The price of protein: Review of land use and carbon footprints from life cycle
assessments of animal food products and their substitutes
, Trudy Rood, Henk Westhoek
PBL Netherlands Environmental Assessment Agency, Bilthoven/The Hague, The Netherlands
Received 13 January 2012
Received in revised form 13 July 2012
Accepted 7 August 2012
Life cycle assessment (LCA)
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
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-
Ó2012 Elsevier Ltd. All rights reserved.
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
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
– 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).
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.
Corresponding author. Address: PBL Netherlands Environmental Assessment
Agency, PO Box 303, NL-3720 AH Bilthoven, The Netherlands. Tel.: +31 302742223.
E-mail address: firstname.lastname@example.org (D. Nijdam).
Food Policy 37 (2012) 760–770
Contents lists available at SciVerse ScienceDirect
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,
– 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
– Some effects have to be discounted over a certain period of
time, such as those on biodiversity and changes in soil organic
– 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.
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
Overview of LCA studies reviewed.
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)
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
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
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
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
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
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
, whereas the
Swedish mix, as used by Cederberg and Flysjö (2004a) resulted in
only 45 g CO
, 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-
In the impact assessment phase of LCAs greenhouse gases are
aggregated to CO
equivalents by GWP (Global Warming Potential)
factors. About half of the studies used the ‘old’ IPCC factors (21 for
and 310 for N
O) (IPCC, 2006), and the other half used the
more recent equivalence factors (25 for CH
and 298 for N
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
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
meat was assumed for
the slaughterhouse and 0.1 kg kg
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
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
for canned mussels (Iribarren et al., 2010a)to
0.9 kg CO
for fresh lobster (Ziegler and Valentinsson,
2008). For frozen cod ﬁllet Ziegler et al. (2003) found 0.6 kg
, and Blonk et al. (2009) found 0.5 kg CO
We used a value of 0.5 kg CO
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.
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
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
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
) Land use (m
) Of which grassland (m
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
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
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
milk (12 studies). Weiske et al. (2006) give an average
of 1.4 kg CO
for milk for the EU-15. In a study by the FAO
(2010), an average of 1.3 kg CO
is calculated for western
Europe. The differences can be traced back to soil condition and
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
meat, which are
reportedly mostly due to the N
O emissions from feed production.
Considerably higher values were presented only by Williams et al.
(2006), with 8.7 kg CO
, and Zhu and Van Ierland (2004),
with 10.6 kg CO
.Zhu and Van Ierland (2004) state that
the high value is caused by a very high energy related CO
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
emission (ca 4 kg CO
) is re-
placed by the value from Kool et al. (2009) (ca 0.5 kg CO
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
emission (2.5 CO
), 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
chicken meat. Only Williams
et al. (2006), presented a much higher outcome (6 kg CO
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
), and the largest by Williams
et al. (2006) (5.5 kg CO
). 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
-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
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
ﬁllet from the Basque
purse seine ﬂeet to over 6 kg CO
ﬁllet from the Galician
bottom trawl ﬂeet. Alaskan pollack also has a relatively small car-
bon footprint (about 3 kg CO
ﬁllet) (Blonk et al., 2009).
Cod shows ranges from 3 to 7 kg CO
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
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
ﬁllet) and has
a land use of 5 m
ﬁllet (Blonk et al., 2009). Land-farmed
trout (two studies) has a carbon footprint of about 7 kg
Vegetarian products (meat substitutes, 1 study, n= 3) have car-
bon footprints ranging from less than 1 kg CO
for a vege-
tal meat substitute to 6 CO
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
for peas and soya (Sheenan et al.,
1998; Nemecek et al., 2005) to 2 kg CO
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
-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
protein for vegetal
products and meat substitute with egg protein to over 2000 m
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.
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
protein Land use m
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
Distribution of the carbon footprint of pork (kg CO
retail meat and percentages).
Kool et al. (2009) Basset-mens and Van der Werf (2005) Eriksson et al. (2005)
-eq % kg CO
-eq % kg CO
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
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
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
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
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
retail meat): 61% for enteric fer-
mentation, 27% for feed production and 12% for manure
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–
, 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
Distribution of the carbon footprint for beef (kg CO
-eq per kg retail meat and percentages) of various production systems.
Suckler calves extensive
Suckler calves semi-extensive
Mixed calves intensive
Culled-dairy cows intensive
Enteric fermentation 44.3 75 19.0 50 6.1 38 4.0 40
from manure 1.2 2 3.8 10 2.2 14 1.1 11
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
Blonk et al. (2008).
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).
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.,
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
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.
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.
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
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
this emission corresponds to only 0.2% of the total carbon footprint
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
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-
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
for NZ milk and 0.83–1.56 for Swedish
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.
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
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
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-
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
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.
Alkemade, R., van Oorschot, M., Miles, L., Nellemann, C., Bakkenes, M., ten Brink, B.,
2009. GLOBIO3: a framework to investigate options for reducing global
terrestrial biodiversity loss. Ecosystems 12, 374–390.
Andersson, K., 2000. LCA of food products and production systems. The
International Journal of Life Cycle Assessment 5, 239–248.
768 D. Nijdam et al. / Food Policy 37 (2012) 760–770
Aubin, J., Papatryphon, E., van der Werf, H.M.G., Chatzifotis, S., 2009. Assessment of
the environmental impact of carnivorous ﬁnﬁsh production systems using life
cycle assessment. Journal of Cleaner Production 17, 354–361.
Ayer, N.W., Tyedmers, P.H., Pelletier, N.L., Sonesson, U., Scholz, A., 2007. Co-product
allocation in life cycle assessments of seafood production systems: review of
problems and strategies. International Journal of Life Cycle Assessment 12, 480–
Basset-Mens, C., van der Werf, H.M.G., 2005. Scenario-based environmental
assessment of farming systems: the case of pig production in France.
Agriculture, Ecosystems & Environment 105, 127–144.
Berlin, J., 2002. Environmental life cycle assessment of Swedish semi-hard cheese.
International Dairy Journal 12, 939–953.
Blanco-Canqui, H., Lal, R., 2008. Principles of Soil Conservation and Management.
Blonk, H., Kool, A., Luske, B., 2008. Milieueffecten van Nederlandse consumptie van
eiwitrijke producten (in Dutch, Environmental effects of Dutch consumption of
protein-rich products). BMA/VROM, Gouda.
Blonk, H., Luske, B., Kool, A., 2009. Milieueffecten van enkele populaire vissoorten
(in Dutch, Environmental effects of some popular ﬁsh species). BMA/VROM,
Brander, M., Tipper, R., Hutchinson, C., Davis, G., 2009. Consequential and
Attributional Approaches to LCA: A Guide to Policy Makers with Speciﬁc
Reference to GHG LCA of Biofuels. Ecometrica Press, Edinburgh.
BSI, 2008. Speciﬁcation for the Assessment of the Life Cycle Greenhouse Gas
Emissions of Goods and Services. British Standards Institute, London.
Carlsson-Kanyama, A., Faist, M., 2000. Energy Use in the Food Sector. ETH, Zurich.
Casey, J.W., Holden, N.M., 2005. Analysis of Greenhouse Gas Emissions from the
average Irish Milk Production System. Agricultural Systems 86, 97–114.
Casey, J.W., Holden, N.M., 2006. Greenhouse gas emissions from conventional, agri-
environmental scheme, and organic Irish suckler-beef units. Journal of
Environmental Quality 35, 231–239.
Cederberg, C., Flysjö, A., 2004a. Environmental Assessment of Future Pig Farming
Systems – Quantiﬁcation of Three Scenarios from the FOOD 21 Synthesis Work.
SIK Report 723. SIK, Göteborg.
Cederberg, C., Flysjö, A., 2004b. Life Cycle Inventory of 23 Dairy Farms in South-
Western Sweden. SIK Report No. 728. SIK, Göteborg.
Cederberg, C., Stadig, M., 2003. System expansion and allocation in life cycle
assessment of milk and beef production. International Journal of Life Cycle
Assessment 8, 350–356.
Cederberg, C., Meyer, D., Flysjö, A., 2009a. Life Cycle Inventory of Greenhouse Gasses
and Use of Land and Energy in Brazilian Beef Production. SIK Report No. 792.
Cederberg, C., Flysjö, A., Sonesson, U., Sund, V., Davis, J., 2009b. Greenhouse Gas
Emissions from Swedish Consumption of Meat, Milk and Eggs, 1990 and 2005.
SIK Report No. 793. SIK, Göteborg.
Davies, R.W.D., Cripps, S.J., Nickson, A., Porter, G., 2009. Deﬁning and estimating
global marine ﬁsheries bycatch. Marine Policy 33, 661–672.
De Vries, M., de Boer, I.J.M., 2010. Comparing environmental impacts for livestock
products: a review of life cycle assessments. Livestock Science 128, 1–11.
Ecoinvent-centre, 2009. The Life Cycle Inventory Data v2.1. Swiss Centre for Life
Cycle Inventories, St. Gallen.
Edwards-Jones, G., Plassmann, K., Harris, I.M., 2009. Carbon footprinting of lamb
and beef production systems: insights from an empirical analysis of farms in
Wales, UK. Journal of Agricultural Science 147, 707–719.
Ekvall, T., Weidema, B., 2004. System boundaries and input data in consequential
life cycle inventory analysis. The International Journal of Life Cycle Assessment
Ellingsen, H., Aanondsen, S.A., 2006. Environmental impacts of wild caught cod and
farmed salmon – a comparison with chicken. International Journal of Life Cycle
Assessment 11, 60–65.
Ellingsen, H., Olaussen, J.O., Utne, I.B., 2009. Environmental analysis of the
Norwegian ﬁshery and aquaculture industry – a preliminary study focusing
on farmed salmon. Marine Policy 33, 479–488.
Eriksson, I.S., Elmquist, H., Stern, S., Nybrant, T., 2005. Environmental systems
analysis of pig production: the impact of feed choice. International Journal of
Life Cycle Assessment 10, 143–154.
FAO, 2009. The State of World Fisheries and Aquaculture 2008. FAO, Rome.
FAO, 2010. Greenhouse Gas Emissions from the Dairy Sector – A Life Cycle
Assessment. FAO’s Animal Production and Health Division, Rome.
Flachowsky, G., Hachenberg, S., 2009. CO
footprints for food of animal origin –
present stage and open questions. Journal für Verbraucherschutz und
Lebensmittelsicherheit 4, 190–198.
Flysjö, A., Cederberg, C., Henriksson, M., Ledgard, S., 2011a. How does co-product
handling affect the carbon footprint of milk? Case study of milk production in
New Zealand and Sweden. The International Journal of Life Cycle Assessment
Flysjö, A., Henriksson, M., Cederberg, C., Ledgard, S., Englund, J.E., 2011b. The impact
of various parameters on the carbon footprint of milk production in New
Zealand and Sweden. Agricultural Systems 104, 459–469.
Flysjö, A., Cederberg, C., Henriksson, M., Ledgard, S., 2012. The interaction between
milk and beef production and emissions from land use change – critical
considerations in life cycle assessment and carbon footprint studies of milk.
Journal of Cleaner Production 28, 134–142.
Garnett, T., 2008. Cooking up a Storm: Food, Greenhouse Gas Emissions and Our
Changing Climate. Food and Climate Research Network, Centre for
Environmental Strategy. University of Surrey, Guildford.
González, A.D., Frostell, B., Carlsson-Kanyama, A., 2011. Protein efﬁciency per unit
energy and per unit greenhouse gas emissions: Potential contribution of diet
choices to climate change mitigation. Food Policy 36, 562–570.
Gronroos, J., Seppala, J., Silvenius, F., Makinen, T., 2006. Life cycle assessment of
Finnish cultivated rainbow trout. Boreal Environment Research 11, 401–414.
Guinée, J.B., Gorrée, M., Heijungs, R., Huppes, G., Kleijn, R., De Koning, A., Van Oers,
L., Wegener Sleeswijk, A., Suh, S., Udo de Haes, H.A., De Bruijn, H., Van Duin, R.,
Huijbregts, M.A.J., 2002. Handbook on Life Cycle Assessment. Operational Guide
to the ISO Standards. I: LCA in Perspective. IIa: Guide. IIb: Operational Annex.
III: Scientiﬁc Background. Kluwer Academic Publishers, Dordrecht.
Haas, G., Wetterich, F., Köpke, U., 2001. Comparing intensive, extensiﬁed and
organic grassland farming in southern Germany by process life cycle
assessment. Agriculture, Ecosystems & Environment 83, 43–53.
Harman, J., Garett, A., Anton, S., Tyedmers, P., 2008. CO
Emissions, Case Studies in
Selected Seafood Product Chains. Brieﬁng Paper. SEAFISH, Grimsby.
Haxsen, G., 2008. Calculating Costs of Pig Production with the InterPIG Network.
Johann Heinrich von Thuenen-Institut (vTI), Federal Research Institute for Rural
Areas, Forestry and Fisheries Institute of Farm Economics, Institute of Market
Analysis and Agricultural Trade Policy.
Hirschfeld, J., Weiss, J., Preidl, M., Korbun, T., 2008. Klimawirkungen der
Landwirtschaft in Deutschland (in German; Climate impacts of German
agriculture). Schriftenreihe des IOW 186/08. IOW, Berlin.
IPCC, 2006. 2006 IPCC Guidelines for National Greenhouse Gas Inventories.
Prepared by the National Greenhouse Gas Inventories Programme. Institute
for Global Environmental Strategies, Kanagawa.
Iribarren, D., Hospido, A., Moreira, M.T., Feijoo, G., 2010a. Carbon footprint of canned
mussels from a business-to-consumer approach: a starting point for mussel
processors and policy makers. Environmental Science & Policy 13, 509–521.
Iribarren, D., Vázquez-Rowe, I., Hospido, A., Moreira, M.T., Feijoo, G., 2010b.
Estimation of the carbon footprint of the Galician ﬁshing activity (NW Spain).
Science of the Total Environment 408, 5284–5294.
Katajajuuri, J., 2007. Experiences and Improvement Possibilities – LCA Case Study of
Broiler Chicken Production. MTT Agrifood Research Finland, Jokioinen.
Kool, A., Blonk, H., Ponsioen, T., Sukkel, W., Vermeer, H., De Vries, J., Hoste, R., 2009.
Carbon Footprints van conventioneel en biologisch varkensvlees (in Dutch,
Carbon Footprints of conventional and organic pork). BMA/WUR, Gouda.
Kramer, K.J., 2000. Food Matters: On Reducing Energy Use and Greenhouse Gas
Emission from Household Food Consumption. University of Groningen.
Lesschen, J.P., van den Berg, M., Westhoek, H.J., Witzke, H.P., Oenema, O., 2011.
Greenhouse gas emission proﬁles of European livestock sectors. Animal Feed
Science and Technology 166–167, 16–28.
Milà i Canals, L., Romanyà, J., Cowell, S.J., 2007. Method for assessing impacts on life
support functions (LSF) related to the use of ‘fertile land’ in Life Cycle
Assessment (LCA). Journal of Cleaner Production 15, 1426–1440.
Marshall, B.M., Levy, S.B., 2011. Food Animals and Antimicrobials: Impacts on
Human Health. Clinical Microbiology Reviews 24, 718–733.
Mollenhorst, H., Berentsen, P.B.M., De Boer, I.J.M., 2006. On-farm quantiﬁcation of
sustainability indicators: an application to egg production systems. British
Poultry Science 47, 405–417.
Nemecek, T., Huguenin-Elie, O., Dubois, D., Gaillard, G., 2005. Okobilanzierung von
Anbausystemen in schweizerischen Acker- und Futterbau. FAL, Zurich.
NEVO, 2010. Dutch Food Composition Database (NEVO-online) v 2011/3.0 Stichting
Nguyen, T.L.T., Hermansen, J.E., Mogensen, L., 2010. Environmental consequences of
different beef production systems in the EU. Journal of Cleaner Production 18,
Ogino, A., Orito, H., Shimada, K., Hirooka, H., 2007. Evaluating environmental
impacts of the Japanese beef cow-calf system by the life cycle assessment
method. Animal Science Journal 78, 424–432.
Pelletier, N.L., Ayer, N.W., Tyedmers, P.H., Kruse, S.A., Flysjö, A., Robillard, G., Ziegler,
F., Scholz, A.J., Sonesson, U., 2007. Impact categories for life cycle assessment
research of seafood production systems: review and prospectus. International
Journal of Life Cycle Assessment 12, 414–421.
Pelletier, N., Tyedmers, P., Sonesson, U., Scholz, A., Ziegler, F., Flysjö, A., Kruse, S.,
Cancino, B., Silverman, H., 2009. Not all salmon are created equal: life cycle
assessment (LCA) of global salmon farming systems. Environmental Science and
Technology 43, 8730–8736.
Pelletier, N., Pirog, R., Rasmussen, R., 2010. Comparative life cycle environmental
impacts of three beef production strategies in the Upper Midwestern United
States. Agricultural Systems 103, 380–389.
Peters, G.M., Rowley, H.V., Wiedemann, S., Tucker, R., Short, M.D., Schulz, M., 2009.
Red meat production in Australia: life cycle assessment and comparison with
overseas studies. Environmental Science & Technology 44, 1327–1332.
Phetteplace, H., Johnson, D., Seidl, A., 2001. Greenhouse gas emissions from
simulated beef and dairy livestock systems in the United States. Nutrient
Cycling in Agroecosystems 60, 99–102.
Ponsioen, T., Broekema, R., Blonk, H., 2010. Koeien op gras, milieueffecten van
Nederlandse en buitenlandse rundvleesproductiesystemen (in Dutch, Cows on
Grass, Environmental Effects of Dutch and foreign Beef Production Systems).
Blonk Milieu Advies, Gouda.
Ramos, S., Vázquez-Rowe, I., Artetxe, I., Moreira, M.T., Feijoo, G., Zufía, J., 2011.
Environmental assessment of the Atlantic mackerel (Scomber scombrus) season
in the Basque Country: increasing the timeline delimitation in ﬁshery LCA
studies. International Journal of Life Cycle Assessment 16, 599–610.
Rockstrom, J., Steffen, W., Noone, K., Persson, A., Chapin, F.S., Lambin, E.F., Lenton,
T.M., Scheffer, M., Folke, C., Schellnhuber, H.J., Nykvist, B., de Wit, C.A., Hughes,
D. Nijdam et al. / Food Policy 37 (2012) 760–770 769
T., van der Leeuw, S., Rodhe, H., Sorlin, S., Snyder, P.K., Costanza, R., Svedin, U.,
Falkenmark, M., Karlberg, L., Corell, R.W., Fabry, V.J., Hansen, J., Walker, B.,
Liverman, D., Richardson, K., Crutzen, P., Foley, J.A., . A safe operating space for
humanity. Nature 461, 472–475.
Roy, P., Nei, D., Orikasa, T., Xu, Q., Okadome, H., Nakamura, N., Shiina, T., 2009. A
review of life cycle assessment (LCA) on some food products. Journal of Food
Engineering 90, 1–10.
Schmidinger, K., Stehfest, E., 2012. Including CO
implications of land occupation in
LCAs-method and example for livestock products. International Journal of Life
Cycle Assessment, 1–11.
Schmidt, J.H., 2008. Development of LCIA characterisation factors for land use
impacts on biodiversity. Journal of Cleaner Production 16, 1929–1942.
Sevenster, M., DeJong, F., 2008. A Sustainable Dairy Sector. Global, Regional and Life
Cycle Facts and Figures on Greenhouse-gas0020Emissions. Commissioned by
the European Dairy Association, CE, Delft..
Sheane, R., Lewis, K., Hall, P., Holmes-ling, P., Kerr, A., Steward, K., Webb, D., 2011.
Identifying Opportunities to Reduce the Carbon Footprint Associated with the
Scottish Dairy Supply Chain – Main Report. The Scottish Government,
Sheenan, J., Camobreco, V., Dufﬁeld, J., Graboski, M., Shapouri, H., 1998. Life Cycle
Inventory of Biodiesel and Petroleum Diesel for Use in an Urban Bus. In: Lab,
N.R.E. (Ed.), USDE Ofﬁce of Fuels Development and USDA Ofﬁce of Energy.
Silvenius, F., Grönroos, J., 2003. Fish Farming and the Environment: Results of
Inventory Analysis. Finnish Environment Institute, Helsinki.
Smaling, E.M.A., Roscoe, R., Lesschen, J.P., Bouwman, A.F., Comunello, E., 2008. From
forest to waste: assessment of the Brazilian soybean chain, using nitrogen as a
marker. Agriculture, Ecosystems & Environment 128, 185–197.
Solomon, S., Qin, D., Manning, M., Chen, Z., Marquis, M., Averyt, K., B, Tignor, M.,
Miller, H., L, 2007. Contribution of working group I to the fourth assessment
report of the intergovernmental panel on climate change. Cambridge University
Spielman, M., Bauer, C., Dones, R., M, T., 2007. Ecoinvent, Transport Services Data
v2.0. Ecoinvent Report No. 14. Swiss Centre for Life Cycle Inventories, St. Gallen/
Stehfest, E.; van den Berg, M., Woltjer, G., Msangi, S., Westhoek, H., 2012. Options to
reduce the environmental effects of livestock production - comparison of two
economic models, accepted for publication in Agricultural Systems.
Steinfeld, H., Gerber, P., Wassenaar, T., Castel, V., Rosales, M., de Haan, C., 2006.
Lifestock’s Long Shadow: Environmental Issues and Options. FAO, Rome.
Suh, S., Weidema, B., Schmidt, J.H., Heijungs, R., 2010. Generalized make and use
framework for allocation in life cycle assessment. Journal of Industrial Ecology
Sutton, M., Howard, C., Erisman, J.W., Billen, G., Bleeker, A., Grennfelt, P., Van
Grinsven, H., Grizetti, B. (Eds.), 2011. The European Nitrogen Assessment.
Cambridge University Press, Cambridge.
Svanes, E., Vold, M., Hanssen, O.J., 2011a. Effect of different allocation methods on
LCA results of products from wild-caught ﬁsh and on the use of such results.
International Journal of Life Cycle Assessment 16, 512–521.
Svanes, E., Vold, M., Hanssen, O.J., 2011b. Environmental assessment of cod (Gadus
morhua) from autoline ﬁsheries. International Journal of Life Cycle Assessment
The Health Council of the Netherlands, 2001. Voedingsnormen energie, eiwitten,
vetten en verteerbare koolhydraten (in Dutch: Reference intakes for energy,
protein, fat and digestible carbohydrates). The Health Council of the
Netherlands, The Hague.
Thomassen, M.A., Dalgaard, R., Heijungs, R., De Boer, I., 2008a. Attributional and
consequential LCA of milk production. International Journal of Life Cycle
Assessment 13, 339–349.
Thomassen, M.A., van Calker, K.J., Smits, M.C.J., de Iepema, G.L., Boer, I.J.M., 2008b.
Life cycle assessment of conventional and organic milk production in the
Netherlands. Agricultural Systems 96, 95–107.
Thrane, M., 2006. LCA of Danish ﬁsh products: new methods and insights.
International Journal of Life Cycle Assessment 11, 66–74.
Trujilo, P., 2008. Using a mariculture sustainability index to rank countries
performances. In: Pauly, D.J.A. (Ed.), A Comparative Assessment of Biodiversity,
Fisheries and Aquaculture in 53 Countries’ Exclusive Economic Zones. The
Fisheries Centre, University of British Colombia, Vancouver, pp. 28–56.
Tyedmers, P., 2004. Fisheries and Energy Use: Encyclopedia of Energy. Elsevier.
Vazquez-Rowe, I., Moreira, M.T., Feijoo, G., 2010. Life cycle assessment of horse
mackerel ﬁsheries in Galicia (NW Spain): comparative analysis of two major
ﬁshing methods. Fisheries Research 106, 517–527.
Vázquez-Rowe, I., Moreira, M.T., Feijoo, G., 2011. Life Cycle Assessment of fresh hake
ﬁllets captured by the Galician ﬂeet in the Northern Stock. Fisheries Research
Vázquez-Rowe, I., Moreira, M.T., Feijoo, G., 2012. Environmental assessment of
frozen common octopus (Octopus vulgaris) captured by Spanish ﬁshing vessels
in the Mauritanian EEZ. Marine Policy 36, 180–188.
Vergé, X.P.C., Dyer, J.A., Desjardins, R.L., Worth, D., 2007. Greenhouse gas emissions
from the Canadian dairy industry in 2001. Agricultural Systems 94, 683–693.
Vergé, X.P.C., Dyer, J.A, Desjardins, R.L, Worth, D., 2008. Greenhouse gas emissions
from the Canadian beef industry. Agricultural Systems 98, 126–134.
Vergé, X.P.C., Dyer, J.A., Desjardins, R.L., Worth, D., 2009. Long-term trends in
greenhouse gas emissions from the Canadian poultry industry. Journal of
Applied Poultry Research 18, 210–222.
Weber, C.L., Matthews, H.S., 2008. Food-miles and the relative climate impacts of
food choices in the United States. Environmental Science and Technology 42,
Weidema, B.P., Wesnaes, M., Hermansen, J., Kristensen, T., Halberg, N., 2008.
Environmental Improvement Potentials of Meat and Dairy Products. EC/JRC/
Weiske, A., Vabitsch, A., Olesen, J.E., Schelde, K., Michel, J., Friedrich, R., Kaltschmitt,
M., 2006. Mitigation of greenhouse gas emissions in European conventional and
organic dairy farming. Agriculture, Ecosystems & Environment 112, 221–232.
Westhoek, H., Rood, T., Van den Berg, M., Janse, J., Nijdam, D., Reudink, M., Stehfest,
E., 2011. The Protein Puzzle. PBL Netherlands Environmental Assessment
Agency, Bilthoven/The Hague.
Williams, A.G., Audsley, E., Sandars, D.L., 2006. Determining the Environmental
Burdens and Resource Use in the Production of Agricultural and Horticultural
Commodities. Natural Resource Management Institute, Cranﬁeld University,
Silsoe Research Institute, Bedford.
Yan, M.J., Humphreys, J., Holden, N.M., 2011. An evaluation of life cycle assessment
of European milk production. Journal of Environmental Management 92, 372–
Zhu, X., Van Ierland, E.C., 2004. Protein chains and environmental pressures: a
comparison of pork and novel protein foods. Environmental Sciences 2003
Ziegler, F., Valentinsson, D., 2008. 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. International
Journal of Life Cycle Assessment 13, 487–497.
Ziegler, F., Nilsson, P., Mattsson, B., Walther, Y., 2003. Life Cycle Assessment of
frozen cod ﬁllets including ﬁshery-speciﬁc environmental impacts.
International Journal of Life Cycle Assessment 8, 39–47.
Ziegler, F., Emanuelsson, A., Eichelsheim, J.L., Flysjö, A., Ndiaye, V., Thrane, M., 2011.
Extended life cycle assessment of southern pink shrimp products originating in
Senegalese artisanal and industrial ﬁsheries for export to Europe. Journal of
Industrial Ecology 15, 527–538.
770 D. Nijdam et al. / Food Policy 37 (2012) 760–770