen et al., 2010). Because grazing animals have been around for a
very long time in such original grasslands, nature has adapted to
grazing, leading to a richness in species. Extensive grazing systems
currently often preserve this biodiversity, albeit mostly by newly
introduced species of grazers. The balance tips when the grazing
gets too intensive and the grasslands may deteriorate (Blanco-Can-
qui and Lal, 2008). Carbon footprints of beef from extensive pro-
duction may seem very high compared to beef from intensive
systems, however this is at least partly dependant on assumptions
regarding allocation and soil carbon sequestration (Flysjö et al.,
2012).
Climate change and land use affect biodiversity on a global
scale. Other relevant issues, such as emissions of nutrients and pes-
ticides may have more local impacts. Eutrophication of water and
soils, for example, has become a serious problem throughout large
parts of Europe (Sutton et al., 2011). Animal husbandry plays a key
role here. Normalization procedures, that is relating LCA ‘scores’ to
national or European emissions, according to the reviewed LCAs,
often indicate that eutrophication of water and soils is a main issue
in animal husbandry and aquaculture. In a review of LCA studies on
livestock products, De Vries and de Boer (2010) found a range of
about 2–20 g of phosphate-equivalents per kilogram of product
(a unit used in LCAs, in which nitrogen and phosphorus com-
pounds are aggregated). Animals with high feed conversion effi-
ciencies, such as poultry, score the best, while pork and beef
score highest. This is in keeping with carbon footprint outcomes
of this review. There does not seem to be a trade-off here. A high
conversion ratio could be enhanced by using hormones and antibi-
otics, although this may result in public health risks (Marshall and
Levy, 2011).
An argument against reduction of the consumption of ruminant
meat often put forward is the unique ability of grazers to convert
grass to high quality human food. In relative intensive production
systems however, beef and dairy cattle is at least partially fed with
feed coming from arable land (Lesschen et al., 2011; Westhoek
et al., 2011). There are significant areas of extensive semi-natural
grasslands in Europe that are only suitable for grazing, but the pro-
duction volumes from these areas are small compared to total pro-
duction of ruminant meat and dairy.
Fishing can have large impacts on marine ecosystems. The re-
moval of large amounts of fish directly affects predators, predated
species and competitors. Bottom trawling generally has very high
discard rates and destructive effects on the seabed (Davies et al.,
2009). In addition to small carbon footprints, European mussel
cultures score well on several other sustainability indicators pub-
lished by the Fisheries Centre of the University of British Columbia
(Trujilo, 2008). Other shellfish, such as oysters and cockles also
score well. Tropical shrimp and prawn cultures and fisheries score
the worst according to this indexation.
Protein quality
When comparing different sources of protein we did not correct
for protein quality. This can be expressed as the protein digestibil-
ity corrected amino acid score (PDCAAS). Generally speaking a lac-
to-ovo vegetarian diet requires a 20% larger intake of protein, and
for a fully vegan diet this is 30% (The health council of the Nether-
lands, 2001). As the modern western diet contains much more pro-
tein than needed (Westhoek et al., 2011), we did not take this
difference into account. It could be argued that this does not apply
to large part of the global population, and even part of the western
population. For people who barely get enough protein in their diet
the quality does matter, and the intake of animal sources can be
lower than vegetal sources. Especially for poultry and some sea-
food protein the environmental ‘distance’ to vegetal sources then
gets very small.
All in all vegetal sources of protein, animal products with high
feed efficiencies and some types of seafood offer chances to
D. Nijdam et al. / Food Policy 37 (2012) 760–770 767
mitigate climate change. Vegetal sources also have the merits of
less eutrophication, less use of pesticides, and less land use, and
do not contribute to animal welfare problems.
LCA methodology
In all studies on meat and milk, the application of the manure is
unanimously attributed to the meat and dairy life cycle in which
this manure is produced. It would however seem reasonable to
attribute some of these emissions to the next cycle, as the manure
application is part of the beginning of a new cycle, whichever prod-
uct this cycle delivers. Taking both the field emissions due to crop
fertilization of feed crops at the ‘cradle’ of the life cycle and manure
application at the other end of the life cycle into account in one life
cycle, results in double counting. The effects of this may be limited,
as any overlap would only concern the field emissions from man-
ure, and not from artificial fertilizer. In The EU agricultural soils re-
ceive 70% more nitrogen from the application of artificial fertilizer
than from manure (Sutton et al., 2011). In South American soy
fields the share of nitrogen from manure is even much lower
(Smaling et al., 2008).
Land use was included by many studies, either as an indicator of
a scarce resource or as an indicator of loss of biodiversity. Specify-
ing the quality and/or intensity of land use, as was done in some
studies by distinguishing several types of land use, seems an
important step forward. An overview of the issues here and an
applicable method for calculating biodiversity impacts was given
by Schmidt (2008).Milà i Canals et al. (2007) proposed an applica-
ble method for including impacts on ‘life support functions’ of land.
In some seafood LCAs the disturbed area of the seabed by bottom
trawling was included (Vázquez-Rowe et al., 2011;Ramos et al.,
2011;Ziegler and Valentinsson, 2008).
An important methodological issue that we encountered in the
LCAs studied is the broad use of default factors for the important
emissions of N
2
O from fertilized soils, rather than actual emission
based on measurements.
The inclusion of changes in soil organic carbon (SOC), mainly
due to land conversion processes less than a hundred years ago,
and the consequent permanent blocking of the natural build-up
of soil organic carbon by permanent land use, may have significant
effects on the outcomes (Blonk et al., 2008; Nguyen et al., 2010;
Flysjö et al., 2012).
Inclusion of carbon field emissions requires detailed local data,
which are usually not available. Also it is difficult to attribute part
of the process of gradual reduction in soil organic carbon to one
season of cultivation of a specific crop. Often a depreciation period
of 20–100 years are used to allocate the carbon emission to 1 year’s
crop (Nguyen et al., 2010). The blocking of the natural build-up of
SOC by land occupation, also referred to as ‘missed potential car-
bon sink’ can be taken into account more easily. Based on some re-
gional characteristics, a modeling approach is presented by
Schmidinger and Stehfest (2012).Williams et al.(2006) mentions
methane emissions from the soil. With an amount of 0.65 kg ha
1
,
this emission corresponds to only 0.2% of the total carbon footprint
of pork.
From the above it may be clear that soil emissions and soil car-
bon sinks, both real and ‘hypothetical’ (i.e. avoided sinks) should be
investigated further.
Most LCAs represent the situation in a given year. In modeled
production chains sometimes multi-year averages were used, e.g.
for grain yields. As agricultural- and fishing yields can vary
strongly in time, the outcomes of environmental analyses can also
differ. For example the carbon footprint of mackerel varied with a
factor five in a time period of 8 years (Ramos et al., 2011). Out-
comes of LCA studies can also differ strongly per region, depending
for example on crop yields and feed conversion factors. Lesschen
et al. (2011) offer an integrated approach for the EU, which provide
both a weighted, pan-European average, as well as results for indi-
vidual countries.
Variations in important emission factors in the dairy chain were
analyzed by Flysjö et al. (2011a,b). They calculated the uncertainty
in the carbon footprint of milk from Sweden and New Zealand with
Monte-Carlo simulations and found a range (2.5–97.5% interval) of
0.6–1.52 kg CO
2
-eq kg
1
for NZ milk and 0.83–1.56 for Swedish
milk.
We did not include LCA of promising new products in our over-
view, like laboratory meat or insect protein, as there are no- or lit-
tle LCA-studies available. An exploratory LCA of insects (crickets)
by Blonk et al. (2008) showed relative low scores on carbon foot-
print and land use. Insects are efficient food converters because
they do not use energy to maintain a high body temperature.
Conclusions
In general, from an analysis of life cycle assessment studies, it
can be concluded that food products of animal origin have higher
climate- and land use related impacts than vegetable products.
Meat substitutes containing egg protein, poultry, eggs and some
seafood products also show small carbon footprints. The largest
impacts per kilogram of product was found for ruminant meat,
both in terms of greenhouse gas emissions and in terms of land
use (occupation). Pork has intermediate impacts. Vegetal products,
certain seafood, and poultry products have relatively small carbon-
and land use related ‘price tags’.
If the land use and carbon footprint are expressed per kilogram
of protein instead of kilogram of product, the differences between
products become less. The range in carbon footprints is approxi-
mately 5–750 kg CO
2
-eq kg
1
protein. These extremes do not rep-
resent large production volumes. Nevertheless the ‘best case’
sources of protein have carbon footprints that are many times
smaller than those of the ‘worst case’ sources. Land use (occupa-
tion) ranges from zero for certain seafood to over 2000 m
2
ykg
1
for beef protein. Per unit of protein vegetal products, certain types
of seafood and poultry products have relatively small carbon foot-
prints. Pork and milk proteins have medium carbon footprints,
while beef protein has a relatively large carbon footprint. Much
land is needed to produce beef protein, although the impact of land
use differs strongly between extensively produced beef (demand-
ing mainly grasslands) and pork and poultry (demanding arable
land). Within categories – such as within beef – also large differ-
ences exist.
The differences in scores, both between and within the various
product categories, present opportunities for lowering the carbon
footprint and land use of our diet. Shifts in consumption from
red meats and high impact seafood towards vegetal sources of pro-
tein, white meats, and sustainable seafood products, as well as im-
proved management within production chains offer a large
mitigation potential.
Appendix A. Supplementary data
Supplementary data associated with this article can be found, in
the online version, at http://dx.doi.org/10.1016/j.foodpol.2012.
08.002.
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770 D. Nijdam et al. / Food Policy 37 (2012) 760–770
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Many studies have analysed the environmental impact of vegan, vegetarian, or reduced meat diets. To date, literature has not evaluated how diet shifts affect environmental impacts by utilising portfolios which reflect personal nutrition preferences. Further, changing diets could alter the available land for non-food uses. This paper defines novel diet portfolios to outline alternative diet transitions and choices within the population and finds their effect on greenhouse gas (GHG) emissions, primary energy use, and land use in Germany. The aim of this study is to capture how these diet shifts affect land availability and increase the options for land-based climate change mitigation strategies. To do so, a contextualisation is made to compare the use of freed-up land for afforestation or biomethane production (with and without carbon capture and storage). The investigated diet portfolios lead to a reduction of the investigated impacts (GHG emissions: 7–67%; energy use: 5–46%; land use: 6–64%). Additionally, afforestation of freed-up land from each diet portfolio leads to further emission removals of 4–37%. In comparison, using the land to produce energy crops for biomethane production could lead to 2–23% further CO²-eq emission reductions when replacing fossil methane. If biomethane production is paired with carbon capture and storage, emission abatement is increased to 3–34%. This research indicates various short-term pathways to reduce GHG emissions with portfolio diet shifts. Utilising freed-up land for climate change mitigation strategies could prove essential to meet climate targets, but trade-offs with, e.g. biodiversity and ecosystem services exist and should be considered.
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Carbon footprint analysis for different freshwater aquaculture practices in different states of India along with Bay of Bengal was done. In West Bengal, the average inputs as carbon equivalent (CE) were [Formula: see text][Formula: see text]kg/ha with an average output of [Formula: see text][Formula: see text]kg live weight/ha. In Odisha, the inputs as CE were [Formula: see text][Formula: see text]kg/ha with an output of [Formula: see text][Formula: see text]kg live weight/ha. In Andhra Pradesh, the average inputs as CE was [Formula: see text][Formula: see text]kg/ha with an average output of [Formula: see text][Formula: see text]kg live weight/ha for Litopenaeus vannamei culture practices, while the same for Indian major carps culture was [Formula: see text][Formula: see text]kg/ha with an output of [Formula: see text][Formula: see text]kg live weight/ha, respectively. The average Global Warming Potential (GWP) of all the carp culture practices in different study places was 1.57[Formula: see text]kg CO 2 -e/kg fish, while the same for [Formula: see text] vannamei culture was 1.87[Formula: see text]kg CO 2 -e/kg shrimp.
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Purpose The study uses machine learning techniques to cluster regional retail egg prices after 2000 in China. Furthermore, it combines machine learning results with econometric models to study determinants of cluster affiliation. Eggs are an inexpensiv, nutritious and sustainable animal food. Contextually, China is the largest country in the world in terms of both egg production and consumption. Regional clustering can help governments to imporve the precision of price policies and help producers make better investment decisions. The results are purely driven by data. Design/methodology/approach The study introduces dynamic time warping (DTW) algorithm which takes into account time series properties to analyze provincial egg prices in China. The results are compared with several other algorithms, such as TADPole. DTW is superior, though it is computationally expensive. After the clustering, a multinomial logit model is run to study the determinants of cluster affiliation. Findings The study identified three clusters. The first cluster including 12 provinces and the second cluster including 2 provinces are the main egg production provinces and their neighboring provinces in China. The third cluster is mainly egg importing regions. Clusters 1 and 2 have higher price volatility. The authors confirm that due to transaction costs, the importing areas may have less price volatility. Practical implications The machine learning techniques could help governments make more precise policies and help producers make better investment decisions. Originality/value This is the first paper to use machine learning techniques to cluster food prices. It also combines machine learning and econometric models to better study price dynamics.
Technical Report
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The agriculture in Germany accounts for more than 13 percent of the German total greenhouse gas emissions. In climate protection strategies (i.e. the energy and climate programme of the federal government), the contribution of the agriculture is usually still neglected. Therefore, the purpose of this study is the evaluation of climate impacts of the agricultural production in Germany, with respect to the most important agricultural products – wheat, pork, beef and milk. The research focuses on, to what extent conventional and organic farming are different in their climate impacts and which advantages and disadvantages can be found in different systems. The performances of the climate assessment show that organic farming normally is more climate friendly than conventional agriculture. That primarily results from large amounts of mineral fertilizer used in the conventional agriculture which causes high greenhouse gas emissions during production and application. On the other hand, the demand for space throughout ecological production processes is higher than in conventional systems. Furthermore, a significant potential for climate protection can be seen in the water logging of drained moorland whose current agricultural utilization leads to extensive greenhouse gas emissions. Altogether, the agriculture could contribute to the attainment of Germany’s climate goals. This could be achieved through changes in production methods as well as abandoning or extensification of the used moorland areas. For this purpose, the study identifies central starting points as well as discusses potential synergy effects and conflicts with environmental and animal protection goals. Although the study focusses on the German agricultural sector, most of its conclusions can be transferred also to other countries where agriculture produces at similar levels of intensity. The central recommendations, namely conversion from conventional high intensity of fertilizer use to organic farming or other practices with lower intensities, the re-wetting of drained moorland, hold to a similar degree for agriculture worldwide. And another central conclusion is valid internationally as well: The conversion to a more climate friendly agriculture will not be possible without changing consumption patterns to reduce the demand for meat and milk products in favor of vegetarian products. This is mainly a challenge in industrialized countries with highly climate-unfriendly consumption patterns.
Book
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This report first presents a systematic overview of the life cycle of meat and dairy products and their environmental impacts, covering the full food chain. It goes on to provide a comprehensive analysis of the improvement options that allow reducing the environmental impacts throughout the life cycle. Finally, the report assesses the different options regarding their feasibility as well as their potential environmental and socioeconomic benefits and costs. The report focuses on improvement options in three main areas: • Household improvements, mainly to reduce food losses (wastage) and to reduce car use for shopping; • Agricultural improvements, mainly to reduce water and air emissions (in particular nitrate, ammonia and methane) and land requirements; • Power savings in farming, food industry, retail, catering, and for household appliances. The study presents the consequences that the adoption of these options might have on a broad range of different environmental issues, including global warming, eutrophication, respiratory health impacts, etc. It shows that when all environmental improvement potentials are taken together, the aggregated environmental impacts (external costs) of meat and dairy products may be reduced by about 20 %. The study has also quantified the economic costs and benefits of implementing the different options.
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Environmental life cycle assessment of Norway lobster (Nephrops norvegicus) caught along the Swedish west coast by creels and conventional trawls—LCA methodology with case study Abstract Background, aim, and scope Two fishing methods, creeling and conventional trawling, are used to target Norway lobster (Nephrops norvegicus), economically the second most important species in Swedish west coast fisheries. The goal was to evaluate overall resource use and environmen-tal impact caused by production of this seafood with the two different fishing methods using life cycle assessment (LCA) methodology. Materials and methods The inventory covered the entire chain starting by production of supply materials and the fishery itself, through seafood auctioning, wholesaling, retailing, to the consumer. That portion of the life cycle occurring on land was assumed to be identical for Norway lobsters regardless as to how they were caught. The functional unit was 300 g of edible meat (i.e., Norway lobster tails), corresponding to 1 kg of whole, boiled Norway lobsters. The seafloor impact of trawling was quantified using a recently developed methodology. Results Major differences were found between the fishing methods with regard to environmental impact: creeling was found to be more efficient than conventional trawling in all traditional impact categories and in the two additional fishery-related categories involving seafloor impact and discarding. Since the quality of the creel-caught Nephrops was higher, the difference was probably even higher than indicated here. Discussion Major improvement potential was identified in the more widespread use of creels and species-selective trawls. The only deficiencies of creel fishing were poorer working environment and safety, and a potentially higher risk of recruitment overfishing. However, these issues could be handled by technological development and fisheries regulations and should not hamper the development of creel fishery. Conclusions Improvement options were identified and quantified for the Swedish Nephrops fishery. The study demonstrates how LCA can be used to compare the envi-ronmental performance of different segments of a fishery. Recommendations and perspectives Shifting to creeling and species-selective trawling would lead to considerably lower discard, fuel use, and seafloor impact while providing consumers with the same amount of Norway lobsters.
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"Principles of Soil Management and Conservation" comprehensively reviews the state-of-knowledge on soil erosion and management. It discusses in detail soil conservation topics in relation to soil productivity, environment quality, and agronomic production. It addresses the implications of soil erosion with emphasis on global hotspots and synthesizes available from developed and developing countries. It also critically reviews information on no-till management, organic farming, crop residue management for industrial uses, conservation buffers (e.g., grass buffers, agroforestry systems), and the problem of hypoxia in the Gulf of Mexico and in other regions. This book uniquely addresses the global issues including carbon sequestration, net emissions of CO2, and erosion as a sink or source of C under different scenarios of soil management. It also deliberates the implications of the projected global warming on soil erosion and vice versa. The concern about global food security in relation to soil erosion and strategies for confronting the remaining problems in soil management and conservation are specifically addressed. This volume is suitable for both undergraduate and graduate students interested in understanding the principles of soil conservation and management. The book is also useful for practitioners, extension agents, soil conservationists, and policymakers as an important reference material.
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Jerome believed that the task of the commentator was to convey what others have said, not to advance his own interpretations. However, an examination of his commentaries on the Prophets shows that their contents are arranged so as to construct a powerful, but tacit, position of authority for their compiler. By juxtaposing Jewish and Greek Christian interpretations as he does, Jerome places himself in the position of arbiter over both exegetical traditions. But because he does not explicitly assert his own authority, he can maintain a stance of humility appropriate for a monk. Here, Jerome may have been a more authentic representative of the tradition of Origen than was his rival, for all that he was willing to abjure Origen's theology.
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Two most critical factors to address in environmental system analysis of future milk production are 1) the link between milk and beef production, and 2) the competition for land, possibly leading to land use change (LUC) with greenhouse gas (GHG) emissions and loss of biodiversity as important implications. Different methodological approaches concerning these factors, in studies on environmental impacts of dairy production, sometimes lead to contradictory results.Increasing milk yield per cow is often one of the solutions discussed in order to reduce GHG emissions from milk production. However, when also accounting for other systems affected (e.g. beef production) it is not certain that an increase in milk yield per cow leads to a reduction in total GHG emissions per kg milk. In the present study the correlation between carbon footprint (CF) of milk and the amount of milk delivered per cow is investigated for 23 dairy farms (both organic and conventional) in Sweden. Use of a fixed allocation factor of 90% (based on economic value) indicates a reduction in CF with increased milk yield, while no correlation can be noted when system expansion is applied. The average CF for two groups of farms, organic and high yielding conventional, is also calculated. When conducting system expansion the CF is somewhat lower for the organic farms (which have a lower milk yield per cow, but more meat per kg milk), but when a 90% allocation factor is used, the CF is somewhat higher for the organic farms compared to the high yielding conventional farms. In analysis of future strategies for milk production, it is suggested that system expansion should be applied, in order to also account for environmental impacts from affected systems. Thus, scenarios for milk and meat production should be analysed in an integrated approach in order to reduce total emissions from the livestock sector.How to account for emissions from LUC is highly debated and there is no current shared consensus. Different LUC methods result in significantly different results. In this study, four different LUC methods are applied, using data for organic milk production and high yielding conventional milk production systems in Sweden. Depending on which LUC method was applied, the organic system showed about 50% higher or 40% lower CF compared to the conventional high yielding system. Thus, when reporting CF numbers, it is important to report LUC-factors separately and clearly explain the underlying assumptions, since the method of accounting for LUC can drastically change the results.