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Agricultural Systems 194 (2021) 103251
Available online 8 September 2021
0308-521X/© 2021 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).
Review
Manure management and soil biodiversity: Towards more sustainable food
systems in the EU
Julia K¨
oninger
a
,
b
, Emanuele Lugato
b
, Panos Panagos
b
, Mrinalini Kochupillai
c
,
Alberto Orgiazzi
b
, Maria J.I. Briones
a
,
*
a
Departamento de. Ecología y Biología Animal, Universidade de Vigo, 36310 Vigo, Spain
b
European Commission, Joint Research Centre (JRC), Sustainable Resources Directorate, Land Resources Unit, I-21027 Ispra, Italy
c
Department of Aerospace and Geodesy, Technical University of Munich (TUM), 80539 Munich, Germany
HIGHLIGHTS GRAPHICAL ABSTRACT
•More than 1.4 billion t y
−1
of manure
are generated in the EU and UK and re-
applied to soils with potential harmful
effects.
•Current European regulations on
manure disposal are scattered across
different legislations and ignore soil
biodiversity.
•This review of 407 documents revealed
that manure quality is more important
to soil biodiversity than manure
quantity.
•To promote soil biodiversity by manure
amendments, a radical transformation
in the way agriculture is conducted is
needed.
•This approach could reinforce EU MS’
commitment to protect soils and
improve future agricultural policies at
the EU level.
ARTICLE INFO
Editor: Mark van Wijk
Keywords:
Common agricultural policy
Environmental policy
Nutrient losses
Soil organisms
Agricultural practices
Sustainability
European Union
ABSTRACT
CONTEXT: In the European Union (EU-27) and UK, animal farming generated annually more than 1.4 billion
tonnes of manure during the period 2016–2019. Of this, more than 90% is directly re-applied to soils as organic
fertiliser. Manure promotes plant growth, provides nutritious food to soil organisms, adds genetic and functional
diversity to soils and improves the chemical and physical soil properties. However, it can also cause pollution by
introducing toxic elements (i.e., heavy metals, antibiotics, pathogens) and contribute to nutrient losses. Soil
organisms play an essential role in manure transformation into the soil and the degradation of any potential toxic
constitutes; however, manure management practices often neglect soil biodiversity.
OBJECTIVE: In this review, we explored the impact of manure from farmed animals on soil biodiversity by
considering factors that determine the effects of manure and vice versa. By evaluating manure’s potential to
enhance soil biodiversity, but also its environmental risks, we assessed current and future EU policy and
* Corresponding author.
E-mail addresses: julia.koninger@uvigo.es (J. K¨
oninger), Emanuele.LUGATO@ext.ec.europa.eu (E. Lugato), Panos.PANAGOS@ec.europa.eu (P. Panagos), m.
kochupillai@tum.de (M. Kochupillai), Alberto.ORGIAZZI@ec.europa.eu (A. Orgiazzi), mbriones@uvigo.es (M.J.I. Briones).
Contents lists available at ScienceDirect
Agricultural Systems
journal homepage: www.elsevier.com/locate/agsy
https://doi.org/10.1016/j.agsy.2021.103251
Received 16 April 2021; Received in revised form 28 July 2021; Accepted 18 August 2021
Agricultural Systems 194 (2021) 103251
2
legislations with the ultimate aim of providing recommendations that can enable a more sustainable manage-
ment of farm manures.
METHODS: This review explored the relationship between manure and soil biodiversity by considering 407
published papers and relevant legislative provisions. In addition, we evaluated whether benets and risks on soil
biodiversity are considered in manure management. Thereafter, we analysed the current legislation in the Eu-
ropean Union relevant to manure, an important driver for its treatment, application and storage.
RESULTS AND CONCLUSIONS: This review found that coupling manure management with soil biodiversity can
mitigate present and future environmental risks. Our analyses showed that manure quality is more important to
soil biodiversity than manure quantity and therefore, agricultural practices that protect and promote soil
biodiversity with the application of appropriate, high-quality manure or biostimulant preparations based on
manure, could accelerate the move towards more sustainable food production systems. Soil biodiversity needs to
be appropriately factored in when assessing manure amendments to provide better guidelines on the use of
manure and to reduce costs and environmental risks. However, radical changes in current philosophies and
practices are needed so that soil biodiversity can be enhanced by manure management.
SIGNIFICANCE: Manure quality in the EU requires greater attention, calling for more targeted policies. Our
proposed approach could be applied by European Union Member States to include soil protection measures in
national legislation, and at the EU level, can enable the implementation of strategic goals.
1. Introduction
1.1. Manure: Boon or bane?
In European legislation, there is no uniformly used denition for
manure: While Regulation EC/1069/2009 on animal by-products de-
nes manure as “any excrement and/or urine of farmed animals other
than farmed sh, with or without litter”, being an organic fertiliser, the
Nitrates Directive (Directive 91/676/EEC) denes manure from farmed
animals as “waste products excreted by livestock: or a mixture of litter
and waste products excreted by livestock, even in processed form”. The
European Union (EU) produces a signicant amount of organic fertiliser
or waste - depending on the denition to be followed: 1.4 billion tonnes
of manure from farmed animals were produced annually in the period
2016–2019 in the EU27 and UK (Eurostat, 2021a). Six large countries
(DE, ES, FR, IT, PL, UK) produce ca. 68% of the total manure, while
Ireland (84 million tonnes) and the Netherlands (73 million tonnes) also
make important contributions. More than 75% of the produced manure
derives from cattle, while pigs and chickens produced ca. 12% each (Fig.
1). However, the numbers of pigs and poultry/broilers increased
signicantly between 2010 and 2018, +25% and +3%, respectivly
(Eurostat, 2019). Due to increasing demand for meat and animal-based
products, as well as an increase in the export of meat and dairy products,
manure is increasingly generated in highly intensive farming systems
(Bernal et al., 2015; Buckwell and Nadeu, 2016) and 4% of European
farms produced 80% of the total amounts of manure in 2018 (Amann
et al., 2018) (See Fig. 1 and Table 1).
The physicochemical properties of manure justify its wide use as a
soil improver and organic fertiliser (Liang et al., 2014; Liu et al., 2010;
Mandal et al., 2007; Saha et al., 2008). It improves soil physical and
chemical properties (Kheyrodin and Antoun, 2011; Loro et al., 1997;
Rayne and Aula, 2020; Unc and Goss, 2004), since it can slow down the
rate of soil pH decline due to its high buffer capacity (Cai et al., 2015)
and/or contribute to decrease aluminium toxicity (de la Luz Mora et al.,
2017). Some studies also suggest that solid manure can raise soil pH due
to the presence of potassium, sodium, magnesium and calcium (L’Her-
roux et al., 1997), calcium carbonates and bicarbonates (Whalen et al.,
2000) and organic anions (Butterly et al., 2013), increasing the buffer
and cation exchange capacities. However, manure’s effects on soil pH
depend on its initial value (Tang and Yu, 1999), the diet of animals
(Butterly et al., 2013), the amount of manure applied (Hao and Chang,
2002), and its treatment prior to its application (Cavalli et al., 2016;
Cavalli et al., 2017). Manure also provides essential mineral nutrients,
such as inorganic nitrogen in the form of ammonium (NH
4+
; Geisseler
et al., 2010), carbon (Francioli et al., 2016), phosphorus and sulphur
(Liu et al., 2020), and metals such as zinc and copper (Bünemann et al.,
2006; Delgado et al., 2012). Consequently, in a long term perspective,
manure could potentially substitute part of mineral fertilisers, decrease
farmers’ costs and reduce the EU dependency on imports of phosphorus
from other countries (Drangert et al., 2018; Garske et al., 2020).
Importantly, manure substantily increases soil carbon stocks (Liu et al.,
2020) and stabilises organic matter content and C/N ratios in a more
stable fraction over the long term (Cui et al., 2018; Gong et al., 2009;
Zhang et al., 2014b), which is the key driving factor for soil microbial
diversity (Cui et al., 2018).
The nutrient added to soils via manure enhances enzyme (Liang
et al., 2014; Watts et al., 2010) and microbial activity (Watts et al.,
2010), as well as the abundance and biomass of soil fauna (Bengtsson
et al., 2005; Birkhofer et al., 2008; van Eekeren et al., 2009). Benets of
increased fungal diversity by bovine manure application, for example,
include minimized dry rot of potatoes (Gle´
n-Karolczyk et al., 2018). A
meta-analysis found a more active and enhanced size of microbial
populations in manures (pig, cattle, chicken, horse) controlling plant-
feeding nematode populations (Liu et al., 2016). However, fertilising
effects of manure and its inuence on soil biodiversity are highly
dependent on its chemical composition, which varies based on animal
feed and species (Kerr et al., 2006) and treatment/processing methods
(Jørgensen and Jensen, 2009; Risberg et al., 2017). Estimates from 2011
suggest that approximately 92.2% of nutrients in manure are being
returned to elds with little or no processing (Foged et al., 2011). Direct
application of manure is related to insufcient manure storage capac-
ities (Buckwell and Nadeu, 2016). In 2010, only 31% of animal farms
had storage facilities for manure and Germany, Italy, Belgium, Malta,
Cyprus and the Netherlands exceeded recommended thresholds for an-
imal density in relation to manure storage capacities (Eurostat, 2013).
However, the amount of processed manure by waste category (for ani-
mal faeces, urine and manure) increased from 8.4 million tonnes in 2004
to 9.52 million tonnes in 2018 (Eurostat, 2021b), including manure
recycled to compost or anaerobic digestates (Eurostat, 2010).
When the dry mass exceeds 20%, manure is considered solid,
whereas slurry has a dry mass ranging between 4% and 20% and manure
with a dry mass below 4% is considered liquid (Eurostat, 2021c). Since
nitrogen (N) in solid manure is bound to organic matter, its minerali-
sation is a relatively slow process, persisting in soils for up to 5–10 years
after its application (Webb et al., 2013). Liquid manure/urine from
herbivores mainly provides mineral nitrogen in the form of ammonium
(NH
4+
; Webb et al., 2013), which is readily plant-available especially in
N-limited environments (Mooshammer et al., 2014). However, excessive
manure applied during long periods will cause the accumulation of
phosphorus and potassium and excessive N surpluses will cause high
leaching rates of N (Geng et al., 2019). Particularly, when combined N
application rates from manure and other N fertilisers exceed plant de-
mand, N made available can exceed plant uptake capacities, resulting in
increased pollution risks (Geng et al., 2019; Meng et al., 2005). Of all
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
3
agricultural emissions, farmed animal manure contributed 18.6% of
methane (CH
4
) and 11.3% of nitrous oxide (N
2
O) emissions in the
EU27 +UK in 2015 (Eurostat, 2018). Agriculture contributes 93.3% of
the overall ammonia emissions in the EU27 +UK in 2013, whereby
animal manure contributed around 58.9% considering the capture,
storage, treatment and use of animal manure (Eurostat, 2015).
Additional pollution risks from manure result from the substantial
use of animal food supplements (e.g., copper and zinc supplements in
pig and poultry farming) in intensive animal farming that end up in the
manure applied to soils (Moral et al., 2008; Provolo et al., 2018). In the
EU, 150 million pigs consume more than 6.2 million tonnes of copper
through feed additives (Panagos et al., 2018). Continuous additions of
heavy metals contaminate soils, groundwater and affect organisms
living in soils, whereby microorganisms are the rst organisms being
harmed (He et al., 2005). Heavy metals become more mobile in acidic
soils (He et al., 2005), which long term applications of mineral fertilisers
contribute to (Czarnecki and Düring, 2015). Manure may contain
pathogens, which is particularly the case in untreated slurry from pigs
kept in high densities (Guan and Holley, 2003; Unc and Goss, 2004;
Venglovsky et al., 2018). Via manure applied to soils, zoonotic
pathogens may enter the human food chain (Bicudo and Goyal, 2003).
Antibiotics, used to maintain animal health and to indirectly boost an-
imal growth (Kümmerer, 2009), also reach soils via manure. Various
antibiotics in manure are mobile (Heuer et al., 2011; Zhao et al., 2019),
harming plants (Zhou et al., 2020) and invertebrates (Sengeløv et al.,
2003; ˇ
Ziˇ
zek et al., 2011). The adsorption of antibiotics depends on the
type of antibiotics and soil properties (pH, clay content, organic matter
and cation exchange (Wang and Wang, 2015)). The widespread use of
antibiotics increases antimicrobial-resistant genes (You and Silbergeld,
2014), which may disseminate in soil bacteria through horizontal gene
transfer (Xie et al., 2017), jeopardizing ecosystem health (Zubair et al.,
2020).
1.2. The relevance of soil biodiversity and the impact of intensive farming
systems
Soil organisms and their diversity are critical drivers for ecosystem
processes (Smith et al., 2015). Their functional diversity drives the
multifunctionality of soil systems (Konopka, 2009; Wagg et al., 2014;
Wardle et al., 2000). Within the soil food web, microorganisms and soil
Fig. 1. Annual manure production (million tonnes) in the European Union and UK and distribution according to main animal types (Period: 2016–2019).
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
4
fauna (micro-, meso- and macrofauna) have a dominant role in shred-
ding, transforming and decomposing soil organic matter (Brussaard,
2012; de Vries et al., 2013; Frouz, 2018). As such, soil organisms drive
nutrient cycling (Osler and Sommerkorn, 2007), including N xation
(Singh et al., 2016), plant nutrient and water uptake (Geisseler et al.,
2010), plant productivity (Van Der Heijden et al., 2008) and pathogen
and antibiotic resistance (Chen et al., 2017; Podmirseg et al., 2019).
Intensive agricultural systems have been shown to cause profound
alterations in soil biota abundance and activities (Orgiazzi et al., 2016a;
Tsiafouli et al., 2015). In the EU, most agricultural soils suffer from soil
biodiversity losses due to soil erosion, climate change, agricultural
intensication and diffuse soil contamination (Jeffery and Gardi, 2010;
Orgiazzi et al., 2016b). Recent studies have shown that the application
of untreated slurry can harm soil biota by introducing antibiotic resis-
tance genes (Chen et al., 2017; P´
erez-Valera et al., 2019) and pathogens
(Goberna et al., 2011) or broad-spectrum antiparasitic controls, thus
reducing the diversity of dung fauna (Adler et al., 2016; Iglesias et al.,
2006). Also, dietary supplements such as heavy metals (Qi et al., 2018)
or hormones (and their hormone-like by-products such as bile-acids)
have been shown to negatively impact soil organisms (Mendelski
et al., 2019). Ammonium-based mineral fertilisers and urea could
accelerate soil acidication (Goulding, 2016) and in turn, the mobility of
metals (Czarnecki and Düring, 2015), and be responsible for the abrupt
declines in microfauna biomass and diversity (Chen et al., 2019). In
relation to this, Rousk et al. (2010) also found bacterial communities to
be strongly impacted by decreases in soil pH. Mineral fertilisers also
signicantly alter the structural and functional community composition
of microbes (Manoharan et al., 2017; Prashar and Shah, 2016; Wang
et al., 2016). For example, they stimulate ammonia-oxidizing bacteria,
which are predominantly responsible for N
2
O emissions (Meinhardt
et al., 2018), increase the abundance of aphid pests (Gagic et al., 2017)
or decrease the activities of arbuscular mycorrhizas (de Souza and
Freitas, 2018) and gut microbiota in collembolans (Ding et al., 2019b).
Negative effects on soil biota increase when mineral fertiliser are applied
over a long period (29 years), compared to the application of manure
treatments consisting of composted farmyard manure and biodynamic
preparations (Faust et al., 2017; Sradnick et al., 2013) and when they
replace carbon-rich organic fertilisers, as was found for nematode
community structure and functions (Liu et al., 2016). Declined soil
biodiversity reduces long term nutrient availability to plants (Wagg
et al., 2014); in fact, intensive farming systems often have low nutrient
retention (Lago et al., 2019; Schrama et al., 2018). Mineral fertilisers do
not add organic carbon to soils. Therefore, the decline of soil organic
matter and soil biodiversity may cause a positive (self-accelerating)
feedback loop with detrimental environmental consequences (Fig. 2).
Table 1
Research protocol leading into research questions and keywords. The Research
Questions in the rst column will be addressed in the corresponding sections.
Research Question
(RQ)
Aim Method Keywords in the
Search
RQ1: Which factors
regulate the direct
and indirect
effects of farm
manures on soil
biodiversity and
their implications
on the fate of
manure
additions?
(Section 3.1)
To identify the
factors
determining the
impact of
manure on soil
biodiversity
including
benets and
threats to soil
biodiversity as
well as the effects
of soil
biodiversity on
the fate of
manure
To perform a
systematic
literature review
on the effects of
manure on soil
biodiversity and
vice versa
Manure
management
and/or animal
faeces and/or
animal dung and/
or animal urine,
benet and/or
harm, soil
biodiversity
RQ2: Which
practices help to
achieve
sustainable
manure
management in
the EU? (Section
3.2)
Recommend best
practices for
integrating soil
biodiversity in
manure
management to
enhance benets
of manure for
soil biodiversity
To examine
sustainable
farming practices
for the role and
integration of
manure
Manure
management,
and/or
sustainable
agriculture, and
European Union
RQ3: What role and
importance, if
any, is attributed
to soil
biodiversity in
current European
legislation on
manure
management?
(Section 3.3)
To investigate
the extent to
which European
policy
instruments
integrate soil
biodiversity and
manure
management
To examine the
integration of
manure
management and
soil biodiversity
in legal
frameworks
Manure
management,
and/or soil
biodiversity, and/
or policy
instruments, and/
or European
Union
RQ4: Which
shortcomings in
regulations and
practices, if any,
currently prevent
sustainable
manure
management in
the EU? (Section
3.4)
To determine
knowledge gaps
and limitations
in current
manure
management
practices and
regulations to
recommend
sustainable
manure
management in
the EU
To evaluate and
combine/match
the ndings
derived from the
two previous
methods
Manure
management,
shortcomings
and/or
limitations, and/
or sustainable
agriculture, and
European Union
Fig. 2. The links between organic matter, nitrogen surpluses and soil biota - impact & consequences (simplication; dependent on various environmental char-
acteristics such as soil properties, climate, type of organic fertiliser and land-use practices).
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
5
Soil organic matter availability and soil biodiversity are closely
interlinked (Thiele-Bruhn et al., 2012): Manure is a source of organic
carbon feeding soil organisms, whereby ecosystem services provided by
microorganisms increase organic carbon pool in the long term (Cui et al.,
2018; Gong et al., 2009; Zhang et al., 2014b) and therefore, can
contribute to more sustainable agriculture on the condition that certain
factors are fullled, which this paper aims to review.
1.3. Manure and its management under current EU legislation
Solutions for manure management are among the priorities of the
European Union (European Commission, 2020; European Commission,
2014). Currently, manure is not regulated under one single EU regula-
tion or directive. Instead, various different directives, regulations and
national Member State laws impact its management and quality. For
example, the Nitrates Directive (Council Directive 91/676/EEC) re-
quires Member States (MS) to implement national legislation governing
the timing and amount of manure in nutrient vulnerable zones (Oenema
et al., 2011), the Fertilising Products Regulation 2019/1009 indirectly
lays down the treatment of manure impacting the quality of organic
fertilisers and the Industrial Emission Directive requires MS to choose
from a set of best available techniques for better handling pig and
poultry manure (2010/75/EU). Although nutrient budgets (Klages et al.,
2020) and other good agricultural practices set by MS (such as allowed
application amount in nutrient vulnerable zones, storage practices) have
improved the use efciency of organic fertilisers, including manure from
farmed animals (Dalgaard et al., 2014; Mihailescu et al., 2014; van
Grinsven et al., 2015), the nitrogen surpluses did not decrease in several
MS between 2010 and 2015 (European Environmental Agency, 2019).
In 2013, 26.6% of the monitored sites in EU had increased nitrate values
(European Commission, 2013a). The European Commission called for
reducing nutrient losses by 50% in its Green Deal (in the Farm to Fork
Strategy and the Biodiversity Strategy 2030) (European Commission,
2020d); therefore, more studies are investigating the recycling and
nutrient recovery potentials of manure (Huygens et al., 2020; Huygens
et al., 2019). Review work (Bloem et al., 2017), reports (Saveyn and
Eder, 2014) and legislation (Fertilising Products Regulation EC 2019/
1009) have focused on enhancing the effectiveness and safety of fertil-
ising products, only including processed manure. Effects on soil biodi-
versity were not considered – mostly since the scope of legislation in
place does not integrate soil biodiversity. While Blaustein et al. (2015)
explored the release and removal of pathogenic microorganisms
deposited with manure, excluding benecial microorganisms and effects
for soil fauna, the effects of manure on soil biodiversity and vice versa
(Chen et al., 2020) have not been considered adequately in literature.
Also, those interactions have not been put into a policy context.
Recently, several publications looked at EU policy developments rele-
vant to sustainable soil management (Montanarella and Panagos, 2021;
R¨
ombke et al., 2016). Sustainable manure management (Hou, 2016;
Malomo et al., 2018) and practices for nutrient recovery from manure at
the EU level (Huygens et al., 2020) gained more importance. Nonethe-
less, scientic papers on sustainable manure management (Hou et al.,
2017; Malomo et al., 2018) and reports produced by the EU and Euro-
pean advisory boards on soil biodiversity neglect the risks and benets
of manure for soil biodiversity. They do not consider the importance of
autochthonous soil biota that are added via manure (EASAC, 2018;
Huygens et al., 2019, 2020; Orgiazzi et al., 2016a).
This review analyses the current directives and regulations in place
for manure management in the European Union, assesses its impact on
soil biodiversity and explores how the incorporation of soil biodiversity
could enhance the solutions currently in place. The relationship between
manure management and soil biodiversity will be explored, both in the
published literature and European legislation, as a means of synthesizing
the main ndings and identifying critical limitations and recommen-
dations. For the latter, we will investigate manure’s potential to be a part
of sustainable soil management enhancing soil biodiversity. Potentially,
this includes guidelines to incorporate better manure management in
future regulatory policies. In particular, the scope of the review is to
focus only on livestock manure and does not include other organic
wastes applied to agricultural lands even in smaller quantities. This re-
view proposes creative solutions that integrate soil biodiversity pro-
motion by addressing manure management at a European scale.
2. Materials and methods
This review follows the method proposed by Torraco (2005) to
analyse existing literature and legislative documents. Therefore, we
divided the literature review process into three phases: planning
(theoretical background), execution (search strategy) and result anal-
ysis. The rst phase denes research questions and a transparent search
strategy to select relevant studies. Thereafter, the relevant information is
extracted by performing systematic research, which is then synthesized
and discussed to develop policy recommendations.
2.1. Research questions
As farm manure additions can have both benecial and harmful ef-
fects on soil biodiversity, it is necessary to study all environmental im-
pacts of animal manures. Accordingly, this study includes both direct
and indirect effects of manure on soil biodiversity, but also their feed-
back responses on the fate of manure. Based on this assumption, the
research questions and keywords dened in this review work are listed
in Table 1.
2.2. Data collection and search strategy
The search was done in well-known literature databases such as the
Web of Science (WOS) and Scopus. Before 1990, only few articles
addressed the effects of manure on soil biodiversity (the rst article on
this topic and recorded in WOS was published in 1991, whereas in
Scopus the rst published article appeared in 1997), therefore our
literature search included the peer-reviewed literature published be-
tween 1990 and February 2021. We also selected peer-reviewed publi-
cations in JSTOR, Routledge (Taylor & Francis), Google Scholar and
Research Gate. A Boolean search by keywords applied a topic search in
WOS and a TITLE-ABS-KEY search in Scopus (see Keywords in Table 1).
Additionally, relevant governmental documents and grey literature from
organisational websites from the European Union were systematically
screened, including reports and statistics published by the European
Commission or by agencies of the European Commission.
Relevant articles were selected based on the following inclusion
criteria: (i) articles referring to manure management in the European
Union and published in peer-reviewed scientic journals between 1990
and 2021 as described above and (ii) grey literature by the European
Commission or by agencies linked to the European Commission or on
behalf of the European Union (e.g., consultancies, assessments); (iii)
legislative documents relevant to manure. The nal database included
407 articles (for the full list see supplementary Table 1). The majority of
the inputs in the database are scientic articles (87%) followed by grey
literature (8%) and legislative documents (5%) (Fig. 3A). There is a huge
increase in documents (publications, grey literature, legislation) with
reference to manure during the last decade as ¾ of the documents were
published in the period 2011–2020 (Fig. 3B).
3. Results
3.1. Factors determining the impact of manure additions to soil
biodiversity in agroecosystems
The main factors determining manure’s impact on soil biodiversity
are manure quality and its management (Blaustein et al., 2015; Büne-
mann et al., 2006). Both can either threaten or enhance soil biodiversity
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
6
(Fig. 1). Variables impacting soil biodiversity are intertwined, therefore,
focusing only on one parameter is not sufcient to predict effects.
Various factors such as the dose, the composition of manure and the type
of soil must be considered (Serani et al., 2020). We divided variables
depending on two management phases:
1) Management I considers variables managed in the pre-spreading
phase (manure type, storage and treatment)
2) Management II considers variables managed in the spreading phase,
including the application methods, quantity and environmental
interactions.
3.1.1. Management I: pre-spreading phase
Decisions during the pre-spreading phase of manure include animal
fodder, medical treatment given to animals, and storage and treatment
of manure, considerably affect the quality of manure (Lekasi et al., 2003;
Ndambi et al., 2019; Tittonell et al., 2010).
Fig. 3. A. Proportion of the inputs (Scientic Articles, Grey literature, Legislative documents) in our Database in the last 30 full years (1990–2021).
B. The trend of inputs in our Database (Scientic Articles, Grey literature, Legislative documents) in the last 30 full years (1990–2020).
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
7
(i) Manure Type: The animal species from which the manure orig-
inates determine the diversity of microorganisms in manure. For
example, bacteria isolated from ruminants are very distinct from swine
manure (Whitehead and Cotta, 2004). Kim et al. (2014) found cattle
manure to contain higher diversity of methanogens compared to pig
manure. Loh et al. (2005) compared the effects of cattle and goat manure
on the abundance and growth rate of earthworms in vermicompost,
nding cattle manure to be more nutritious for earthworms. This indi-
cated enhanced mineralisation, whereby the nitrogen content increased
during the vermicomposting (Loh et al., 2005). A 13-year study found
apart from pasture management, both poultry litter and cattle manure to
increase soil microbial biodiversity (Yang et al., 2019a). Ansari and
Mahmood (2017) found both goat manure and poultry manure with
inoculated rhizobium to increase microbial population and enzyme.
Murchie et al. (2015) found cattle slurry benecial for earthworm
populations, whereas pig slurry had no effect. However, Sun et al.
(2016) found pig manure to diversify the fungal community more than
cow manure when applied together with mineral fertilisers.
The content of organic carbon in manure varies by the animal species
from which it originates. Moral et al. (2005) found horse, pig and rabbit
manures to have the highest total organic carbon content, which was
mostly determined by the organic matter concentration (manure con-
sisted of faeces and straw from the bedding). The fraction of easily
biodegradable organic compounds, which is also highly water-soluble,
was highest in goat, chicken and horse manure. Amendments to
manure enhance its organic matter content, which can either be added
manually (Michel Jr et al., 2004) or by incorporating the bedding ma-
terial such as straw, woodchips and sawdust (Petric et al., 2009).
Replacing proteins with wheat in pig diets had no impact on tested
Collembola and earthworm species (Serani et al., 2020). Both cattle
manure and swine manures add high amounts of organic carbon, but
cattle manure contained higher amounts of dissolved organic carbon
(Kim et al., 2014).
Feed supplements may be responsible for the presence of toxic
contaminants in manure (Bolan et al., 2004) and heavy metals (Guan
et al., 2011; Leclerc and Laurent, 2017), such as zinc and copper added
in intensive piglets and poultry farming to promote their growth and
prevent diarrhea (Gans et al., 2005; Jensen et al., 2016). A high con-
centration of copper and zinc can lead to their soil accumulation,
particularly when the high density of pigs is combined with clay and
alkaline soils (higher pH; Panagos et al., 2016). Once introduced to soils,
heavy metals can alter the function of soil organisms, e.g., shifting from
metal-sensitive to metal-resistant individuals (Giller et al., 1998).
Depending on the initial state of soil biodiversity, metal stress may in-
crease or decrease microbial diversity (Giller et al., 2009). However,
Gans et al. (2005) found that already small heavy metal concentrations
signicantly harmed microbial diversity. Zinc, copper, arsenic, manga-
nese and cadmium signicantly inuence the composition of microbiota
(Li et al., 2020). An increasing concentration of copper sharply decreases
bacterial richness and evenness (Nunes et al., 2016). Also, Xu et al.
(2019) found metals to signicantly inhibit the microbial activity but
also the production of microbial biomass carbon. In the same study, the
fungal population reacted more sensitively to metal toxicity Xu et al.
(2019). In high amounts of pig manure amended to soils, the zinc and
copper concentration reduced earthworms (Segat et al., 2019), possibly
due to affected reproduction rates (Segat et al., 2015). Earthworms are
the soil invertebrates most harmed by an elevated copper concentration
(Naveed et al., 2014) and heavy metals were found to accumulate in
earthworms after the application of pig manure, depending on the
bioavailability of metals in pig manure (Li et al., 2010). Earthworms
avoided soils spiked with copper and zinc (Lukkari et al., 2005) and
concentrations of 272 and 156 mg kg
−1
for Cu and Zn, respectively
caused a 100% mortality. Domínguez-Crespo et al. (2012) found
323.11 mg kg
−1
Cu and 84.12 mg kg
−1
Zn to cause the death of 100% of
earthworms. Apart from metals, also hormones, fed to cattle to enhance
growth and milk production (Bartelt-Hunt et al., 2012), may end up in
manure (Kjær et al., 2007), posing a potential risk for invertebrates
(Mendelski et al., 2019).
Fig. 4. Benecial and harmful manure practices for soil biodiversity (on the left) result in sustainable manure management if the checklist provided on the right
is followed.
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
8
Veterinary products fed to animals are often poorly metabolized, e.
g., between 30 and 90% of antibiotics are excreted in urine and faeces
(Alcock et al., 1999). When introduced to soils, antibiotics signicantly
alter soil microbial community biomass (Eguchi et al., 2012; Hammes-
fahr et al., 2011) and composition (Ding and He, 2010; Hammer et al.,
2016). For example, the antibiotic sulfamethoxazole resulted in an
initial decrease in bacterial growth rates, selecting more tolerant species
(Demoling et al., 2009). Combining copper and the antibiotic oxytet-
racycline signicantly declined functional microbial biodiversity (Kong
et al., 2006). Also, antibiotics reduced earthworm gut-associated mi-
crobial diversity (Ma et al., 2019) and antibiotic resistance genes
accumulated due to horizontal gene transfers (Ding et al., 2019a). Toxic
effects of pig manure were signicantly reduced in soils with a higher
organic matter content (Segat et al., 2015).
(ii) Storage: Post defecation, manure is exposed to environmental
conditions such as temperature, moisture, exposure to air and ultraviolet
radiation, starting a natural biodegradation process altering bacterial
community composition (Oliver et al., 2010) and organic carbon content
(Moral et al., 2005). The storage of manure by composting or ageing
impacts emissions and nutrients and its effect on soil biodiversity. With
an increasing storage time, pathogens in poultry manure decrease,
whereby reduced water content decreases microbial mobility (Brooks
et al., 2009). Without manure storage, the antibiotic monensin levels
exceeded thresholds for triggering effects on soil organisms (ˇ
Ziˇ
zek et al.,
2014). The study also found composting to accelerate monensin
degradation compared to manure ageing. However, storing manure
containing the antibiotic sulfonamide for 60 days did not prevent
sulfonamide-resistance genes (Heuer et al., 2008). After 175 days of
storage, the abundance of resistance genes decreased and thus was faster
declining after application to soils. Hammesfahr et al. (2011) found
sulfonamide antibiotic compounds in pig manure to signicantly inu-
ence the N cycle due to alteration in soil microbial community structure.
Storing contaminated manure in dark at 20 ◦C for 6 months increased
the concentration of sulfonamide antibiotic compounds due to their
adsorption to manure compounds. When added to soils, antibiotic
compounds are remobilized upon microbial degradation of manure. In
the same study, nitrifying bacteria were inhibited in stored manure,
possibly due to antibiotic resistance formed during storage. Applied to
soils, stored contaminated manure reduced the bacterial biomass and
inhibited nitrifying bacteria affected by antibiotics, while fungal
biomass in manure increased during the storage (Hammesfahr et al.,
2011).
(iii) Treatment: Chemical, thermal or physical treatments are used
to reduce contamination risks of fresh manures (Kumar et al., 2013).
Those treatments change manure composition and, therefore, nutrient
content, emissions and leaching potentials (Sommer and Hutchings,
2001) and soil organisms (García-Gonz´
alez et al., 2016). Based on the
literature reviewed, we list the impact of different manure treatments on
Table 3
The impact of soil biodiversity on the environmental threats caused by manure application: ++ large positive impact; +positive impact; −negative impact; −- large
negative impact; 0 neutral (neither positive nor negative impact); + − No clear position in literature; NA refers to not available studies. Techniques separating manure
into solid and liquid fraction allowing their separate management is not covered in the table since the impact on the environment and biodiversity depends on the
fraction and the technique.
Environmental Threats (relevant to soils)
Impact of biodiversity Emissions NH
3
Leaching Heavy Metal soil contamination Pathogens (Salmonella) Antibiotic resistance genes Carbon losses
Microbial biomass +− ++ ++ ++ ++ +−
Genetic diversity + + + ++ ++ ++
Soil fauna + + + ++ ++ ++
Table 2
The impact of manure treatment on the environment (with focus on soils) and biodiversity (for low to medium amounts of manure, not exceeding 25 t ha
−1
): ++ large
positive impact; þpositive impact; −negative impact; −- large negative impact; 0 neutral (neither positive nor negative impact); + − No clear position in literature;
NA refers to no available studies. Techniques separating manure into solid and liquid fraction allowing their separate management are not covered in the table since the
impact on the environment and biodiversity depends on the fraction and the technique.
Impact Environmental impacts with focus on soils (see supplementary text in Appendix A
for more details)
Soil Biodiversity
Manure Treatments
NH
3
Loss
Heavy
metal soil
pollution
Salinisation Antibiotics Pathogens Soil
organic
Carbon
content
Microbial
biomass
Genetic
diversity
Soil
fauna
Plant-
parasitic
nematodes
Raw application (from animals
farmed in stables, excluding
untreated manure by grazing
animals)
– ¡ ¡ ¡ ¡ þ þ¡ þ¡ ¡ þ
Aerobic composting of the solid
fractions (Aerobic microorganisms
decompose organic matter,
occurring naturally when manure is
stored in heaps)
¡ þ þþ þ þ þþ þ þþ þþ þ¡
Biostimulant Fermentation
(Naturally-occurring acidication
e.g., compost teas)
NA ¡ þ þ þþ þþ þþ þþ þþ þþ
Anaerobic digestion (Microbial
degradation of organic matter to
biogas, as methane and carbon
dioxide)
¡ ¡ þþ ¡ ¡ þ 0 þ¡ ¡ þ
Additives and other pre/
treatments (e.g., acidication
through the addition of chemical
compounds such as sulfuric acid)
þNA NA NA ¡NA NA NA 0 þþ
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
9
Table 4
Manure management in EU legislation, including Directives and Regulations.
Legislation Scope of the legislation in terms of manure management
Animal By-Products and Derived Products Regulation 1069/
2009
Allows the direct soil application of unprocessed manure (that is not placed on the market) to soil, notwithstanding the
quality of manure (type, storage, treatment). Manure application must follow the precautionary principle, considering
the particular environmental condition. The regulation also sets rules for the disposal, collection and transport of
manure, allowing the transfer of manure from one farm to another within the same MS without a health certicate
tracking the manure’s origin or quality. Sterilisation and production in a certied plant are needed to put fertilisers
based on manure on the market.
The Common Agricultural Policy (CAP) Regulation (EU) No
1306/2013
The majority of direct payments of the CAP encourages quantity over quality, favouring vast amounts of manure
which on the other hand might lead to “inappropriate disposal practices”, including the application of raw manure or
its application not being well-timed (Buckwell and Nadeu, 2016; Lazcano et al., 2008). Also, the CAP offers Member
States the option of a voluntary coupled support for animal-based food (dairy products, beef, veal, sheepmeal and goat
meat (European Commission, 2021a) aiming to prevent the import of animal products that were produced under sub-
optimal social and environmental conditions. The rural development fund of Pillar II allows MS to spend up to 30% of
their budget in activities, such as the support of manure processing or storage, including anaerobic digesters (
European Commission, 2017). Also, subsidies for organic farming are part of Pillar II (see below). With the CAP
2021–2027, MS can decide upon eco-schemes. The list of recommended eco-schemes includes “other practices related
to GHG emissions” including improved manure management and storage (European Commission, 2021b). Within the
conditionality of pillar I, the Good Agricultural Environmental Conditions (GAEC), introduces GAEC 5, the Farm
Sustainability Tool (FaST) (European Commission, 2018). Nutrient management requires data “such as data on soil,
the proximity of protected areas and legal limits on the use of nutrients” (European Commission, 2019). FaST takes
into account soil information obtained using on-site tests (soil and manure samples) and an inventory of all nutrient
sources (e.g., crop residues), application techniques and handling and storage of manure (PwC, 2019). When MS
decide for GAEC 5, the nutrient management with the FaST tool is compulsory for any agricultural land, not only for
Nitrates Vulnerable Zones (European Commission, 2018), aiming to make nutrient management more adapted to local
site conditions (EASAC, 2018) and to reduce “nutrient leakage in groundwater and rivers, as well as positively
contributing to soil quality and reducing greenhouse gas emissions” (European Commission, 2019).
Air Quality and National Emission Ceilings Directive 2016/
2284
MS may establish good agricultural practices by including manure to improve the nutrient status and soil structure,
addressing animal husbandry practice, livestock housing, manure storage and spreading techniques, and how to
improve nutrient management, for example, by promoting (a) investments for low emission animal housing systems,
(b) investments for low emission manure storage systems, (c) feed additives to reduce methane emissions, (d)
investment for low-emission manure spreading techniques and (e) investment for on-farm bio-digesters.
EU Nitrates Directive 91/676/EEC The most critical EU regulation for managing manure and its impact on the environment (Van Grinsven et al., 2012).
The Directive is Statutory Management Requirement 1 of the conditionality of the CAP – a prerequisite for direct
payments that are either area- or animal-based (EU Commission). The Directive aims to protect water quality by
designating nutrient vulnerable zones, requiring MS to propose Good Agricultural Practices in high-risk zones to
prevent eutrophication measures, i.e., by considering amounts, storage, time of the year and local weather for manure
application (Musacchio et al., 2020; Serebrennikov et al., 2020). The limits of applicable N in high-risk zones is max.
170 kg/ha of N and in certain zones of France, manure needs to be treated before its application to soils (Loyon, 2017).
Some MS provide farmers with exemptions if manure disposal does not harm ecosystems (Serebrennikov et al., 2020;
Svanb¨
ack et al., 2019). Time limits for manure application per hectare according to climatic and geological
characteristics prohibits the application of manures in winter or on steep slopes (Buckwell and Nadeu, 2016;
Louwagie et al., 2011). The implementation of the Directive varies greatly, for example, whether digestates from
anaerobic digestions are included and thus regulated (European Commission, 2013).
Organic Production Schemes Council Regulation (EEC) No
2092/91
Prohibits using manure obtained from factory-farmed animals in organic agriculture, but the denition is left to MS or
organic control agencies (Løes et al., 2017). The limited availability of certain essential nutrients such as P allows
organic farmers to apply mineral P, manure from conventional agricultural systems, organic residues if composted or
anaerobically digested and other animal residues, including bones and meat (Løes et al., 2017). Varying regulations
and limits of permitted conventional manure complicate the comparisons within MS, making consumer education
difcult. Antibiotics in Organic Farming can only be used if alternative treatments fail, they must be approved by a
veterinarian, and no animal can receive more than three treatments per year (Duval et al., 2020), except for
vaccinations, treatments against parasites and compulsory eradication schemes. If the animal’s life cycle is less than
one year, only one treatment is possible. Stricter regulations decreased the use of antibiotics signicantly (Wagenaar
et al., 2011). While the usage of copper-based fungicides was reapproved in 2018, the usage is limited to 4 kg year
−1
and incldudes copper on the list of candidates for substitution, which means once an alternative is found, copper will
be banned (Droz et al., 2021; EU 2018/1981). The amount of permitted copper was reduced by a factor of ten
compared to the 1980–1990 period (Karimi et al., 2020).
Fertilising Products Regulation (EC) No 2003/2003 replaced
by Regulation (EU) 2019/1009
It lays down rules for EU fertilising products to access markets. Since Regulation 2003/2003 only covers inorganic
fertilisers, standards such as composting, standards and quality widely differ between MS (Bernal et al., 2017; Cesaro
et al., 2015). This regulation will be replaced by the Fertilising Products Regulation 2019/1009 in July 2022 (
European Commission, 2019), which will impose stricter limits for mineral fertilisers (for example cadmium content)
and will allow fertilising products from recycled bio-wastes and other secondary raw materials (Huygens et al., 2019;
Paungfoo-Lonhienne et al., 2019), aiming to reduce the pressure on biological resources through the recycling of
biowastes (Bell et al., 2018; Paungfoo-Lonhienne et al., 2019). Waste Products, according to 2008/98/EG (including
manure), will lose their status as waste as soon as they are transformed into other products (Directive 2008/98/EC).
For example, its treatment (such as fermenting or composting) will convert manure into a fertilising product. Labelled
as a Circular Economy product, the Regulation allows for the recycling of manure as a secondary waste product. This
will increase the access to fertilising products based on manure. The update will introduce limits for pollutants (such as
copper and zinc) in composts and digestates (the proposal does not regulate raw manure). For the anaerobic digestion
of manure, temperature-time proles for thermophilic digestion, glass/metal and plastic contents are regulated and
limit heavy metals, polycyclic aromatic hydrocarbons and pathogens. According to the regulation, any given organic
soil improver must have a dry matter content higher than 20% and the content of organic carbon must exceed 7.5% of
the total dry mass (Rodrigues et al., 2020).
Veterinary Medicinal Directive 2001/82/EC replaced by
Regulation EU/2019/6
Foresees the authorisation, testing, documentation and traceability of veterinary medicinal products, including, e.g.,
broad-spectrum anti-parasitics such as Ivermectin (Adler et al., 2016) but also hormones and antibiotics. The
regulation EU/2019/6 will replace the Directive 2001/82/EC in January 2022 (European Medicine Agency, 2019).
(continued on next page)
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
10
the environment, focusing on soils and soil biodiversity in Table 2 (for
more details on effects of different manure treatments on the environ-
ment see supplementary Table 2).
While heavy metals cannot be removed by treatments, antibiotics
(Zhang et al., 2014a), antibiotic-resistant genes (Sengeløv et al., 2003)
and hormones (Kjær et al., 2007) can be reduced via certain treatment.
Contaminants in slurry were found toxic to earthworms and enchy-
traeids (Segat et al., 2019; Segat et al., 2020) and collembolans in sub-
tropical soils (Segat et al., 2020). Walsh et al. (2012) found slurry to
reduce microbial growth and fungal diversity. High amounts of
ammonia had negative effects on soil organisms as found for cattle slurry
on collembolans (Pommeresche et al., 2017), earthworms (Van Vliet and
De Goede, 2006) and arbuscular mycorrhizal fungi (Kahiluoto et al.,
2000; Wentzel and Joergensen, 2016). Compared to solid manure, the
distribution of bacterial species in slurry was more evenly dominated by
certain species when urine was mixed with faeces (Cotta et al., 2003).
Several studies found raw liquid manure benecial for soil organisms
(Bosch-Serra et al., 2020; Plaza et al., 2004; Ponge et al., 2013; Wentzel
and Joergensen, 2016), e.g., by increasing bacterial biodiversity (van
der Bom et al., 2018). The application of raw liquid manure can lead to
toxic conditions for plant-parasitic nematodes due to high amounts of
organic acids, high ammonia compounds and low C/N ratios (Thoden
et al., 2011). Grifths et al. (1998) found cattle slurry to shift nematode
diversity from plant-parasitic to bacterivorous species. Neufeld et al.
(2017) found higher microbial biomass and a higher activity of
cellulose-degrading enzymes after applying solid manure compared to
the application of liquid dairy slurry. Increasing the dose of raw and
composted pig manure (from 0 to 100 t ha
−1
) enhances the toxicity of
raw manure on collembolans (Maccari et al., 2020). Earthworms were
more abundant after applying composted manure compared to un-
treated solid manure (Hurisso et al., 2011).
Composted amendments provide organic matter for soil organisms,
introduce benecial microbes (Semple et al., 2001) and reduce amounts
of antibiotic resistant genes (Tien et al., 2017). Earthworms are more
abundant if composted manure is added rather than when slurry is
applied (Rollett et al., 2020). Farmyard manure increases bacterial
biodiversity, particularly those better adapted to nutrient-rich environ-
ments (Francioli et al., 2016). However, larger amounts of composted
manure did not increase the population, likely due to salinity stress
(Hurisso et al., 2011). When compost is fermented (immersion-extrac-
tion-oxygenation) in a liquid (mostly water) for a certain period (from
few hours up to two weeks), compost teas are produced (De Corato,
2020b), enhancing fertilising effects of manure (Pane et al., 2014).
Compost teas can be produced under aeration (e.g., by manually stir-
ring) or under anaerobic conditions (Scheuerell and Mahaffee, 2002)
and additives may be added such as minerals (e.g., rock dust) and
nutrient additives including kelp extract and/or humic acids (De Corato,
2020a; Pane et al., 2014). Depending on the conditions, the fermenta-
tion process enhances the content of matured, stabilised organic matter
(humic substances) and soluble mineral nutrients (Pant et al., 2012).
However, it also multiplies compost microbiota which effectively sup-
presses soil-borne plant diseases and pathogens (De Corato, 2020b). For
example, a compost tea based on fermentation of amino acid fertiliser
from rapeseed meal with compost from pig manure effectively
controlled Verticillium wilt affecting cotton (Lang et al., 2012) and
aerated compost tea with added nutrients (kelp extract, rock dust, and
humic acid) contributed to control grey mold in Geranium (Scheuerell
and Mahaffee, 2006). Pant et al. (2012) found the highest number of
active microbial populations in compost teas based on aged vermi-
compost. The survival rate of grown strains depends on the quality of the
inoculant (compost feedstocks, compost age, water ratio, fermentation
time, and added nutrients), the soil environment (pH) and climatic
factors (temperature). Consequently, the results of fermentations are
variable and unpredictable (Scheuerell and Mahaffee, 2002). However,
preparations based on local inputs that sustain indigenous soil biota are
better adapted to the local environmental conditions. They are more
likely to thrive under agricultural practices than the commercial inoc-
ulum sources frequently used in biostimulant products (Valliere et al.,
2020). Further research is needed to improve the application and us-
ability of compost teas (De Corato, 2020b). Due to its properties and
because compost teas enhance plant growth (De Corato, 2020a; Pane
et al., 2014), compost teas follow the denition of biostimulants sug-
gested by du Jardin (2015).
For biogas production, organic material in manure is anaerobically
digested by anaerobic microorganisms decomposing carbon in the bio-
reactor. Digestates are the “waste” product of the process, whereby ef-
fects to soil organisms differ. Compared to no fertiliser application and
undigested slurry, anaerobic digestates stimulate bacterial growth while
no effect on fungi was found (Walsh et al., 2012). In another study,
Barduca et al. (2020) found anaerobic digestates to increase fungal ac-
tivity while not changing microbial biomass. Also, no harmful effects
have been observed on soil organisms after nine days (Johansen et al.,
2013) or after one month of its application (Podmirseg et al., 2019).
However, Westphal et al. (2016) found only a minimal impact of
anaerobic digestion on nematode, bacterial and fungi communities in
soil one month after application of digestates. Another six-week exper-
iment comparing the fertilisation effect of anaerobic digestates of food
waste with mineral N-fertiliser, compost and farmed animal manures
showed that the high amount of NH
4+
resulting from anaerobic diges-
tion harmed earthworms (Rollett et al., 2020). Digestates are less
decomposable than undigested manure (Cavalli et al., 2017) since more
carbon sources and nutrients for soil organisms are limited after diges-
tion (M¨
oller, 2015). Other studies found negative effects on fungi
(estimated as reduced concentrations of ergosterol (Walsh et al., 2012;
Wentzel and Joergensen, 2016)) and a decreased carbon content
compared to conventional slurry, which may affect soil biota (Ernst
et al., 2008). Mukhuba et al. (2018) found anaerobic digestion to reduce
the diversity of culturable bacteria. Anaerobic digestion does not reduce
heavy metals, so that digestates often contain high amounts of heavy
metals (Nkoa, 2014), which may harm bacterial diversity (Gans et al.,
2005) and signicantly decrease soil biodiversity (Anjum et al., 2017).
Similarly, a study on Collembola found a signicant drop in the density
of collembolans three days after the application of anaerobically
digested cattle slurry, which may be caused by high NH
4+
contents.
Seven weeks from the application, numbers did only slightly recover
(Pommeresche et al., 2017). Compared to thermophilic digestates,
leading to more rapid digestion under higher temperatures (around
55 ◦C), digestates of the slower mesophilic digestation at around 37 ◦C
contained more microorganisms (pathogenic and benecial; Qi et al.,
2018).
Acidifying swine slurry increases the concentration of volatile fatty
acids at low pH, limiting microbial metabolism and, hence, pathogen
Table 4 (continued )
Legislation Scope of the legislation in terms of manure management
Among the main objectives of the updated regulation (EU/2019/6) antimicrobial resistance should be tackled through
prudent usage. As a measure, for example, certain antimicrobials may be restricted for use in animals, prioritising
them for human treatment. It requires MS to collect data on antimicrobial medicinal products used in animals.
Industrial Emissions Directive (2010/75/EU) The Directive regulates manure management for larger pig and poultry farms. To obtain a permit, the farms must
comply with a series of conditions related to animal housing and manure management (storage, processing
techniques, application techniques) to reduce overall adverse environmental impacts.
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
11
activity and effectively controlling plant-parasitic nematodes (Mahran
et al., 2009). A recent study found a little overall impact of acidied
slurry on collembolans, although a larger amount of slurry may create
favourable habitats for these organisms (D’Annibale et al., 2019).
3.1.2. Management II: spreading phase
During the spreading phase of manure, decisions regarding the
application techniques, the quantity of manure and environmental in-
teractions (e.g., physiochemical properties and climate) may also impact
soil biodiversity (Bünemann et al., 2006; Unc and Goss, 2004).
(i) Application: When manure is supercially applied on broad
surfaces, e.g., with broadcasting methods allowing uniform manure
application such as through box spreaders, tank wagons, tow hoses and
irrigation systems (Eurostat, 2021c), the microorganisms contained are
more likely exposed to UV light and desiccation, reducing their con-
centrations (Forslund et al., 2011). In the case of supercial manure
application, the likelihood of run-off is higher (Lamba et al., 2013; Sis-
tani et al., 2010), increasing the release and transport of pathogenic
microorganisms. Subsurface application such as injection prevents run-
off (Lamba et al., 2013) and increases the survival chances of microor-
ganisms and the risk of groundwater contamination, especially when
manure contains pathogens and is untreated (Forslund et al., 2011;
Semenov et al., 2009). De Goede et al. (2003) found that the surface
application of slurry enhanced soil respiration, while slit injection
negatively impacted epigeic earthworms closer to the soil surface.
However, the densities of deeper dwelling anecic and endogeic earth-
worms were equal or higher after slit injection compared to surface
application on 12 different farms. The same authors estimated that slit
injection damaged 15% of epigeic earthworms at depths 0–5 cm and
30% of epigeic earthworms at depths 0–10 cm. Van Vliet and De Goede
(2006), comparing broadcaster and slit injected manure, found no
impact on the abundance of Enchytraeidae and nematodes. However,
one week after broadcaster manure application, nematodes in the stage
of dauerlarvae disappeared. Slit cutting in the presence of manure had
no impact on earthworm abundance, but when combined with manure,
broadcasting increased the abundance of earthworms. However, in the
absence of manure, slit cutting declined the abundance of earthworms
by 30% one week after the application. Broadcasting affected particular
juveniles possibly due to salts present in slurry.
Effects of manure application on the abundance of earthworms
depend on soil moisture and season. Wet conditions negatively impacted
epigeic earthworms (De Goede et al., 2003; Van Vliet and De Goede,
2006), most likely since they are attracted to the soil surface and thus
easier damaged by the machines. Slurry applied together with chalk and
clay minerals to 18 different elds resulted in a higher abundance of
Collembola and Acari in broadcasted compared to slit injected elds.
Microarthropods responded in particular to the slurry application
method rather than slurry manure type (Jagers op Akkerhuis et al.,
2008). Slurry is frequently applied using tractors, which poses a high
risk of subsoil compaction (Thorsøe et al., 2019), negatively impacting
soil biodiversity, the biomass of soil organisms and enzyme activity
(Nawaz et al., 2013).
(ii) Quantity: The continuous application of manure increases mi-
crobial biomass, soil respiration, nutrients such as potassium, phos-
phorus, magnesium (Segat et al., 2019) and mineral and organic N pools
in soils (Pang and Letey, 2000; Yu et al., 2021). When plant nutrient
requirements are exceeded, manure can contribute to the soil accumu-
lation of soluble salts, heavy metals, P and N (Huygens et al., 2019).
Ammonium was found to be among the main toxicity factors in organic
materials, reducing the survival and reproduction of invertebrates
(Renaud et al., 2017) and collembolans (Maccari et al., 2020). When pig
slurry is applied to high N demanding crops following their maximum
demands (2 t dry matter ha
−1
), soil biota is likely harmed (Domene et al.,
2008).
High amounts of applied manure also increase the content of path-
ogenic bacteria concentrations (Hruby et al., 2016) and heavy metals
such as cadmium (Belon et al., 2012; Carne et al., 2020), zinc and copper
(Segat et al., 2019). In France, zinc in manure contributed to 78% of
overall zinc contamination in soils (Belon et al., 2012). Testing various
doses of pig manure on earthworms (ranging 0 to 300 t ha
−1
) and
enchytraeids (ranging 0 to 100 t ha
−1
), Segat et al. (2020) found
increasing doses to reduce juveniles of both groups signicantly.
Another study tested the effects of a single application of 20, 50 and
150 t ha
−1
of pig manure on soil fauna, whereby the highest amount
altered soil fauna composition signicantly, increasing insect larvae,
dipterans and mites while decreasing ants, collembolans, earthworms
and enchytraeids (Segat et al., 2019).
The doses of unstabilised swine manure had a more signicant
impact on earthworms compared to the effects of varying diets and the
use of antibiotic growth-promoting additives (Serani et al., 2020).
Also, for collembolans, toxicity increases with the quantity of pig
manure, whereby doses of 50, 75 and 100 t ha
−1
caused 100% mortality
(Maccari et al., 2020). However, Porhajaˇ
sov´
a et al. (2018) found 25 or
50 t ha
−1
to not impact the presence of beetles (Coleoptera). The ap-
plications of 600 kg N ha
−1
of pig manure led to a higher abundance of
nematodes, compared to 150 kg N ha
−1
, possibly due to an increase in
microbial biomass (Jiang et al., 2013).
(iii) Environmental interactions: Testing the application of swine
manure in four different soils, Segat et al. (2015) found that the soil
texture and organic matter alter the toxicity of swine manure for
earthworms. Compared to total N and total carbon, Jiang et al. (2013)
found pH to have the most considerable effect on nematodes and fun-
givores (the study does not specify considered species of fungivores).
Bacterivores, plant parasites and microbial functional groups were pri-
marily affected by total carbon. The pH and cation exchange capacity
are the most critical parameters affecting the bioavailability of metals
and thus their toxicity (Lock and Janssen, 2001). Extreme soil pH values,
both acidic and alkaline soil conditions, could lead to pollution. Toxicity
by high pH affected the reproduction of Collembola (Crouau et al., 1999)
and pH above 9 caused a 100% mortality of the earthworm Eisenia fetida
(Kaplan et al., 1980). However, the success rate of slurries against plant-
parasitic nematodes depends on non-ionized organic acids which are
predominantly found at low pH (Thoden et al., 2011), so that swine
manure was more successful in reducing plant-parasitic nematodes in
acidic soils compared to alkaline soils (Lazarovits et al., 2001). The
mobility of microorganisms in soils applied via manure signicantly
increases with rain or irrigation (Brooks et al., 2009). Also, slurry
applied to frozen (Rocard et al., 2018) or waterlogged soils (Weslien
et al., 1998) is more susceptible to run-offs, increasing the risk of Sal-
monella spreading (Hruby et al., 2018). Both climate and soil properties
impact the variation of soil organisms, for instance explaining respec-
tively 17.3% and 24.7% in nematode community composition (Jiang
et al., 2013).
3.1.3. Effects of soil biodiversity on manure management
Over the last years, edaphic conditions and environmental hetero-
geneity have been increasingly investigated for their role in determining
the effects of manure additions on pollution, nutrient availability, and
soil biodiversity. The abundance of native, resident soil microbes is
crucial for pathogen suppression (Goberna et al., 2011; P´
erez-Valera
et al., 2019; Podmirseg et al., 2019). Soils with high numbers of native
soil bacteria are more likely to outcompete allochthonous species
(P´
erez-Valera et al., 2019; van Elsas et al., 2012), preventing the
dissemination of antibiotic resistance genes (Chen et al., 2017). Also,
earthworms contribute to reduce the broad-spectrum antibiotic oxytet-
racycline in manure after its application to soil, possibly activating and
introducing microbes that can degrade antibiotics (Ravindran and
Mnkeni, 2017). Soil bacterial diversity reduces the colonisation ability
of the most common animal faecal pathogens that cause infections, such
as the toxin-producing E. coli bacteria (Jiang et al., 2002; Xing et al.,
2020). Pathogenic suppression is much lower in perturbed soils, where
their indigenous microora has been severely reduced (Moynihan et al.,
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
12
2013). Those effects were conrmed in sterilised soils (Xing et al., 2020).
In undisturbed agricultural systems (e.g., no-tillage), local bacteria in
the upper soil layer form a barrier against antibiotic resistance bacteria
(P´
erez-Valera et al., 2019), whereas tilled soils increase the risk of
signicantly decreasing earthworm populations, as reported by Briones
and Schmidt (2017).
Soil organisms play a crucial role in detoxifying soils, such as bac-
teria using metals in their metabolism for their growth (Ahemad, 2019);
the large biomass of fungi reducing them to fewer toxic forms through
chemical oxidation or reduction processes (Pinedo-Rivilla et al., 2009;
Siddiquee et al., 2015); or earthworms immobilising them by adsorbing
them to their cell walls (Ayangbenro and Babalola, 2017), which may be
boosted by vermicomposting (Swati and Hait, 2017). The diversity of
soil microorganisms are crucial to remove or detoxify contaminants
(Garbisu et al., 2017). When soils’ biodiversity was reduced through
fumigation practices, soils were less resilient to persistent stress by
copper contamination (Grifths et al., 2000). Increased soil fauna di-
versity enhances plant nutrient uptake and reduces NH
3
losses (Geisseler
et al., 2010; Singh et al., 2016; Van Der Heijden et al., 2008). Earth-
worms crucially improve soil structure and stabilise soil organic matter,
thus enhancing the retention of carbon and N (Jouquet et al., 2011; Lago
et al., 2019). Moreover, through their interactions with microorganisms,
earthworms reduce ammonia volatilisation and increase soil N and P
availability (Maldonado et al., 2019). This has led Lago et al. (2019) to
conclude that more diverse food webs are critical for N retention in
agricultural soils. Without this wide variety of biological activities,
fewer nutrients are retained within the soil matrix and end up leached
into groundwaters, as it has been found when soils are sterilised (Cars-
well et al., 2018) or when earthworms are excluded (Butenschoen et al.,
2009). Also, for the immobilization of heavy metals, indigenous mi-
croorganisms play a crucial role as potential bioremediators (Li et al.,
2020).
Soil biodiversity increases carbon stabilisation by stimulating the
activity of microbial communities (Dignac et al., 2017), which enhances
the mineralisation of organic matter in the short term and the seques-
tration of carbon in organo-mineral complexes in the long term (Brad-
ford et al., 2013; Six et al., 2006). Earthworms’ fast incorporation of
organic matter into cast aggregates prevents decomposition (Vidal et al.,
2016). Changes in community composition may have large impacts on
carbon dynamics (Nielsen et al., 2011), e.g., the losses of nematodes
declined carbon cycling signicantly (Barrett et al., 2008). The loss of
microbial diversity decreases carbon cycling, for instance, by reducing
decomposition and carbon respiration (De Graaff et al., 2015). In a
laboratory approach, a decreasing microbial diversity highly affected
carbon cycling, whereby the signicance of the diversity effect increased
with nutrient availability (Maron et al., 2018).
Finally, soil biology plays a crucial role in preventing environmental
threats related to manure application (see Table 3). To enhance the
safety of soil organic amendments, Renaud et al. (2017) proposed to
combine chemical and ecotoxicological data for an assessment of the
toxicity of amendments. The study recommends to consider varying
sensitivities and contaminant thresholds for various organisms but also
to consider diverse routes of exposure during the assessment of toxicity
of organic amendments, including manure.
3.2. Best practices for integrating soil biodiversity in managing manure
3.2.1. The role of biostimulants, biodynamic farming and agroecological
practices
Soil fauna and microbial metabolism can be signicantly accelerated
by using fewer pesticides and mineral fertilisers (Birkhofer et al., 2008;
Hartmann et al., 2015; Yang et al., 2019b), avoiding tillage (Briones and
Schmidt, 2017) or by providing high-quality sources of nutrients to soil
organisms (organic fertilisers/mulching; Carrillo et al., 2011). Also,
adding external organisms may enhance the local genetic diversity
(Brussaard et al., 2007; du Jardin, 2015; Sangiorgio et al., 2020). In
organic farming systems (Bengtsson et al., 2005; Bünemann et al.,
2006), manure plays a crucial role as a direct source of nutrients for soil
organisms. It also adds genetic biodiversity and benecial microorgan-
isms to soils (Marschner et al., 2003; Zhen et al., 2014). The biodiversity
in manure fertilised elds exceeds the number of mineral fertilised soils
by 50% (excluding non-predatory insects and pests) (Bengtsson et al.,
2005). The mean number of earthworm species was 126% higher and
the number of individuals 156% higher in a diverse organic system than
in a monocrop conventional eld, applying mineral fertilisers (Feledyn-
Szewczyk et al., 2019).
In organic viticulture, soil microorganisms were three to four-fold
higher than in conventional viticulture (Karimi et al., 2020) mostly
due to more available organic carbon applied by farmyard manure for
soil organisms (Okur et al., 2016) - despite the usage of copper-based
fungicides and bactericides, applied against downy-mildew. These
treatments are currently still allowed in organic farming due to the lack
of alternatives. While extremely high copper concentrations were found
to negatively affect soil biodiversity, allowed amounts (<4 kg per ha)
are considered to not harm soil biodiversity (Karimi et al., 2021).
However, other studies found that even low copper contamination can
alter the spatial distribution of soil bacteria and fungi (Mackie et al.,
2013) and cause severe damages on yeasts (Grangeteau et al., 2017).
More extensive comparative studies and research for alternatives are
needed (Karimi et al., 2020).
In ancient traditional knowledge systems in India (Biswas, 2020;
Randhawa and Kullar, 2011), Europe (Bogaard et al., 2013; Krausmann,
2004) and South America (Birk et al., 2011), manure has been used as a
main organic fertiliser or even as an ingredient for biostimulants.
Traditional methods emphasise (i) quality of manure (e.g., source,
including feed of the animal whose manure is used), (ii) method used for
its treatment, including proportions of various ingredients used for such
treatment and (iii) its management, i.e., the duration for which it can be
stored, the amount to be used, and time/frequency of application. Such
practices are applied, for example, in Natural Farming practices in India.
Natural Farming refers to agro-ecological farming practices based on
Indian traditional knowledge and combines organic farming practices
with farm-made biostiumulant preparations (Das et al., 2020).
Most formulations used in Natural Farming have (cow) manure
(particularly manure from indigenous breeds) as a critical ingredient.
These formulations transform manure via fermentation into a potent
biofertiliser that signicantly enhances soil biodiversity. Apart from cow
manure, farmer-made biostimulants are based on local, site-specic in-
puts such as sugar (e.g., ripe fruits), proteins (e.g., pea our) and local
soil. The application of these formulations improves the soil physical,
chemical, and biological properties (Das et al., 2020; Devarinti, 2016;
Liao et al., 2019), which are crucial to extract phosphorous from organic
wastes (Szogi et al., 2015). Natural Farming biofertilisers have been
shown to increase the availability of nutrients for sunowers while
decreasing the contaminants’ concentration, such as chloride and sulfate
(Iftikhar et al., 2019).
In the EU, organic production, that includes also the use of Biody-
namic preparations ((EC) No 834/2007), allows the use of manures and
composts complemented with compost teas based on manure (M¨
ader
et al., 2002). The effectiveness of biodynamic preparations developed
into a controversial scientic debate (Faust et al., 2017). Some studies
nd these preparations to enhance plants’ self-regulating processes
(Morau et al., 2020) and increase soil organisms’ activity (Giannattasio
et al., 2013; Spaccini et al., 2012). Zaller and K¨
opke (2004) found the
decomposition signicantly enhanced after 100 days in plots that
received biodynamic preparations based on manure mixed with various
owers, compared to other farmyard manure treatments. By contrast,
others nd no difference in the effects between composted cattle manure
and biodynamic preparations (Faust et al., 2017; Hendgen et al., 2018;
Krauss et al., 2020; Sradnick et al., 2013). However, these studies often
focus on microbial biomass (Faust et al., 2017) or considered the mi-
crobial functional diversity only (Hendgen et al., 2018; Sradnick et al.,
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
13
2013). M¨
ader et al. (2002) considered the overall soil biodiversity,
stating a higher diversity in biodynamic farming. Nevertheless, the trials
conducted on organic and biodynamic practices sourced manure from
different sources, limiting the signicance of the results (Faust et al.,
2017).
Both Natural Farming in India and some organic farming practices in
the EU combine the addition of microorganisms gained from manure
with more sustainable agricultural practices that link ecology with food
production, the so-called agroecological practices (Chavarria et al.,
2018; Migliorini and Wezel, 2017; Thiele-Bruhn et al., 2012). Agro-
ecological practices are recommended due to enhanced resilience to
ecologic pressure, such as pest attacks and weather extremes (Altieri
et al., 2015) and economic and social pressure on farmers (Van der Ploeg
et al., 2019). The same study provides empirical evidence for the eco-
nomic potential of agroecological practices in the EU by simultanously
sustaining employment levels and increasing incomes. Since agroeco-
logical farming systems operate on a regional basis, direct distribution,
marketing and fairer prices increase consumer trust (Dumont et al.,
2016) as well as (small) farmer prots. Parallel to agroecological
movements (bottom-up movements, e.g., in Belgium; Stassart et al.,
2018), there is increasing political support for these practices and
institutional dynamics are driving the process: France has integrated
agroecological practices at the national environmental policy level
(Ajates Gonzalez et al., 2018; Moudrý et al., 2018), and the EU mentions
explicitly agroecological farming practices in its Biodiversity Strategy
for 2030 as a possible solution to maintain productivity while increasing
soil fertility and biodiversity.
3.2.2. Recommendations for sustainable manure management
To prevent environmental threats relevant to soil biodiversity, we
synthesised the optimal sequence of management steps that should be
considered in manure management that consider the response on soil
biodiversity (see the left side in Fig. 4).
Firstly, manure quality must be assessed and only applied without
further treatment when:
(i) heavy metals that have been added to the animal feed do not
exceed the given tresholds,
(ii) no antibiotics have been used in the previous week(s),
(iii) the local climatic/edaphic conditions are favourable (soils are
not frozen or extremely wet after heavy rainfall events), and.
(iv) the quantities applied are not excessive.
However, if animals show any signs of illness or have been treated
with antibiotics or hormones, manure must be treated before being
applied to soils. Here it must be noted that also just storing manure
without further treatment does not reduce antibiotic compounds in
manure adequately (Hammesfahr et al., 2011). Depending on the type of
pollutant agent, different treatment strategies need to be adopted.
Manure as fertiliser should only be anaerobically digested if no an-
tibiotics were added to the diet of farmed animals, othewise antibiotic-
resistant genes may remain in the digestates (Tien et al., 2017). Keeping
ill animals separately from healthy animals prevents the mass-usage of
antibiotics, while permitting the separate storage of contaminated and
uncontaminated manure. This, in turn, permits the selective usage of
manure from healthy animals as an organic fertiliser. Also, to reduce the
toxic effects of pig manure in soils, organic matter should be added (e.g.,
straw or other plant residues; Segat et al., 2015).
Manures from animals treated with heavy metals (copper/zinc)
should not exceed the permited limits of copper in organic farming
(<4 kg per ha). Ding et al. (2021) determined dosages of copper and zinc
as thresholds that should not be exceeded: 5 mg/kg Cu and 50 mg/kg Zn
for animals in the range of 9–15 kg; 4 mg/kg Cu and 50 mg/kg Zn for
animals between 15 and 25 kg and 4 mg/kg Cu and 10 mg/kg Zn for
animals in the range of 25–60 kg. Since heavy metals cannot be removed
from manure by any currently known treatment methods, the amount of
copper needs to be calculated based on dietary additions and threshold
limits. Also, mineral zinc may be replaced by organic zinc – resulting in
improved faecal scores (Bouwhuis et al., 2017). The incidence of diar-
rhea of piglets can be reduced by decreasing stress among animals (e.g.,
by reducing the animal counts, more space and access to the outdoor;
Martínez-Mir´
o et al., 2016).
Careful attention should also be paid to the amounts of manure
added since several studies have highlighted the negative correlation
between manure quantities and soil biodiversity (Maccari et al., 2020;
Segat et al., 2019), which, in turn, is closely associated with the
increased concentration of pollutants in manure. Accordingly, when the
type and concentrations of these pollutants are unknown, applied
manure should not exceed 25 t ha
−1
(Segat et al., 2019). Applying less
amounts of manure might reduce other benets of this organic fertiliser
on soil quality, such as C sequestration. However, recent studies have
shown that the positive effects of manure application (in large quanti-
ties) on soil carbon storage are strongly dependent on agricultural
management (e.g., tillage, additional mineral fertilisation) and climatic
and soil abiotic condtions (e.g., soil pH and initial C content) (Gross and
Glaser, 2021). Indeed, climatic and edaphic conditions also need to be
considered before manure application. Particularly, clay-rich alkaline
soils are prone to metal accumulation, while acidic soils pose additional
environmental risks because enhanced metal mobility can lead to
groundwater pollution. Similarly, as there is a high amount of ammonia
in the digestates, manure should not be applied to alkaline soils without
adequate pre-treatment to prevent ammonium (NH
4
) emissions.
In the context of manure application methods, the separation of the
solid and liquid segments of manure through osmosis and scrabbing
permits better recovery of N from post-processing (Finzi et al., 2020).
Manure treatment practices must be included in subsidy systems to
ensure the economic feasibility of the practices. Manure should not be
applied by means of slit or broadcasting methods after raining events to
prevent harmful effects on earthworms (De Goede et al., 2003; Van Vliet
and De Goede, 2006).
In summary, to avoid ecotoxicological effects, thresholds must be set
that consider both the quantity of manure and the local climatic and
edaphic conditions. These thresholds should be based on quantitative
assessments, including the responses of various soil organisms (ranging
from micro to meso- and macro-fauna) to different manure management
practices under different soil types, agricultural practices and climatic
conditions. Manure management recommendations should meet sus-
tainable farming methods (see 3.2.1) to fully protect and promote soil
biodiversity and to achieve sustainable manure management (Fig. 4).
Only sustainable manure management can counteract any potential
environmental risks posed by farmyard manure additions.
Despite the potential of these recommendations to achieve more
sustainable farming systems, in order to be fully adopted and imple-
mented by farmers, we would need a radical transformation in the way
agriculture is conducted. Coupling better manure quality and increasing
soil taxonomical and functional diversity could provide a self-regulating
system that maintains current production levels (Thiele-Bruhn et al.,
2012; Lavelle et al., 2016).
3.3. The relevance of soil biodiversity in European legislation on manure
management
The main instrument that drives agricultural production in the EU is
the Common Agricultural Policy (CAP), combining economic incentives
with regulatory requirements (Paleari, 2017). For example, the Nitrates
Directive is a Statutory Management Requirement, which is part of the
conditionality of the CAP. This needs to be fullled by farmers to receive
subsidies. Within the conditionality of the CAP, Member States (MS)
must decide certain Good Agriculture Environmental Conditions
(GAEC), which then are mandatory for farmers. The post-2020 CAP
introduces in GAEC 5, the Farm Sustainability Tool (FaST). Based on
information of selected farms, the type of crops, the number of animals
and the amount of manure, the tool provides data for nutrient man-
agement, including “data on soil, the proximity of protected areas and
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
14
legal limits on the use of nutrients” (European Commission, 2019). The
tool aims to facilitate nutrient management and adapt it to local site
conditions (European Commission, 2019). Apart from its conditionality,
the CAP provides mostly voluntary options (such as eco-schemes and
agri-environment-climate payments) to protect and enhance soil
biodiversity.
Various other regulations (e.g., Regulation 1069/2009 for the
transport, trade and disposal of manure), apply directly to all MS while
directives are implemented at the MS-level into national legislation, e.g.,
the Nitrates Directive, the Air Quality and National Emission Ceilings
Directive and the Industrial Emissions Directive. For the implementation
at the MS-level, MS decide on best practices for manure management
and application based on the agronomic context (European Commission,
2014), for example, including nutrient balances (European Commission,
2000) and preferences for manure treatment (Hou et al., 2018). In leg-
islations targeting manure, there is no generally applicable denition of
manure quality. Different legislations address various aspects related to
manure (e.g., air emissions, nutrient losses to water, manure application
and management), but there is no overarching piece of legislation that
bundles all different aspects related to manure management. An analysis
of relevant legislation found in total eight policies (Table 4), impacting
manure management and consequently soil biodiversity (see Table 4; for
an overview of existing legislation on manure management and soil
biodiversity; see supplementary Table 3). Only the Organic Production
Scheme Regulation (2092/91) currently combines manure management
with farming practices that may enhance soil biodiversity.
Currently, there is no unied plan of N budgeting in the EU, and
indicators for nutrient management (such as the calculation for N sur-
pluses), water quality (Klages et al., 2020) and calculations for manure-
N availability (Webb et al., 2011) greatly vary at farm level. For
example, for the calculation of the nutrient use efciency, only 13 MS
(out of EU27 +UK) consider soil type, 14 MS consider crop type, 13 MS
consider the time of application, 8 MS consider climate, 7 MS consider
the method of application, 23 MS consider manure type and 7 MS ensure
the long term availability of the indicator (Webb et al., 2011). Similarly,
MS include digestates differently generated through anaerobic diges-
tion. For example, Spain and Italy do not address digestates specically,
while Austria, Belgium, Germany, Denmark, France, Lithuania, the
Netherlands and Poland handle digestates as fertiliser, soil amendment
or soil improver. The limitation set by the Nitrates Directive of 170 kg/
ha of N applicable to nitrate vulnerable zones includes digestates in
Ireland, Belgium, Denmark, Estonia and the Netherlands (if the input
material consists of at least 50% of manure) and Sweden (if input ma-
terial consists of 100% of manure). Austria, Bulgaria, Germany, Finland,
France, Italy, Latvia, and Slovenia omit digestates in their calculation;
the remnant MS did not answer the survey by the Joint Research Centre
(European Commission, 2013b). Some MS use very high-efciency rates
for digestates (Wallonia in Belgium calculated a 100% efciency; (Webb
et al., 2011)), which is not consistent with research results (M¨
oller and
Müller, 2012) and may increase leaching risks. Bloem et al. (2017)
recommended not to calculate the amount of applicable digestates or
manure based on its N content because of the low N:P ratios that favour
P surpluses and possible runoff. Few MS set maximum values of appli-
cable P such as Sweden (Bloem et al., 2017), Denmark (Wall et al.,
2011), Belgium and the Netherlands (Amery and Schoumans, 2014).
Also, the importance attributed to soil biodiversity varies among MS.
Although several countries collect data on soil organisms independently
(R¨
ombke et al., 2016), there is a lack of a standardised methodology at
the EU level (van Leeuwen et al., 2017). To overcome some of these
challenges, the European Commission has launched the EU Soil Obser-
vatory (Montanarella and Panagos, 2021). Also, the research activities
of the Joint Research Centre (JRC) in the European Commission aims at
standardising soil biodiversity monitoring at the EU level (Orgiazzi
et al., 2018). A soil biodiversity module has been introduced in LUCAS
topsoil survey and almost 1000 soil samples have been analyzed for soil
biodiversity through metabarcoding and metagenomics (Orgiazzi et al.,
2018). In particular, in 630 samples, including agricultural soils, met-
agenome was also sequenced. This permitted investigation into the
impact of management practices on soil functional genes (e.g., antimi-
crobial resistance genes). While the proposal for a Soil Framework
Directive, including threats to soil biodiversity, was rejected in 2007 due
to the blocking of ve Member States (MS), soil protection is currently
regulated at the MS level following the principle of subsidiarity (R¨
ombke
et al., 2016). However, the protection of soil biodiversity has recently,
gained greater prominence and was directly addressed in the Biodiver-
sity Strategy for 2030, within the EU Green Deal (Montanarella and
Panagos, 2021). Recently, the rst global soil biodiversity assessment
was presented by the Food and Agriculture Organisation (FAO, I, et al.,
2020). The EU has expressed the intention to improve sustainable soil
management that may promote, amongst other, soil biodiversity: the
“Farm to Fork Strategy” aims to reduce pesticide usage by 50%, fertiliser
application by 20% and nutrient losses by at least 50%, while the
Biodiversity Strategy 2030 targets to increase high-diversity landscape
features by at least 10% and organic farming from 7.2% in 2019 to 25%
in 2030 (European Commission, 2020d).
3.4. Analysis of manure-relevant EU legislation for potential integration
of soil biodiversity issues
Analysis of the legislation in place revealed the following limitations
that complicate sustainable manure management in the EU:
- Currently, the concentration of manure production is too high to be
absorbed by the local environment, leading to losses and unsus-
tainable management. Apart from approaching the unsustainable
sizes of animal farmed in the EU, prevailing manure practices such as
the application of raw manure and the often subsidised anaerobic
digestion must be further examined for their potential negative ef-
fects as soil amendments due to the quality of nitrogen sources, the
risk of contamination with heavy metals and antibiotics and the
potential nutrient leaching due to changed soil chemistry. Challenges
posed by a declining quality and an increasing amount of manure are
not adequately addressed in existing legislations, which also do not
include the potential negative effects on soil biodiversity (see sup-
plementary Table 3). Because policy frameworks regulating manure
management are scattered across various legislations, the complexity
of achieving sustainable soil management is increased (Thorsøe
et al., 2019).
- Only Organic Production Schemes regulate manure quality in com-
bination with agricultural practices that are benecial for soil or-
ganisms. Other legislation does not consider soil biodiversity in their
scope and consequently neither in their recommended practices. For
example, regulations permit the immediate application of high
amounts of slurry despite the lack of long-term studies on its effects
on soil organisms (Podmirseg et al., 2019). Furthermore, even in
organic farming, the application of conventional manure is
permitted, potentially introducing hormones, heavy metals, micro-
plastics and antibiotic gene resistance to organic soils, thereby
compromising environmental goals and quality standards of Organic
Production Schemes (Løes et al., 2017). The recycling of conven-
tional manure in organic agriculture remains an issue that needs to
be addressed in future policy developments. To address drug resi-
dues, upstream actions (drug administration to animals) may be most
suitable, which will be introduced in the update of the Veterinary
Medicinal Regulation EU/2019/6 coming into place in 2022.
- From 2022 onwards, the Fertilising Products Regulation will come
into place aiming to facilitate cross-border transport of high-quality
products. The Regulation will imply stricter quality controls for
organic soil improvers (e.g., maximum amounts of copper (300 mg/
kg dry matter) and zinc (800 mg/kg dry matter), minimum carbon
content) featuring the potential of pre-treated manure as a safe,
hygienic fertiliser product. However, thresholds have not been
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
15
calculated based on soil biota taxonomic data and no similar
thresholds for the application of raw slurry are available. This
Regulation will facilitate access to markets for organic fertilisers
based on manure, however, it will not overrule national legislation
since the Regulation is based on the principle of optional harmo-
nisation. While an additional material class is under development
which group all animal by-products, currently, the regulation only
considers composting and anaerobic digestion, both of which are
time and resource-intensive (i.e., requiring a large surface area for
logistics and high investments; (Alegbeleye and Sant’Ana, 2020).
Likely, farms with small amounts of manure apply anaerobic diges-
tion less frequently (Foged et al., 2011), but many MS subsidizes
anaerobic digestion (Banja et al., 2019; European Commission,
2012).
- The EU prioritises inputs from animal farming for biogas production
(European Commission, 2012) to limit indirect land-use changes and
prevent the conversion of abandoned or set-aside land providing
biomass for biogas production (Tamburini et al., 2020). For example,
the Renewable Energy – Recast to 2030 (RED II) restricts feedstocks
with “high indirect land-use change-risk” (Directive (EU) 2018/
2001). However, research on anaerobic digestion does not consider
the long term effects on local soil conditions and often ignores the
potential risks of pollutants in digestates (Grando et al., 2017; Meyer
and Edwards, 2014; Podmirseg et al., 2019).
- The Nitrates Directive is currently the most critical tool to mitigate
nutrient leaching from manure. Despite its standards set to reduce N
losses, full compliance has not yet been achieved (Buckwell and
Nadeu, 2016). For example, a stakeholder network analysis in the
Lombardy plain nds steady or increasing N surpluses calling for
better governance and monitoring due to farmers’ attitudes towards
inadequate N management (Musacchio et al., 2020). In France, in the
Seine basin, actions to reduce nitrates in water tables were not
effective and diffuse agricultural pollution has not decreased since
the 1990s. Since 2013, the number of water tables that have been
downgraded due to exceeding nitrate thresholds has doubled (Bou-
leau et al., 2020). The European Commission urged France (Euro-
pean Commission, 2020c), Italy (European Commission, 2020a),
Belgium and Spain (European Commission, 2020b) to comply with
nitrate thresholds for losses to water tables set in the Nitrates
Directive. If thresholds are not met, MS are brought in front of the
Court of Justice of the European Union, as was the case for Germany
(European Ombudsman, 2019). Legal processes led to stricter rules
for applying fertilisers (Bouleau et al., 2020) and Germany imple-
mented a new national fertiliser law in 2020 (Schulz, 2020).
- The Industrial Emissions Directive regulates manure management
for larger pig and poultry farms, including nutritional management
of animals, the collection, storage, treatment and land application of
manure (Directive 2010/75/EU), aiming to reduce overall adverse
environmental impacts. It does not set limits for heavy metals, which
pose a particular threat to soil biodiversity and the functions
provided.
- While certain exibility at the MS level is crucial to consider spatial
characteristics, certain standards should be applied EU-wide.
Currently, climate and soil conditions are not yet adequately
considered for manure management at the MS level. For example,
only a few MS consider the spatial soil properties and long term ef-
fects in calculating manure-N availability (Webb et al., 2011). Esti-
mates of nitrogen budgets provided by MS do not consider nutrient
accumulation and leaching to harm soil biodiversity (Klages et al.,
2020). Also, the effects of pre-treatment on nutrient availability are
ignored (Bloem et al., 2017) and the recovery of nutrients besides N
(e.g., in GAEC conditionality) is not yet included in current EU reg-
ulations (Garske et al., 2020).
- Amendments adopted by the European Parliament on 24 October
2017 on the proposal for a regulation laying down rules on the
making available on the market of CE marked fertilising products
((EC) No 1069/2009 and (EC) No 1107/2009 (COM(2016)0157 –C8-
0123/2016–2016/0084(COD)), recommended the inclusion of
recital 5a: “(5a) To ensure effective use of animal manure and on-
farm compost, farmers should use those products which follow the
spirit of ‘responsible agriculture’, favoring local distribution chan-
nels, good agronomic and environmental practice and in compliance
with union environmental law, … The preferential use of fertilisers
produced on-site and in neighboring agricultural undertakings
should be encouraged.” Such a provision can yield benets rather
than further environmental losses, if farmers are adequately trained
for on-farm treatment and application of manure based natural fer-
tilizers and biostimulants. Unfortunately, no such farmer training
curriculums have been designed and this proposal has been dropped
by EU 2019/2009.
4. The way forward in future EU policies of manure
management
At present, integrated manure management that addresses both the
risks and benets for soil biodiversity sits beyond the scope of legisla-
tion, despite increasing evidence drawing attention to the link between
both components (as highlighted in this review). Since manure man-
agement is a balancing act on the crossroad of conicting interests, so-
cially acceptable solutions (Burton, 2007) require the commitment from
the scientic community, policymakers, farmers and other stakeholders
involved (Gebrezgabher et al., 2014).
The scientic community emphasises the importance of considering
the environmental impacts of the entire manure management process
because integrated manure management may reduce environmental
hazards by 65% while doubling nutrient efciency (from 33 to 70%; De
Vries et al., 2015). However, besides addressing climatic emissions,
holistic manure management also needs to include the effects on below-
ground biodiversity. This review has shown that solutions combining
manure management and soil biodiversity enhancement simultaneously
could lower environmental risks. For example, applying raw manure
with biofertiliser or biostimulants may enhance plants’ nutrient uptake
and decreases the contamination risks derived from the nutrients surplus
(Paungfoo-Lonhienne et al., 2019; Schoebitz and Vidal, 2016; Thonar
et al., 2017). However, recommendations often focus on emissions only.
For example, the recommended manure incorporation practices to
reduce NH
3
emissions (Amann et al., 2018; UNECE, 2015) do not
consider the harmful effects on soil organisms. This shortcoming could
be overcome by coupling manure injection with climatic conditions
(prevent injection after raining events) to protect soil-surface dwelling
organisms.
Furthermore, since manure quality greatly diminishes with
increasing numbers of farmed animals (due to usage of antibiotics,
heavy metals), reducing the animal stocks is a key factor for reducing
undesired environmental impacts, including those on soil biodiversity.
While certain consequences of heavy animal farming can be ltered by
better treatment methods (e.g., pathogens), heavy metals cannot be
ltered by treatment methods and remaining antibiotic-resistant genes
have been found in digestates (Tien et al., 2017). Therefore, priorities
should focus on quality rather than quantity. This goes hand in hand
with reducing the consumption of meat and dairy products with more
environmental-friendly and nutritious plant-based diets, which simul-
taneously contribute to food security (De Boer and Aiking, 2011). To
prevent the replacement of farmed animals in the EU with cheaply im-
ported meat and dairy products, education and awareness campaigns
need to support innovation of food culture.
In addition to reducing animal density, better quality manures can be
obtained through a better monitoring of unprocessed manure, as well as
better manure application practices – comparable to those existing for
food composition, fertiliser and biostimulants formulations. A proposal
would be to better inform farmers by adding relevant information on
packaging (labelling), including the nutrient content, the source (type of
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
16
animal, place of origin, year), the quality as well as management of
manure that will be placed on the market using the Fertilising Product
Regulation. Standardising manure from highest to lowest quality based
on source, treatment method and expected impact on benecial soil
organisms would support more mindful manure application by farmers –
reducing negative impacts of low-quality manure and maximising pos-
itive impacts of high-quality manure. While organic fertilisers currently
still occupy a niche market (horticulture, oriculture, gardens), with an
increasing share of organic farms (EU goal of 25% until 2030), the
market is likely to grow. The grading and labelling of manure (in-
gredients) can help to better monitor the effects of short- and long-term
application of specic types of manure in various agricultural conditions
(soil, crop, climate). Such standardised procedures will allow tracking
the relocation of manure and its nutrients (Bernal et al., 2017; Musac-
chio et al., 2020), enabling redistribution of nutrients from surplus areas
to arable areas (Hanserud et al., 2017), as applied in the Netherlands
(Van der Straeten et al., 2011). Also, a transparent, standardised manure
quality will facilitate the transport to nutrient-decient regions (Luos-
tarinen et al., 2018). However, control and monitoring mechanisms
require extensive nancial resources. As a trade-off, local, traceable
manure chains must be created. In Austria, for example, a stand-
ardisation for the manufacturing and labelling applies for compost,
classifying compost depending on input materials, including heavy
metals (RIS Austria, 2001).
Soil biodiversity needs to be appropriately included when budgeting
manure amendments to provide better guidelines on the use of manure
and to reduce costs and environmental risks. Recommended values and
thresholds of manure amendments must consider the state of soil
biodiversity according to different land management practices. To move
towards integrated manure practices, alternative manure treatment
methods such as acidication of manure (Fangueiro et al., 2015; Ti et al.,
2019), or by combining anaerobic digestion with solid–liquid separation
and N removal (Finzi et al., 2020), need to be expanded while consid-
ering the long term effects of manure management on soil organisms
(Bünemann et al., 2006). Further, farming systems based on traditional
knowledge (TK) such as biodynamic farming or Indian Natural Farming
that adopt a systems approach to managing and enhancing soil and seed
biodiversity need to be studied more closely. Additionally, research
must address potential threats in manure management (pathogens,
antibiotic resistance genes, leaching of nutrients and heavy metals) as
well as benets (acting as biofertiliser, enhancing soil biodiversity,
adding organic matter), while considering soil-specic characteristics
and local status of soil biodiversity.
In parallel, standardised monitoring schemes of soil biodiversity
(such as LUCAS) will help to better understand the effects of land-use
practices, including manure practices, on soil biota. Monitoring soil
biodiversity needs to incorporate taxonomic and functional soil biodi-
versity to obtain information on the effects of land-use practices on
ecosystem services (Felipe-Lucia et al., 2020). Coupling a harmonised
methodology to determine the quality, amounts and treatment of
manure and manure-based fertiliser products with soil biodiversity data
could be integrated into the Farm Sustainability Tool for Nutrients
(FaST) or similar nutrient management tools for monitoring nutrient
budgets. The FaST proposed in the framework of the CAP Good Agri-
cultural and Environmental Conditions (GAECs), aims to facilitate sus-
tainable use of fertilisers for all farmers in the EU while boosting the
digitalisation. The mandatory usage of the FaST Nutrient Management
Tool from 2021 onwards brings the opportunity to move beyond the
amount and timing of manure eld application by integrating manure
quality, its origin and site characteristics (Buckwell and Nadeu, 2016).
FaST can play a role in accounting for the amount of manure produced
by farm and should also incorporate manure management in the
nutrient budget. For example, recommended values and thresholds of
manure amendments must consider the state of soil biodiversity ac-
cording to different land management practices and provide better
guidelines on the use of manure to reduce costs and environmental risks.
The tool also offers a great potential to educate farmers in linking soil
biodiversity and manure management and enhance soil biodiversity
functions in the next CAP update. The FaST is part of the post-2020
Common Agriculture Policy, which gives more exibility and re-
sponsibility to MS in regulating manure management and soil biodi-
versity according to their national needs. However, it is unknown how
MS will respond towards voluntary actions such as eco-schemes.
Data provided by the LUCAS would allow establishing thresholds
and indicators to prevent threats posed by inappropriate manure prac-
tices. Indicators are crucial as quantitative targets in regulatory in-
struments, and could contribute to better manure treatment and nutrient
management (Daxini et al., 2018; Savage and Ribaudo, 2013). There-
fore, either the scope of current legislations need to be expanded, or a
new legislation is required that considers the impact of manure on soil
biodiversity. Both strategies require the same extent of the legal process
and would allow implementing the direction set at a strategic European
political level, where the importance of protecting soil biodiversity is
increasing. At a strategic level, the EU is increasingly aware of the
importance of protecting soil biodiversity: for example, the Biodiversity
Strategy 2030 explicitly mentions the critical role of soil biodiversity.
Moreover, the aims of both the Farm to Fork Strategy and Circular
Economy Action Plan indirectly protect and promote soil biodiversity.
For example, the Biodiversity Strategy 2030 explicitly mentions the
critical role of soil biodiversity.
Since the Fertilising Product Regulation opens a market for manure
and biostimulant products and will also more strictly regulate mineral
fertilisers from 2022 onwards, the access and value of organic fertilisers
are likely to increase. From an economic point of view, sustainable
manure management may contribute to partially replace mineral fer-
tilisers by increasing soil biodiversity and, consequently, nutrient ef-
ciency. This replacement could decrease the dependencies on imports of
mineral P (European Commission, 2018; Svanb¨
ack et al., 2019) and N;
therefore, it will request fewer fossil energy sources (European Com-
mission, 2013c; Smil, 2004). A trade-off of replacing mineral fertilisers
with manure will require wider knowledge-management as well as
knowledge dissemination. For implementing manure management
practices, farmers’ awareness of soil biodiversity is crucial (Sere-
brennikov et al., 2020). A study in Spain, Italy, Denmark and the
Netherlands found that legislations are the key factor contributing to
manure treatment (Hou et al., 2018). However, too strict regulations or
a lack of knowledge prevent the implementation of composting practices
(Viaene et al., 2016). Knowledge of on-farm composting or other
manure treatment methods needs to be spread, e.g., through training
and agricultural extension work (De Corato, 2020a).
Such a replacement would impose positive long-term effects and
sustainable soil practices replacing short-term gains derived from min-
eral fertiliser usage. At the farm level, change in agricultural practice is
either driven by personal motivation, economic incentives outweighting
the costs for adaption or by regulatory instruments such as nes, loss of
payments or reputation as a threat of potential consequences of non-
compliance (Prager and Posthumus, 2010). Consequently, lower prices
and higher efciency of bio-based fertilisers could counteract the lack of
trust towards such products and convince farmers to substitute (part of)
mineral with organic fertilisers (Tur-Cardona et al., 2018). Also, current
behavioural adaptations of European farmers often focus on economic
aspects (Dessart et al., 2019), neglecting psychological characteristics
such as attitudes, norms, perceived control and resources that positively
inuence the intentions of farmers (Daxini et al., 2018). For more
J. K¨
oninger et al.
Agricultural Systems 194 (2021) 103251
17
effective agri-environmental policies, the socio-economic aspects of
farmers’ decision-making must be addressed (Dessart et al., 2019; Sie-
bert et al., 2006).
5. Conclusions
Considering soil biodiversity in manure management provides a win-
win solution for enhancing agricultural productivity, reducing farmers’
costs and enabling positive environmental effects (fewer fertilisers, an
increase of soil organic carbon, mitigation of contamination risks and
improved nutrient efciency rates, particularly when compared to
mineral fertilisers as shown in two recent meta-analysis (Luo et al.,
2018; Zhang et al., 2020)). All these together can accelerate the move
towards sustainable food systems in the EU. Soil biodiversity can be
increased by implementing more sustainable farming practices such as
high diversity farming, limiting mechanisation and pollutants and by
adding microorganisms cultivated in manure (e.g., as done in biody-
namic farming practices). Microorganisms in manure that are adapted to
local environmental and soil properties are found to promote soil
biodiversity in the long term, to the extent that soil biodiversity in
manure fertilised soils exceeds that of inorganically fertilised soils by
50% (Bengtsson et al., 2005).
Since manure can also threaten soil biodiversity, potential risks
should be addressed, e.g., by treating contaminated manure or by
avoiding the application of manure containing heavy metals. This re-
view provides an overview of benecial and harmful manure practices
vis-`
a-vis soil biodiversity. These practices should be selectively applied,
monitored and considered by policy makers to better protect and
meaningful enhance soil biodiversity. Our ndings may contribute to
policy developments as projections for 2030 predict an increase in liquid
slurry application (European Commission, 2014). This, in turn, may
cause signicant harm to soil biodiversity due to high concentration of
heavy metals, pathogens and antibiotics. As a rst step to tackle the EU’s
manure relevant problems, it is important to prioritise the manure
quality over quantity. Putting quantity over quality harms soil biodi-
versity and therefore, sustainable food systems. Soil biodiversity needs
to be appropriately included when budgeting manure amendments to
provide better guidelines on manure use and to reduce costs and envi-
ronmental risks. To counteract current mismanagement, increased
monitoring and integrated research are needed.
At a strategic level, the EU is increasingly aware of the importance of
protecting soil biodiversity. However, to achieve the ambitious goals of
the Farm to Fork Strategy and the Circular Economy Action Plan, new/
updated regulatory instruments are needed. At present, holistic manure
management measures that address both the risks and benets for soil
biodiversity are beyond the scope of existing legislation. Either the scope
of current legislations needs to be expanded, or new legislation is
required that considers the impact of manure on soil biodiversity.
Due to the high amounts of manure that are currently being pro-
duced, their disposal is prioritised in policy agendas rather than recog-
nising and building on its multiple benets. It is time to adopt a more
holistic approach that seizes the benets of farm manures for soil
biodiversity, as an additional strategy for soil protection in light of
current and future environmental challenges.
Declaration of Competing Interest
The authors declare that they have no known competing nancial
interests or personal relationships that could have appeared to inuence
the work reported in this paper.
Acknowledgements
The study received funding from the SoildiverAgro project nanced
by the European Union’s Horizon2020 research and innovation pro-
gramme under grant agreement No. 817819. The authors thank Dries
Huygens for valuable feedback and Leonidas Liakos for his assistance.
Authors also acknowledge funding for open access charge from Uni-
versidade de Vigo/CISUG.
Appendix A. Supplementary data
Supplementary data to this article can be found online at https://doi.
org/10.1016/j.agsy.2021.103251.
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