SOILSOILSOILSOIL2199-398XCopernicus GmbHGöttingen, Germany10.5194/soil-1-665-2015Biogeochemical cycles and biodiversity as key drivers of ecosystem services provided by soilsSmithP.pete.smith@abdn.ac.ukhttps://orcid.org/0000-0002-3784-1124CotrufoM. F.RumpelC.PaustianK.KuikmanP. J.ElliottJ. A.McDowellR.GriffithsR. I.AsakawaS.BustamanteM.HouseJ. I.SobockáJ.HarperR.PanG.WestP. C.GerberJ. S.ClarkJ. M.AdhyaT.ScholesR. J.ScholesM. C.https://orcid.org/0000-0001-5537-6935Institute of Biological & Environmental Sciences, Scottish Food Security Alliance-Crops and ClimateXChange, University of Aberdeen, 23 St Machar Drive, Aberdeen, AB24 3UU, UKDepartment of Soil and Crop Sciences & Natural Resource Ecology Laboratory, Colorado State University, Fort Collins, CO 80523-1499, USACNRS, IEES (UMR UPMC, CNRS, UPEC, INRA, IRD) and Ecosys (UMR INRA, AgroParisTech), Campus AgroParisTech, Bâtiment EGER, 78850 Thiverval-Grignon, FranceAlterra Wageningen UR, P.O. Box 47, 6700AA Wageningen, the NetherlandsEnvironment Canada, National Hydrology Research Centre, Saskatoon, Saskatchewan, S7N 3H5, CanadaAgResearch, Invermay Agricultural Centre, Private Bag 50034, Mosgiel, New ZealandCentre for Ecology & Hydrology, Maclean Building, Benson Lane, Crowmarsh Gifford Wallingford, OX10 8BB, UKGraduate School of Bioagricultural Sciences, Nagoya University, Chikusa, Nagoya 464-8601, JapanDepartamento de Ecologia, Universidade de Brasília, I.B. C.P. 04457. Campus Universitário Darcy Ribeiro – UnB. D.F. CEP: 70919-970 Brasília, BrazilCabot Institute, School of Geographical Sciences, University of Bristol, University Road, Bristol, BS8 1SS, UKNational Agriculture and Food Centre Lužianky, Soil Science and Conservation Research Institute Bratislava, Gagarinova 10, 827 13 Bratislava, SlovakiaSchool of Environmental Science, Murdoch University, South Street, Murdoch WA, 6150, AustraliaInstitute of Resources, Environment and Ecosystem of Agriculture, Nanjing Agricultural University, 1 Weigang, Nanjing 210095, ChinaGlobal Landscapes Initiative, Institute on the Environment (IonE), University of Minnesota, 325 Learning & Environmental Sciences, 1954 Buford Ave, St. Paul, MN 55108, USASoil Research Centre, Department of Geography and Environmental Science, School of Archaeology, Geography and Environmental Science, University of Reading, Whiteknights, Reading, RG6 6DW, UKSchool of Biotechnology, KIIT University, Bhubaneswar – 751024, Odisha, IndiaGlobal Change and Sustainability Research Institute and School of Animal, Plant and Environmental Studies, University of Witwatersrand, Private Bag 3, Wits 2050, South AfricaP. Smith (pete.smith@abdn.ac.uk)19November20151266568528April20151June20152November20159November2015This work is licensed under a Creative Commons Attribution 3.0 Unported License. To view a copy of this license, visit http://creativecommons.org/licenses/by/3.0/This article is available from https://soil.copernicus.org/articles/1/665/2015/soil-1-665-2015.htmlThe full text article is available as a PDF file from https://soil.copernicus.org/articles/1/665/2015/soil-1-665-2015.pdf
Soils play a pivotal role in major global biogeochemical cycles (carbon,
nutrient, and water), while hosting the largest diversity of organisms on
land. Because of this, soils deliver fundamental ecosystem services, and
management to change a soil process in support of one ecosystem service can
either provide co-benefits to other services or result in trade-offs. In
this critical review, we report the state-of-the-art understanding
concerning the biogeochemical cycles and biodiversity in soil, and relate
these to the provisioning, regulating, supporting, and cultural ecosystem
services which they underpin. We then outline key knowledge gaps and
research challenges, before providing recommendations for management
activities to support the continued delivery of ecosystem services from
soils.
We conclude that, although soils are complex, there are still knowledge gaps,
and fundamental research is still needed to better understand the
relationships between different facets of soils and the array of ecosystem
services they underpin, enough is known to implement best practices now.
There is a tendency among soil scientists to dwell on the complexity and
knowledge gaps rather than to focus on what we do know and how this
knowledge can be put to use to improve the delivery of ecosystem services. A
significant challenge is to find effective ways to share knowledge with soil
managers and policy makers so that best management can be implemented. A
key element of this knowledge exchange must be to raise awareness of the
ecosystems services underpinned by soils and thus the natural capital they
provide. We know enough to start moving in the right direction while we
conduct research to fill in our knowledge gaps. The lasting legacy of the
International Year of Soils in 2015 should be for soil scientists to work
together with policy makers and land managers to put soils at the centre of
environmental policy making and land management decisions.
Introduction
Soils play a critical role in delivering a variety of ecosystem services
(Scholes and Scholes, 2013). Management aimed at improving a particular
ecosystem service can either provide co-benefits to other services or
result in trade-offs (Robinson et al., 2013). Examples of some of the
synergies and trade-offs (Smith et al., 2013), the role of soils in
supporting ecosystem services, and their role in underpinning natural
capital (Dominati et al., 2010; Robinson et al., 2009, 2014) have recently
been reviewed. The ability of soils to provide services is principally
conferred by two attributes: the range of biogeochemical processes that
occur in the soil and the functionality of soil biodiversity. In the
following subsections we present the state-of-the-art understanding and
knowledge gaps on carbon, nutrient, and water cycling in soil, as well as the role of soils
as a habitat for organisms and as a genetic pool. We clarify how the
biogeochemical processes provide regulating, provisioning, and supporting
services, as well as the role of biodiversity (genetic diversity, functional
diversity, and abundance and activity of organisms) in supporting these
services. These functions collectively confer soil health, which is critical
for the underpinning of cultural services, among other things. A range of
soil services have been identified including soil as a source of raw
materials such as sand or clay, a surface for building infrastructure, and
an archive for landscape development and history of human soil use (e.g. Blum, 2002), but here we focus on those that map on to ecosystem services
listed in the Millenium Ecosystem Assessment (MA) (Millennium Ecosystem
Assessment, 2005).
The MA classified ecosystem services into supporting, regulating,
provisioning, and cultural services, and this categorization is widely used,
and though the scheme was not designed to fit all assessments (Fisher et
al., 2009), it has been modified for use in national ecosystem assessments
(e.g. UKNEA, 2011). More recently, the Common International Classification
of Ecosystem Services (CICES; Haines-Young and Potschin, 2012) was developed
to support environmental accounting in the European Union and in the United
Nations Statistical Division (European Commission et al., 2013, 2014). A major difference between the MA and the CICES
classification systems is that CICES does not include supporting services
(see below), which are treated as intermediate steps in the delivery of
final goods and services (Haines-Young and Potschin, 2012). In this review,
we include supporting services, since they are often referred to in the
literature, while accepting the CICES observation that supporting services
are not of direct benefit of people, although they are of great indirect
benefit. The MA supplemented by UKNEA (2011) for supporting services,
provides definitions and examples of provisioning, regulating, supporting,
and cultural services as follows.
Provisioning services are “physical products obtained from
ecosystems” and include food (including wild-harvested seafood and game,
cultivated crops, wild foods, and spices), raw materials (including timber,
pulp, skins, animal and vegetable fibres, organic matter, fodder, and
fertilizer), genetic resources (including genes for crop improvement and
health care), freshwater, minerals, medicinal resources (including
pharmaceuticals, chemical models, and test and assay organisms), energy
(hydropower, biomass feedstocks including biofuels, wood, and charcoal), and
ornamental resources (including fashion; handicraft; jewellery; pets;
worship; decoration; and souvenirs like furs, feathers, ivory, orchids,
butterflies, aquarium fish, shells, etc.).
Regulating services are “benefits obtained from the regulation of
ecosystem processes” and include carbon sequestration and climate
regulation, waste decomposition and detoxification, pollutant immobilization
and detoxification, purification of water and air, regulation of water flow
(including flood alleviation), and pest and disease control.
Supporting services are “ecosystem services that are necessary for
the production of all other ecosystem services” and include soil
formation, nutrient cycling, water cycling, primary production, and habitat
for biodiversity.
Cultural services are “nonmaterial benefits people obtain from
ecosystems through spiritual enrichment, cognitive development, reflection,
recreation, and aesthetic experiences” and include cultural (including use
of nature as motif in books, film, painting, folklore, national symbols,
architectural, advertising, etc.), spiritual and historical (including use
of nature for religious or heritage value or sense of place), recreational
experiences (including ecotourism, outdoor sports, and recreation), and science
and education (including use of natural systems for school excursions and
scientific discovery). Examples of cultural services underpinned by soils
are the terra preta soils of the Amazon Basin, representing the historical
cultural heritage of the region before European settlers; Histosols, which
are a vital component of peatland landscapes, underpinning the landscape/amenity value of these valued wild areas; and soils used as building
material for traditional houses providing cultural heritage values, such as
the mud brick houses in Bam in Iran and Shibam in Yemen. Since this paper
focuses on biogeochemical cycling and soil biota, cultural services are not
discussed further in detail in this review.
Schematic representation of where soil carbon, nutrient,
and water cycles, and soil biota underpin ecosystem services (adapted from
Smith et al., 2014). Role in underpinning each ecosystem service shown by C, soil carbon; N, soil nutrients; W, soil water; and B, soil biota.
Only soil carbon, nutrient, and water cycles, and soil biota are considered,
so the figure does not represent a comprehensive overview of soil ecosystem
services, which have been reviewed recently elsewhere (e.g. Robinson et
al., 2013, 2014).
Figure 1 summarizes the ecosystem services underpinned by soils. In the
following sections, we examine the state-of-the-art understanding of carbon,
nutrient, and water cycles and biodiversity in soils, and show how these
underpin the provisioning, regulating, supporting, and cultural ecosystem
services described above. We then discuss the knowledge gaps across all of
these areas, recommend key foci for future research, and present
recommendations for practices and policies to support the continued delivery
of these ecosystem services from soils.
Management actions affecting the soil carbon cycle and their impact
on ecosystem services.
Management action or other driver of changeProvisioning service impactRegulating service impactSupporting service impactCultural service impactLand use change (conversion of forest/grassland/wetland to cropland)Increased production of food, fibre, and energy crops; reduced availability of natural raw materials; potential change in hydrology/water availabilityDecreased soil C sequestration and storage – increased GHG flux; increased erosion and sediment yield – reduced regulations of water flow and qualityPrimary production may be changed; nutrient recycling reduced if no inputs, increased if there are inputsLower recreation value; may have impact on cultural value in recreating diverse landscapesLand use change (establishment of forest or grassland on agricultural land)Raw material provision may be increased; agricultural production likely decreased (but not always, e.g. agroforestry)Increased C sequestration; increased regulation of water flow and qualityPrimary production may be changed; increased water recyclingIncreased recreation value; may have impact on cultural value in recreating diverse landscapesIntensified nutrient management through fertilization and limingIncreased production of food and other raw materialsEffect on net soil C sequestration uncertain; increased GHG flux from fertilizer production and use; water and air pollutionIncreased primary production; increased nutrient recyclingSoil amelioration using organic amendments such as compost and biocharIncreased food production; more raw materials; more water available for plant growthIncreased C sequestration; increased water purification valueIncreased primary production; increased nutrient cycling; improved water infiltration and retentionDiversification of crop production systems (i.e. more perennials, reduced bare fallow)Potential impact on agricultural production (±); more diverse productsIncreased C sequestration; increased purification valueChanged primary production; increased nutrient retention; improved water infiltration and retentionImproved cultural value from more diverse landscapesReplacement of hay forage production with pasture use on grasslandsNo impactEffect on C sequestration uncertainIncreased recreation value; may have impact on cultural value in recreating diverse landscapesImproved grazing managementIncreased food production; reduced runoff and improved water useIncreased C sequestration; increased purification value; water flow regulationIncreased primary production; improved water infiltration and retentionSoils and the carbon cycle
Soil C stocks: Carbon (C) storage is an important ecosystem function of soils that has
gained increasing attention in recent years. Changes in soil C impacts on,
and feedbacks to, the Earth's climate system through emissions of
CO2
and CH4 as well as storage of carbon removed from the atmosphere during
photosynthesis (climate regulation; Table 1). Soil organic matter itself
also confers multiple benefits for human society, e.g. enhancing water
purification and water holding capacity, protecting against erosion risk,
and enhancing food and fibre provision through improved soil fertility
(Table 1; Pan et al., 2013, 2014).
Soil is an important C reservoir that contains more C (at least
1500–2400 PgC) than the atmosphere (590 PgC) and terrestrial vegetation (350–550 PgC)
combined (Schlesinger and Bernhardt, 2013; Ciais et al., 2013), and an
increase in soil C storage can reduce atmospheric CO2 concentrations
(Table 1; Whitmore et al., 2014). All three reservoirs of C are in constant
exchange but with various turnover times, with soil as the largest active
terrestrial reservoir in the global C cycle (Lal, 2008). Carbon storage in
soils occur in both organic and inorganic form. Organic C stocks in the
world's soils have been estimated to comprise 1500 Pg of C to 1 m depth and
2500 Pg to 2 m (Batjes, 1996). Recent studies have shown that the soil C pool to 1 m
depth may be even greater and could account for as much as 2000 Pg. These
higher values are mainly based on increased estimates of the C stored in
boreal soils under permafrost conditions (Tarnocai et al., 2009), in which
decomposition is inhibited by low temperature, lack of oxygen, and low pH in
waterlogged soils, e.g. peats (Smith et al., 2010). Although the highest
C
concentrations are found in the top 30 cm of soil, the major proportion of
total C stock is present below 30 cm depth (Batjes, 1996). In the northern
circumpolar permafrost region, at least 61 % of the total soil C is stored
below 30 cm depth (Tarnocai et al., 2009). Peatlands are particularly
important components of the global soil carbon store, covering only 3 % of
the land area but containing around 500 PgC in organic-rich deposits
ranging from 0.5 to 8 m deep (Gorham, 1991; Yu, 2012), with storage in deeper
layers as yet unquantified.
In arid and semi-arid soils, significant inorganic C can be present as
carbonate minerals (typically Ca/MgCO3, called “calcrete” or
“caliche” in various parts of the world), formed from the reaction of
bicarbonate (derived from CO2 in the soil) with free base cations,
which can then be precipitated in subsoil layers (Nordt et al., 2000). Soils
derived from carbonate-containing parent material (e.g. limestone) can also
have significant amounts of inorganic C. The inorganic C pool globally is
large, estimated to be ∼750PgC to a depth of 1 m (Batjes,
1996). However, in most cases, changes in inorganic C stocks are slow, are
not amenable to traditional soil management practices, and do not play a
significant role in terms of most ecosystem services (though a major
exception is the geoengineering proposal to add finely ground silicate
minerals to soils, which will then weather to carbonates, taking up
CO2
in the process; Köhler et al., 2010). Thus, further discussion of soil C
in this review will focus on soil organic C.
The net balance of soil C depends on the inputs of C to soils relative to C
losses. Losses can occur via mineralization (i.e. decomposition), leaching of
dissolved C, and carbonate weathering (Smith, 2012; Schlesinger and
Bernhardt, 2013). Thus, the soil organic C stock may either increase or
decrease in response to changes in climate and land use practices (Smith et
al., 2015). Furthermore, rates of SOC stock change in different parts of the
profile can vary for different soils and types of perturbation, because some
portion of the C stored in soil, mainly in topsoil, turns over rapidly,
while other soil C fractions can have a long residence time (von Lützow
et al., 2008; Rumpel and Kögel-Knabner, 2011). The accumulation of
stabilized C with long residence times in deep soil horizons may be due to
continuous transport, temporary immobilization and microbial processing of
dissolved organic matter within the soil profile (Kalbitz and Kaiser, 2012),
and/or efficient stabilization of root-derived organic matter within the
soil matrix (Rasse et al., 2005). The process of soil formation – i.e. the
development of depth, horizons, and specific properties – is itself a
supporting service (Table 1).
High SOC content also improves other chemical and physical soil properties,
such as nutrient storage (supporting service), water holding capacity
(supporting and regulating service), aggregation, and sorption of organic or
inorganic pollutants (regulating service). Carbon sequestration in soils may
therefore be a cost-effective and environmentally friendly way to not only
store C for climate regulation but also enhance other ecosystem services
derived from soil, such as agricultural production, clean water supply, and
biodiversity (Table 1; Pan et al., 2013) by improving soil organic matter
(SOM) content and thereby soil quality (Lal, 2004). Moreover, processes
which improve SOM may themselves provide services, e.g. use of cover crops,
which can provide provisioning or water regulation services while improving
soil C (Table 1). SOM or soil carbon are widely used proxy variables for
soil health (e.g. Kibblewhite et al., 2008).
C cycling: Carbon enters the soil as aboveground or belowground plant litter
and exudates. C input is not homogenous within the soil profile. Whereas
topsoil receives higher amounts of aboveground litter, subsoil C originates
from root C as well as dissolved C, transported down the soil profile. Root
C has a greater likelihood of being preserved in soil compared to shoot C,
and was therefore hypothesized to account for most of the SOC (Rasse et al.,
2005). The majority of plant litter compounds pass through and are modified
by the soil biota. Thus, SOM is composed of plant litter compounds as well
as microbial and, to a smaller extent, faunal decomposition products (Paul,
2014). It is a complex biogeochemical mixture comprising molecules derived
from organic material in all stages of decomposition. Some organic matter
compounds, including microbial decomposition products, may be stabilized for
centuries to millennia by binding to soil minerals or by physical occlusion
into micro-aggregates (von Lützow et al., 2008), for example with iron
oxyhydrates (Zhou et al., 2009), or through protection by occlusion within
soil aggregates (Dungait et al., 2012). The inherent chemical recalcitrance
of some plant litter compounds (e.g. lignin) has a minor influence on their
longevity in soil (Thévenot et al., 2010), whereas the location of SOM
within the soil matrix has a much stronger control on its turnover (Chabbi
et al., 2009; Dungait et al., 2012). Mineral-associated SOM is predominantly
composed of microbial products (Miltner et al., 2012). Therefore, microbial
use efficiency of plant inputs largely determines SOM
stabilization through interaction with the mineral phase (Cotrufo et al.,
2013), in addition to the environmental controls discussed elsewhere in this
section. In peatlands, organic matter is stabilized by high water tables
that slow down biological activity and decomposition. SOM is mineralized to
carbon dioxide (CO2) in aerobic environments, or reduced to methane
(CH4) in anaerobic environments. Soil CO2 efflux, resulting from
SOM mineralization, and from rhizosphere respiration and inorganic C
weathering, is the largest terrestrial flux of CO2 to the atmosphere
(∼60PgC; the sink of carbon on the other hand contributes
to the climate regulation service; Smith, 2004). This flux is an order of
magnitude larger than anthropogenic CO2 emissions due to fossil fuel
burning and land use change (1.1 PgCyr-1; Ciais et al., 2013). Under anaerobic
conditions, CH4 is formed by methanogenic microorganisms. A proportion
of this CH4 is oxidized to CO2 by methanotrophic microorganisms,
but a proportion can be emitted from the soil surface (Reay et al., 2010).
Since CH4 is many times more potent as a greenhouse gas than
CO2
on a per-molecule or per-mass basis (Ciais et al., 2013), soil CH4
emissions and their mitigation play an important role in climate regulation
(Table 1).
Fire may affect many ecosystem services, including C sequestration. For
fires in natural ecosystems, a decrease in soil C storage is often observed
initially, but through positive effects on plant growth, as well as input of
very stable pyrogenic C, C storage may increase at longer timescales
(Knicker, 2007). An additional long-term C pool in many soils is
pyrogenic carbon (PyC), formed from partially combusted (i.e. pyrolysed)
biomass during wildfires or other combustion processes (Schmidt and Noack,
2000). Globally, soils are estimated to contain between 54 and 109 Pg PyC
(Bird et al., 2015). Some of this PyC has a highly condensed aromatic
structure that retards microbial decay, and can thus persist in soils for
relatively long periods (Singh et al., 2012). Soil amended with industrially
produced PyC (biochar) as a climate mitigation technique often shows no
increase in soil respiration despite the additional carbon, the reduced
ecosystem carbon turnover results in increased soil carbon storage (Stewart
et al., 2013). PyC additions to soil affect regulating ecosystem services,
such as C sequestration, nutrient cycling, and adsorption of contaminants.
However, PyC properties, and as result their effect on ecosystem services,
may be strongly dependent on fire conditions.
Factors influencing soil C storage: Fundamentally, the amount of C stored in a given soil is determined by the
balance of C entering the soil, mainly via plant production but also through
manures or amendments such as organic sludge or biochar, and C leaving the
soil through mineralization (as CO2), driven by microbial processes,
and to a lesser extent leaching out of the soil of dissolved carbon and
carbonate weathering. Locally, C can be lost or gained through soil erosion
or deposition, leading to a redistribution of soil C, at landscape and
regional scales (van Oost et al., 2007).
Consequently, the main controls on soil C storage are the amount and type of
organic matter inputs, the efficiency by which this is used by microbes, and
the capacity of the soil to retain it by physical or chemical stabilization
(Cotrufo et al., 2013). In most natural and agricultural ecosystems, plant
productivity and subsequent death and senescence of biomass provide the
input of organic C to the soil system (Table 1). Thus, higher levels of
plant residue inputs will tend to support higher soil organic carbon stocks,
and vice versa (Paustian et al., 1997), though this does not continue indefinitely
(Zvomuya et al., 2008). Plants also affect soil C cycling by their specific
mycorrhizal associations (Brzostek et al., 2015). Shifts in specific
mycorrhizal associations affect SOM storage by contributing to both SOM
formation and decomposition. Ectomychorrizhal turnover is a dominant process
of SOM formation (Godbold et al., 2006), possibly due to the more
recalcitrant nature of the chitin in fungal tissues, compared to the
cellulose and lignin in plant residues. In arbuscular mycorrhizal fungi, it
has been suggested that glomalin, a highly resistant glycoprotein, has an
active role in aggregate formation and SOM stocks (Rillig, 2004). Symbiotic
mycorrhizal fungi can also directly impact the turnover of organic matter by
the production of exo-enzymes (Averill et al., 2014; Finzi et al., 2015).
In many regions of the world, SOM accumulates because of inhibition of
microbial SOM decomposition, due to cold, dry, or anoxic conditions
(Trumbore, 2009). In general, when water is not limiting, higher soil
temperatures increase the rate of microbial decomposition of organic matter.
Thus soil temperature is a major control of SOM storage in soil C cycle
models (Peltoniemi et al., 2007). The temperature sensitivity of SOM
decomposition is not, however, as straightforward as represented in most
models but varies between the many different forms of chemical and physical
protection of organic matter in soil (Conant et al., 2011; Zheng et al.,
2012). Water influences soil C storage through several processes. Moist, but
well-aerated, soils are optimal for microbial activity and decomposition
rates decrease as soils become drier. However, flooded (saturated) soils
have lower rates of organic matter decay due to restricted aeration and thus
often have very high amounts of soil C (e.g. peat soils). High precipitation
may also lead to C transport down the soil profile as dissolved and/or
particulate organic matter, as well as lateral transport through soil
erosion and deposition. During dry periods, SOM decomposition is decreased,
but after rewetting there may be an accelerated pulse of CO2 emission
in aerobic soils (Borken and Matzner, 2008), whereas drought and lowering
water tables may increase decomposition in naturally anaerobic peats
(Freeman et al., 2001; Clark et al., 2012). However, the effect of drought is
not only direct via soil microbial activity. There are feedback loops
concerning drought and C storage via plant activities, such as litter input and
rhizodeposition. Drought was found to affect plant litter composition
(Sanaullah et al., 2014), plant C flow and root exudation (Sanaullah et al.,
2012), as well as the resulting enzyme activities in the rhizosphere
(Sanaullah et al., 2011).
C cycling in soils is strongly linked to the cycling of N and P. Since the
C:N:P stoichiometry in SOM is generally lower than in plant material –
i.e. there is more N and P per unit C – C generally accumulates in aerobic soil
where nutrients are not limiting (Alberti et al., 2014). Nevertheless, an
increase in organic C is often accompanied by increased N resource use
efficiency in croplands (Pan et al., 2009), especially when SOC is increased
with biochar (Huang et al., 2013). In nutrient-limited peatlands, inputs of
nitrogen and/or phosphorus within the tolerance levels of sensitive plant
species have increased rates of carbon accumulation (Aerts et al., 1992;
Turunen et al., 2004; Olid et al., 2014). The relationship between nutrients
and C cycling is not straightforward, since nutrients are also needed by
soil microbes to degrade SOM. Thus, nutrient addition can either decrease or
increase C storage, depending on the initial SOM stoichiometry, the ability
of the soil minerals to stabilize microbial products of decomposition, and
the simultaneous effects on plant productivity and organic matter inputs to
soils.
Management actions affecting soil nutrient cycles and their impact
on ecosystem services.
Management action or other driver of changeProvisioning service impactRegulating service impactSupporting service impactCultural service impactIntensive addition of mineral fertilizersIncreased food, fibre, and feedstock productionReduced water quality through eutrophication, reduced air quality through emission and volatilization of reactive N gasesIncreased primary production; alteration of the nutrient and C cycling; possible reduction of biodiversityUse of organic soil amendments (e.g. manure, composts and biochar)Increased food, fibre, and feedstock production; may increase water retentionIncrease C sequestrationIncrease nutrient retentionImplementation of no-tillIncrease nutrient retentionPrecision agricultureIncrease efficient production of foodReduced GHG emissions per unit productionReduce consumption of water and nutrient by improving use efficiencyPrescribed use of fire for pasture managementIncrease feedstock productionIncrease C sequestration by conversion to BCReduce N recycling by storing black nitrogenUse of biological soil supplementsStimulate productivity; act as fertilizersMay improve pest and disease controlImproved nutrient cycling
The amount and type of clay particles (and to a lesser extent silt
particles) are the major factors controlling the quantity and composition of
soil C (Sollins et al., 1996; von Lützow et al., 2006). Clays are mainly sheet-like
crystals of silicon and aluminium, known as phyllosilicates, often located
as skins coating soil aggregates. In clay-rich soils, higher organic matter
content and a greater concentration of O-alkyl C derived from
polysaccharides may be expected compared to sandy soil, which are
characterized by lower C contents and high concentrations of alkyl C (Rumpel
and Kögel-Knabner, 2011). Aliphatic material may be responsible for the
hydrophobicity of soils, which can lead to reduced microbial accessibility
and therefore increased C storage (Lorenz et al., 2007). Many of the
OM-matrix interactions are driven by expandable and non-expandable
phyllosilicates, which interact with organic compounds through their large
surface areas, micropores, and micro-aggregation, particularly in acid
soils. In neutral and calcareous soils, polyvalent cations (especially
Ca2+) predominate in the interaction mechanism, forming bridges between
the largely negatively charged SOM and negatively charged phyllosilicates
(Cotrufo et al., 2013). Short-order silicates, like allophane, provide some
of the strongest organo-mineral interactions and stabilize both proteins and
carbohydrate monomers, though their occurrence is very geographically
restricted (Buurman et al., 2007; Dümig et al., 2012: Mikutta and Kaiser,
2011). Pedogenic oxides (for example iron oxyhydrates in rice paddies)
usually act as a coating of soil mineral particles and stabilize carbon,
contributing to a higher C storage and stability compared to other soils
(Song et al., 2012).
Bioturbation (the mixing of soil by organisms) may further influence the
amount as well as the chemical nature of soil C. It greatly influences the
heterogeneity of soils by creating hotspots of carbon and biological
activity. On biologically active sites, incorporation and transformation of
organic compounds into soil is usually enhanced, leading to more
organo-mineral interactions and increased C storage (Wilkinson et al.,
2009).
Microbial decomposition of SOM may be stimulated by the input of labile
(easily decomposed) organic matter through the priming effect (Jenkinson et
al., 1971). Positive priming refers to greater mineralization of otherwise
stable C through shifts in microbial community composition and activity
(Fontaine et al., 2003). However, in some cases, the addition of organic
matter to soil may also impede mineralization of native SOM (negative
priming effect), thereby protecting SOM from its decomposition. Plant
communities (Table 1) are the main controlling factors of these processes
because they influence organic matter input and microbial activity by their
effects on soil water, labile C input, pH, and nutrient cycling (Kuzyakov et
al., 2000).
By storing and cycling C, nutrients, and water, soils provide supporting
services like soil formation and nutrient and water retention, which
underpin both primary production and landscape hydrology (the processes
which deliver provisioning services such as food, fibre, and water; Table 1),
in addition to the regulating services such as climate regulation already
discussed (Fig. 1). To ensure that soils continue to provide these key
services, soil will require to be managed for C preservation – thus
mitigating climate change – while simultaneously permitting continued SOM
recycling (Table 1). Janzen (2006) pointed to this dilemma, that there is a
trade-off between improved soil fertility to support the provisioning
services of food/timber production and the regulating service of soil carbon
sequestration aiding climate regulation. Despite knowledge on which
practices are likely to lead to improved SOC status, a better understanding of
the controls on SOM distribution, stabilization, and turnover will help to
better target these practices. This will be an important contribution to the
mitigation of greenhouse gases, while assuring decomposition and, with it,
the cycling of nutrients necessary to support food production. Table 1
summarizes management actions affecting the soil carbon cycle and their
impacts on ecosystem services.
Global (a) nitrogen (N) and (b) phosphorus (P) fertilizer use
between 1961 and 2012 split for the different continents in Mt P per year;
plotted from FAOSTAT data (FAOSTAT, 2015).
Soils and nutrient cycles
Soils support primary production among other services, which in turn
delivers the provisioning services of food and fibre production (Table 2).
As such, soils are vital to humanity since they provide essential nutrients,
such as nitrogen (N), phosphorus (P), and potassium (K) and many trace
elements that support biomass production, which is essential for the supply
of human and animal food, for energy and fibre production and (future)
feedstock for the chemical industry (Table 2). Since the 1950s, higher
biomass production and yield increases have been supported by fertilizers
derived from mined minerals or industrial synthesis (Fig. 2).
Intensification of agricultural practices and land use has in many regions
resulted in a decline in the content of organic matter in agricultural,
arable soils (Table 2; Matson et al., 1997; Smith et al., 2015). In some
areas, extensive use of mineral fertilizers has led to atmospheric
pollution, greenhouse gas emissions (e.g. N2O, very important for
climate regulation), water eutrophication, and human health risks (Galloway
et al., 2008), thereby negatively affecting the regulating services of soil,
air, and water quality (Table 2; Smith et al., 2013). During the 21st
century, it is likely that the human population and demand for food, feed,
and energy will rise. In order to sustain biomass production in the future,
and to avoid negative environmental impacts, fertile soils need to be
preserved and soil fertility needs to be restored where lost. This can be
done through both the recycling and accumulation of sufficient amounts of
organic matter in soils (Janzen, 2006), through a combination of plant
production and targeted additions of organic and mineral amendments to soils
(see Sect. 2).
The soil function “fertility” refers to the ability of soil to support and
sustain plant growth, which relates to making available N, P, other
nutrients, water, and oxygen for root uptake. This is facilitated by (i) their
storage in soil organic matter, (ii) nutrient recycling from organic to plant
available mineral forms, and (iii) physical–chemical processes that
control their sorption, availability, displacement, and eventual losses to
the atmosphere and water (Table 2). Managed soils are a highly dynamic
system and it is this very dynamism that makes the soil work and supply
ecosystem services to humans. Overall, the fertility and functioning of
soils strongly depend on interactions between the soil mineral matrix,
plants, and microbes; these are responsible for both building and decomposing
SOM, and therefore for the preservation and availability of nutrients in
soils (Cotrufo et al., 2013). To sustain this service, the cycling of
nutrients in soils must be preserved (Table 2).
Applied and excess nitrogen and phosphorus in croplands. Nitrogen
and phosphorus inputs and excess were calculated using a simple mass balance
model (West et al., 2014), extended to include 175 crops. To account for both
the rate and spatial extent of croplands, the data are presented as kilogram per
hectare of the landscape. (a) Applied nitrogen, including N
deposition; (b) applied phosphorus; (c) excess nitrogen; and (d) excess phosphorus.
After carbon, N is the most abundant nutrient in all forms of life, since it
is contained in proteins, nucleic acids, and other compounds (Galloway et
al., 2008). Humans and animals ultimately acquire their N from plants, which
on land is mostly taken up in mineral form (i.e. NH4+ and
NO3-) from the soil. The parent material of soils does not contain
significant amounts of N (most other nutrients such as P largely originate
from the parent material). New N mostly enters the soil through the fixation
of atmospheric N2 by a specialized group of microorganisms. However,
the largest flux of N within the soils is generated through the continuous
recycling of N internal to the plant–soil system: soil mineral N is taken up
by the plant, is fixed into biomass, and eventually N returns in the form
of plant debris to the soil. Here microorganisms decompose it, mineralizing
part of the N and making it newly available for plant growth, while
transforming the other part into SOM, which ultimately is the largest stock
of stable N in soil. Generally, N cycles tightly in the system with minimal
losses. Nitrogen is lost from the soil to the water system by leaching and
to the atmosphere by gas efflux (NH4, N2O, and N2). In most
terrestrial natural ecosystems, N availability limits productivity. Through
the cultivation of N2 fixing crops, the production and application of
mineral N fertilizer, the increasing application of animal manure from livestock
and bio-wastes, and the unintentional deposition of atmospheric reactive
N
(ultimately derived from industrial-era human activities), humans have
applied twice as much reactive N to soils as the N introduced by natural
processes, significantly increasing biomass production on land (Vitousek and
Matson, 1993; Erisman et al., 2008). In some regions of the world, mineral
fertilizer is applied in excess of plant requirement, but in other regions,
in particular in Sub-Saharan Africa, where economic constraints limit the use
of fertilizers, productivity is strongly limited by soil available N and
other nutrients, notably P and K (N and P; Fig. 3).
Phosphorus derived from parent material, through weathering, cycles
internally in the plant–soil system between biochemical molecules (e.g. nucleic acid, phospholipids) and mineral forms after decomposition
(e.g. H3PO4). In soils, P is among the most limiting of nutrients,
since it occurs in small amounts and is only available to plants in its
dissolved ionic forms, which promptly react with calcium, iron, and aluminium
cations to form highly insoluble compounds. Largely in these forms, P is
lost to the aquatic system through erosion and surface runoff. Losses may
also occur in dissolved form, for instance via subsurface flow and groundwater
(McDowell et al., 2015). An important form of loss is in the export of
organic P in agricultural products. Due to widespread agricultural
P
deficiencies, humans started to mine “primary” P from guano or rock
phosphate deposits and added it to soils in the form of mineral fertilizer
(Fig. 2). This external input has led to positive agronomic P balances
(MacDonald et al., 2011) and excesses of P and N in many regions (West et
al., 2014; Fig. 3). There are large variations across the world, with high
surpluses in the USA, Europe, and Asia and deficits in Russia, Africa, and
South America (Fig. 3). Since plant P uptake is a relatively inefficient
process with roughly 60 % of the total P input to soils not taken up in
the short term, a 3-fold increase in the export of P to water bodies has
been estimated, with significant impacts on water quality (Bennett et al.,
2001).
Soil functions related to the water cycle and ecosystem
services.
Soil functionMechanismConsequenceEcosystem serviceStores (storage)Water held in soil pores supports plant and microbial communitiesBiomass production Surface protectionFood Aesthetics Erosion controlAccepts (sorptivity)Incident water infiltrates into soil with excess lost as runoffStorm runoff reductionErosion control Flood protectionTransmits (hydraulic conductivity)Water entering the soil is redistributed and excess is lost as deep percolationPercolation to groundwaterGroundwater recharge Stream flow maintenanceCleans (filtering)Water passing through the soil matrix interacts with soil particles and biotaContaminants removed by biological degradation/retention on sorption sitesWater quality
Clearly, management practices need to be implemented that sustain, restore,
or increase soil fertility and biomass production by promoting the accrual
of SOM and nutrient recycling, applying balanced C amendments and
fertilization of N, P, and other nutrients to meet plant and soil
requirements, while limiting the addition of excess fertilizer and retaining
nutrients in the soil–plant system (Table 2). C, N, and P cycling in
soils is coupled by tight stoichiometric relationships (e.g. relatively
fixed C:N:P in plants and microorganisms; Güsewell, 2004); thus their
management needs to be studied in concert. Nutrient management has been
extensively studied, with the aim of identifying and proposing management
practices (e.g. precision agriculture) that improve nutrient use efficiency
and productivity and reduce potentially harmful losses to the environment
(Table 2; van Groenigen et al., 2010; Venterea et al., 2011). Yet, our ability to predict
the ecosystem response to balanced fertilization is still limited, and
effectiveness and reliability would benefit from continued monitoring of
efforts. Further benefits are anticipated from improved plant varieties with
root morphologies that have better capacity to extract P from soils or use
it more efficiently, perhaps in concert with mycorrhizal symbionts.
Fertilization with nutrients other than N and P has been less well explored
within the realm of understanding soil organic matter responses to
agricultural C inputs and the potential to restore and increase soil organic
matter (e.g. Lugato et al., 2006). Hence, we stress the importance of an
integrated approach to nutrient management, which supports plant
productivity while preserving or enhancing SOM stocks, and reducing nutrient
losses to the atmosphere or water resources. Several issues exist where
prediction and optimization of performance would benefit from relevant and
continued data acquisition for the range of climate and environmental and
agro-ecological conditions. Table 2 summarizes some management actions
affecting soil nutrient cycles and their impacts on ecosystem services.
Soils and the water cycle
Soils provide important ecosystem services through their control on the
water cycle. These services include provisioning services of food and water
security, regulating services associated with moderation, and purification of
water flows, and they contribute to the cultural services of landscapes/water
bodies that meet recreation and aesthetic values (Table 3; Dymond, 2014). At
the pedon to hillslope scale, water stored in soil is used for
evapotranspiration and plant growth that supplies food, stabilizes the land
surface to prevent erosion, and regulates nutrient and contaminant flow. At a
catchment and basin scale, the capacity of the soil to infiltrate water
attenuates stream and river flows and can prevent flooding, while water that
percolates through soil can replenish groundwater that can maintain water
supplies and sustain surface water ecosystems while promoting a continued
flow during periods of reduced precipitation (Guswa et al., 2014).
The soil functions of accepting, storing, transmitting and cleaning of water
shown in Table 3 are inter-related. Soil water storage depends on the rate
of infiltration into the soil relative to the rate of precipitation. Soil
hydraulic conductivity redistributes water within and through the soil
profile. The infiltration rate and hydraulic conductivity both depend on the
water stored in the soil. The initially high rate of infiltration into dry
soil declines as the soil water content increases and water replaces air in
the pore space. Conversely, hydraulic conductivity increases with soil
moisture content as a greater proportion of the pores are transmitting
water. Water content and transmission times are also important to the
filtering function of soil because contact with soil surfaces and residence
time in soil are important controls on contaminant supply and removal
(McDowell and Srinivasan, 2009).
The quantity of water which a soil can store depends on the thickness of the
soil layer, its porosity, and soil matrix–water physical interactions. The
latter are expressed as a water retention curve, the relationship between
the soil water content and the forces holding it in place. The porosity and
water retention curve are in turn influenced primarily by the particle size
distribution and the soil bulk density, but also by the amount of SOM and the
macropores created by biotic activity (Kirkham, 2014).
Management actions affecting the soil water cycle and
their impact on ecosystem services.
Management action or other driver of changeProvisioning service impactRegulating service impactSupporting service impactCultural service impactLand use change (increase change of agricultural to urban)Decreased biomass; decreased availability of water for agricultural useIncreased impervious surface; decreased infiltration, storage, soil-mediated water regulationDecreased genetic diversity; reduction of rainfall recycling, e.g. in the tropicsDecreased natural environmentLand use change (increase change of arable to intensive grassland)Increased yield of animal over vegetable proteinIncreased C sequestration; greater requirement of water; stress on ecosystem health of downstream waterwaysIncreased genetic diversity associated with mixed pasturesChange from traditional values and aesthetic valueIrrigation (increase)Increased biomass over dryland agriculture; decreased availability of water for urban useIncreased C sequestration, but decreased filtration potentialImproved habitat for plant speciesInfrastructure alters landscape decreasing spiritual connection with catchmentDrainage (increasing in marginal land)Decreased soil saturation; increased biomass; removal of wetlandsDecreased C sequestration, denitrification, and flood attenuationBetter habitat for productive grassland plants, but loss of genetic diversityDecreased recreational potential (e.g. ecotourism)
Optimum growth of most plants occurs when roots can access both oxygen and
water in the soil. The soil must therefore infiltrate water, drain quickly
from saturation to allow air to reach plant roots, and retain and
redistribute water for plant use. An ideal soil for plant production depends
on the climatic conditions. Soil structural stability and porosity are also
important for the infiltration of water into soil. In addition to soil
texture, organic matter improves soil aggregate stability (Das et al.,
2014). While plant growth and surface mulches can help protect the soil
surface, a stable, well-aggregated soil structure that resists surface
sealing and continues to infiltrate water during intense rainfall events
will decrease the potential for downstream flooding resulting from rapid
overland flow. Porosity (especially macropores of a diameter ≥75µm)
controls transmission of water through the soil. In addition to
total porosity, the continuity and structure of the pore network are as
important to these functions as they are in filtering out contaminants in
flow. Furthermore, the soil must support biota that will degrade the
compounds of interest or have sorption sites available to retain the
chemical species. Soil organic matter is important for these roles and,
together with mineral soil (especially the clay fraction), provides sorption
sites (Bolan et al., 2011). Flow through macropores, which bypass the soil
matrix, where biota and sorption sites are generally located, can quickly
transmit water and contaminants through the soil to groundwater or
artificial drains, but for filtering purposes, a more tortuous route through
the soil matrix is more effective (McDowell et al., 2008). There are
multiple other links between soil biota and soil water, with water potential
in particular having a pivotal role in the structure, growth, and activity of
the soil microbial community (Parr et al., 1981).
Management of soil alters the ecosystem services provided by water (Table 4).
Soil conservation and sustainable management practices to combat
desertification help to retain soil organic matter, structural stability,
infiltration, and profile water holding capacity. The promotion of soil as a
C sink to offset greenhouse gas emissions generally helps to maintain or
improve soil hydrological functions as well. Deforestation, overgrazing, and
excessive tillage of fragile lands, however, will lead to soil structural
deterioration and a loss of infiltration, water retention, and surface water
quality (Table 4; Steinfeld et al., 2006). Anthropogenic modifications to
the water cycle can aid soil function. In dry regimes, inadequate soil
moisture can be mitigated through supplementary irrigation, and where
waterlogging occurs it can be relieved by land drainage. However, irrigation
and drainage can have consequences for water regulation services. Irrigation
that enables a shift to intensive land use can increase the contaminant load
of runoff and drainage (Table 4; McDowell et al., 2011). Furthermore,
drainage of wetland soils has been shown to reduce water and contaminant
storage capacity in the landscape and can increase the potential for
downstream flooding, as well as increasing the potential for GHG emissions due to the
rapid decomposition of SOC in soil and dissolved organic C in drainage water
(IPCC, 2013). The removal of surface or groundwater for irrigation disrupts
the natural water cycle and may stress downstream ecosystems and
communities. Irrigation of agricultural lands accounts for about 70 % of
ground and surface water withdrawals, and in some regions competition for
water resources is forcing irrigators to tap unsustainable sources.
Irrigation with wastewater may conserve fresh water resources, but the fate
of waterborne contaminants in soil and crops is a potential concern (Sato
et al., 2013).
Soils as a habitat for organisms and as a genetic resource
Soils represent a physically and chemically complex and heterogeneous
habitat supporting a high diversity of microbial and faunal taxa. For example,
10 g of soil contains about 1010 bacterial cells, representing more
than 106 species (Gans et al., 2005). Up to 360 000 species of animals
live predominantly in the soil – a large fraction of all animal species
(Decaëns et al., 2006). These complex communities of organisms play
critical roles in sustaining soil and wider ecosystem functioning, thus
conferring a multitude of benefits to global cycles and human
sustainability. Specifically, soil biodiversity contributes to food and
fibre production, and is an important regulator of other soil services
including greenhouse gas emissions, water purification (Table 5; Bodelier,
2011), and supporting services such as nutrient cycling. Stocks of soil
biodiversity represent an important biological and genetic resource for
biotechnological exploitation. Previous methodological challenges in
characterizing soil biodiversity are now being overcome through the use of
molecular technologies, and currently significant progress is being made in
opening the “black box” of soil biodiversity (Allison and Martiny, 2008)
with respect to providing fundamental information on normal operating ranges
of the biodiversity under different soil, climatic, and land use scenarios.
Addressing these knowledge gaps is of fundamental importance, firstly as a
prelude to understanding wider soil processes, but also to better inform the
likely consequences of land use or climatic change on both biodiversity and
soil ecosystem services.
The development of molecular technologies has led to a surge in studies
characterizing soil biodiversity at different scales – from large landscape
scale surveys to specific, locally focused studies using manipulation, or
contrasting of specific land uses. The large-scale surveys yield the broader
picture, and conclusions are emerging identifying the importance of soil
parameters in shaping the biodiversity of soil communities (Fierer and
Jackson, 2006). In essence, the same geological, climatic, and biotic
parameters which ultimately dictate the supporting service of soil
formation are also implicated in shaping the communities of soil biota,
thus regulating the spatial structure of soil communities observed over
large areas (Griffiths et al., 2011). Locally focused experimentation
typically reveals more specific changes with respect to local land use or
climate. Most studies have focused on assessing one component of soil
diversity. Next-generation high throughput sequencing now allows the
analyses of “whole soil food webs”, permitting a thorough interrogation of
trophic and co-occurrence interaction networks. The challenge is to
consolidate both approaches at various scales in order to understand the differing
susceptibility of global soil biomes to change.
It is essential to link these new biodiversity measures to specific soil
functions in order to understand the pivotal roles of soil organisms in
mediating soil services. The development of in situ stable isotope tracer methods
(e.g. Radajewski et al., 2000) to link substrate use to the identified
active members serves to clarify the physiological activity of these
organisms. Additionally, whole-genome shotgun metagenomic sequencing is now
becoming an increasingly cost-effective approach to assessing the
biodiversity of functional genes in soils (Fierer et al., 2013), potentially
allowing for a trait-based rather than taxon-based approach to understanding
soil biodiversity, akin to recent approaches applied to larger and more
readily functionally understood organisms above ground. It is becoming
increasingly apparent that functionality and biodiversity co-vary with other
environmental parameters. Thus manipulative experimentation is required to
determine the fundamental roles of soil biodiversity versus other co-varying
factors in driving soil functionality. Table 5 summarizes management actions
affecting the soil biota and their impacts on ecosystem services.
Management actions affecting the soil biota and their
impacts on ecosystem services.
Management action or other driver of changeProvisioning service impactRegulating service impactSupporting service impactCultural service impactLand use change of natural vegetation to agricultural intensificationChanged genetic resources; changed production of (precursors to) industrial and pharmaceutical productsDecreased C sequestration; changed pest and disease controlChanged elemental transformationChanged diversity of soil organisms (e.g. elimination of some soil organisms)Use of organic amendmentsIncreased genetic resources, decreased production of (precursors to) industrial and pharmaceutical productsIncreased C sequestrationIncreased soil formation, increased primary production by phototrophs, changed elemental transformationIncrease in soil organismsUse of broad spectrum bioactive agrochemicalsDecreased genetic resources, decreased production of (precursors to) industrial and pharmaceutical productsPossible decreased waste decomposition and detoxificationDecreased primary production by phototrophs, changed elemental transformationDecreased diversity of soil organisms (e.g. elimination of some soil organisms)Pollution by heavy metals or xenobioticsDecreased genetic resources, decreased production of (precursors to) industrial and pharmaceutical productsPossible decreased waste decomposition and detoxificationDecreased primary production by phototrophs, changed elemental transformationDecreased diversity of soil organisms (e.g. elimination of some soil organisms)Climate change (global warming)Possible decreased C sequestrationChanged elemental transformationKnowledge gaps and research needs concerning soil carbon, nutrient, and water
cycles, and the role of soil biodiversity
Soil carbon cycle: Substantial progress has been made in recent years towards more
fundamental understanding of the processes controlling soil C storage and in
improving and deploying predictive models of soil C dynamics that can guide
decision makers and inform policy. However, it is equally true that many new
(and some old) gaps in our knowledge have been identified and research needs
articulation. New research on soil C dynamics has been driven in part by
increasing awareness of (1) the importance of small-scale variability for
microbial C turnover (Vogel et al., 2014), (2) interactions between the C
cycle with other biogeochemical cycles (Gärdenäs et al., 2011), and
(3) the importance of soil C, not only at the field scale but also at regional
to global scales (Todd-Brown et al., 2013). The most cited gaps in basic
knowledge include plant effects on SOM storage and turnover; controls on
microbial efficiency of organic matter processing, including biodiversity,
association/separation of organic matter, and decomposing microbial
communities in the mineral soil matrix (Bardgett et al., 2008); the role of soil
fauna in controlling carbon storage and cycling, dynamics of dissolved
organic carbon, and its role in determining C storage and decomposition
(Moore et al., 2031; Butman et al., 2014); black C stabilization and
interactions of black C including biochar with native soil C and mineral
nutrients; and the role of soil erosion in the global C cycle (Quinton et
al., 2010). For predictive modelling and assessment, the most frequently cited
knowledge gaps are closer correspondence of measured and modelled SOM
fractions (Zimmermann et al., 2007), improved modelling of C in subsurface
soil layers, distributed soil C observational and monitoring networks for
model validation, more realistic and spatially resolved representation of
soil C in global-scale models, and the response to climatic extremes
(Reichstein et al., 2013).
Soil nutrient cycles: In the second half of the 20th century, higher biomass yields were
supported by higher use of fertilizer (N, P) inputs. Today, at the beginning
of the 21st century, this is not considered sustainable. Alternatives are
needed that will use inherent soil fertility and improved resource use
efficiencies, and to prevent losses of N and P. Examples in agriculture
include ecological intensification and new crop varieties with improved
ability to extract P and use from soils. At the food system level, more
effective nutrient management would benefit from a focus on a “5R
strategy”: (1) realign P and N inputs, (2) reduce P and N losses to
minimize eutrophication impacts, (3) recycle the P and N in bio-resources,
(4) recover P (and N) from wastes into fertilizer, and (5) redefine
use and use efficiency of N and P in the food chain including
diets and regional and spatial variability (e.g. Snyder et al., 2014).
Soil water: The soil management practices that maintain the ecosystem services of food
and water provision, flow regulation, water purification, and aesthetic
value within the soil and water cycle are well known. However, their
application is not universal and poor management leads to a loss of
function. Under scenarios of increased climatic variability with more
extremes of precipitation and increased severity of droughts, soil functions
will be stressed and the level of good soil management will be required to
improve (Walthall et al., 2012). Research into these interactions, as well as
future proofing of current good practice, is required.
Soil biota: Despite recent advances in knowledge regarding stocks and changes in soil
biodiversity, global-scale syntheses are still largely absent. Indeed, many
of these highly pertinent issues were raised more than 20 years ago
(Furusaka, 1993), and to date none of these factors have been unravelled
fully. Key barriers to syntheses are the lack of concerted soil surveys
addressing multiple functions with standardized methodologies. New
technologies for soil biodiversity assessment generate large data sets of
gene sequences which are typically archived in publicly accessible
databases. The adoption of such approaches for soil function measurements
alongside deployment of agreed standard operating procedures (e.g. as
developed in the recent, EU-funded EcoFINDERS project) could serve to
address these gaps. Ultimately, new methods are revealing the high
sensitivity of change in soil biological and genetic resources from threats
such as management, and we now need to recognize the distinct types of
organisms found in different soils globally and understand their functional
roles in order to predict vulnerability of these resources to future change.
Recommendations for management activities to support the continued delivery
of ecosystem services from soils
Best management practices that support one facet of soil functioning tend to
also support others. Building SOM, for example, enhances soil C, soil
nutrient status, improves water holding capacity, and supports soil biota
(Lal, 2004; Smith, 2012). Similarly, preservation of natural ecosystems, and
prevention of degradation or conversion to intensive agriculture, almost
always benefits soil C, nutrients, water, and biota. These synergies, and the
fundamental role of soil, make the goal of supporting soil function more
straightforward than the goal of maximizing multiple ecosystem services,
which often involves trade-offs (Robinson et al., 2013; Smith et al., 2013).
For example, in terms of the provisioning service of food, the highest
per-area yields are often obtained under intensive cropping, with large
external inputs of mineral fertilizer, other agro-chemicals (such a
pesticides and herbicides), and sometimes water through irrigation (West et
al., 2014), with the most intensive forms of agriculture occurring in
greenhouses, where external inputs of fertilizers, water, and energy can be
extremely high (Liu et al., 2008). Though intensive cropping produces high
per-area yields, it is not the best management system for a range of other
ecosystem services, potentially adversely affecting supporting services
(e.g. soil formation through erosion), regulating services (e.g. climate
regulation through greenhouse gas emissions; air, water, and soil quality
through leaching of agrochemicals; pollination through adverse impacts on
pollinators), and cultural services (e.g. reduced aesthetic value of the
landscape through large-scale monoculture; Smith et al., 2013). Balancing
the trade-offs between different ecosystems services is, therefore, more
difficult than designing management strategies that support soil C,
nutrients, water, and biota. Tables 1, 2, 4, and 5 present some examples of
management activities that affect a range of soil functions, and a number of
beneficial management actions occur in most/all of the tables. The most
important of these beneficial management activities are described below.
Land cover and use change
A number of meta-analyses (Wei et al., 2014; Guo and Gifford, 2002; Don et
al., 2011) show that natural systems lose carbon when converted to
agriculture, with the exception of forest to pasture conversion, where some
studies indicate carbon gain (Guo and Gifford, 2002) while others indicate
carbon loss (Don et al., 2011). Given the link between organic matter and
soil carbon, nutrients, water, and biota, conversion of natural systems to
agriculture is likely to adversely impact all of these factors. Protection
of natural ecosystems, therefore, benefits soil carbon, nutrients, water, and
biota. Rewilding of surplus agricultural land would be expected to enhance
soil carbon, nutrients, water, and biota, as seen for set-aside land or
reforestation of former cropland (Don et al., 2011). In the absence of land
cover/land use change, improved management of agricultural soils can
improve soil carbon, nutrient, water, and biota (Smith et al., 2015), as
described below.
Improved agricultural management
Reducing soil disturbance (e.g. through reduced or zero-tillage) is often done to improve soil
moisture retention to enhance soil function, and can also increase SOC stocks
(West and Post, 2002; Ogle et al., 2005), though the C benefits of no-till
may be limited to the top 30 cm of soil and some authors argue that the C
benefits have been overstated (Powlson et al., 2014). Baker et al. (2006)
found similar soil C in conventional and no-till systems, suggesting that C
accumulation is occurring at different depths in the soil profile under
different management schemes. Given the tight coupling of soil C and N,
increased organic matter also tends to increase nutrient supply, and also
enhances water holding capacity (Lal, 2004) which in turn improves the
delivery of ecosystem services, and can increase soil biota. Zero tillage
also gives rise to greater earthworm and arthropod populations (House and
Parmelee, 1985). Perennial crops also reduce the need for annual tillage,
and can provide similar benefits. Cultivation of perennial plants with
improved rooting systems is likely to increase soil C stocks in C-depleted
subsoil horizons (Kell, 2012). Land use change, such as removal of perennial
plants and subsequent cultivation, were found to affect both short-lived and
long-lived C pools (Beniston et al., 2014).
Maintaining ground cover through improved residue management, and use of cover crops during
traditional bare fallow periods, helps to improve C returns to the soil,
prevent erosion and surface sealing, maintain soil nutrients and soil
moisture, and support an active level of soil biota (Lal, 1997). Similar
benefits can be achieved through well-designed rotations and use of
perennial crops or agroforestry (e.g. Mbow et al., 2014).
Use of organic amendments increases SOM content (Lal, 2004; Smith, 2012; Gattinger et al., 2012),
which, as described above, benefits soil C, nutrients, water, and biota.
Organic amendments traditionally include crop residues, animal manures,
slurries, and composts. These organic matter additions were found to improve
C storage and other regulating ecosystem services if repeated regularly.
Recent developments, such as the use of biochar or hydrochar from the
pyrolysis or hydrothermal carbonization of crop residues or other biomass,
can increase SOC stocks and can also reduce soil N2O emissions and
enhance soil fertility (Zhang et al., 2010), which could be effective over
multiple years (Liu et al., 2014). However, the properties of these
materials and their net effect on ecosystem services are strongly dependent
on production conditions (Wiedner et al., 2013; Naisse et al., 2015). Soil
amendment with compost and biochar or their mixture may be particularly
useful for increasing the regulating and supporting services of degraded
soils (Ngo et al., 2014). Biochar, in conjunction with bioenergy production,
is at this stage one of the most promising technologies for achieving the
large-scale negative carbon emissions required by the middle of the century to prevent
global mean temperatures from increasing above 2 ∘C, though this
is controversial (Fuss et al., 2014).
Optimized timing and rate of fertilizer application: Intensification has increased annual global flows of N and P to more than
double natural levels (Matson et al., 1997; Smil, 2000; Tilman et al.,
2002). In China, N inputs to agriculture in the 2000s were twice that in
1980s (State Bureau of Statistics-China, 2005). Optimizing the timing and
rate of fertilizer applications ensures that the nutrients are available in
the soil at a time when the plant is able to take them up, which limits
nutrient loss, hence reducing the risk of water pollution and downstream
eutrophication (Carpenter et al., 1998). Fertilizer decision support tools
can help in implementing optimized nutrient management, as can soil testing (to
establish soil nutrient status before fertilization), and precision farming,
to ensure that nutrient additions are targeted where needed. Subsurface
application of slurries to reduce ammonia volatilization can increase
nitrous oxide emissions, so there can be trade-offs associated with this
practice (Sutton et al., 2007).
Optimized use of agrochemicals: Reduction in use of broad spectrum bioactive agrochemicals will benefit soil
biota. The under-application of pesticides and herbicides could also
plausibly have net negative environmental impact, if it means that more land
needs to be brought into production (Carlton et al., 2010, 2012).
Optimization of agrochemical applications will also reduce water pollution
through leaching.
Water management: Irrigation of dryland agriculture can increase productivity and C returns
to the soil, with the benefits to soil carbon, nutrients, water, and biota
discussed above, but it can decrease filtration potential and increase the
risk of soil salinization (Ghassemi et al., 1995; Setia et al., 2011). In
waterlogged marginal lands, drainage can increase productivity and thereby
increase carbon returns to the soil while at the same time decreasing methane and nitrous
oxide emissions. If wetland soils are drained, oxidation of organic soils
will lead to large losses of soil C and the nutrients associated with
it, decreasing the ability of these soils to carry out services like water
purification (e.g. through denitrification). Drainage of peatlands has been
associated with increased runoff and flood risk (Ballard et al., 2012). In
terms of biodiversity, productivity of drained marginal lands can increase
at the expense of plant genetic diversity.
Improved grazing management (e.g. optimized stocking density) can reduce soil degradation and thereby
maintain and enhance organic matter content (McSherry and Ritchie, 2013),
benefiting soil C, nutrients, water, and biota as described above. Higher
productivity and deep-rooted grasses can similar effects (Kell, 2012), while
also modifying water use efficiency, but potentially at the expense of plant
genetic diversity. Reduction in grazing density can reduce soil compaction
and therefore increase infiltration and water storage and reduce the risk of
runoff and flooding downstream (Marshall et al., 2009). Fire management can
also increase soil C and nutrient status of soils (e.g. Certini, 2005).
Conclusions
Many practices are known to enhance all or most of the functions of soils
considered in this review, which is encouraging for our efforts to protect
soils into the future. Soils are complex, there are still knowledge gaps
(outlined in Sect. 6), and fundamental research is still needed to better
understand the relationships between different facets of soils and the array
of ecosystem services they underpin. There is a tendency to dwell on the
complexity and knowledge gaps rather than to focus on what we do know and
how this knowledge can be put to use to improve the delivery of ecosystem
services. While more knowledge is required on where specific agricultural
systems are best placed to utilize and deliver ecosystem services most
efficiently in order to protect and enhance our soils in the long term, best
practices are well characterized and many can be implemented immediately.
Despite a growing population and increasing demands for resources, enough is
known to discriminate the extremes of beneficial and detrimental
agricultural practices, as well as their interactions with different types of
soils. A significant challenge is to find effective ways to share this
knowledge with soil managers and policy makers, so that best management can
be implemented. A key element of this knowledge exchange must be in raising
awareness of the ecosystems services underpinned by soils and thus the
natural capital they provide (Robinson et al., 2013). We know enough to
start moving in the right direction, while we conduct research to fill in
our knowledge gaps. Therefore, a challenge to soil scientists is to better
communicate what we do know while we carry out research to better
understand the things that we do not know. The lasting legacy of the
International Year of Soils in 2015 should be for soil scientists to work
together with policy makers and land managers in order to put soils at the centre of
environmental policy making and land management decisions.
Acknowledgements
The input from P. Smith and P. J. Kuikman contributes to the EU-funded FP7 project
SmartSOIL (grant agreement no. 289694), and that from P. Smith and R. I. Griffiths to the
NERC-funded U-Grass project (NE/M016900/1). The input from P. C. West and J. S. Gerber was
supported by the Gordon and Betty Moore Foundation, and that from P. Smith, J. S. Gerber, and
P. C. West contributes to the Belmont Forum/FACCE-JPI-funded DEVIL project
(NE/M021327/1). Input from G. Pan was supported by funding from the Priority
Academic Program Development of Jiangsu Higher Education Institutions,
China. J. I. House was funded by a Leverhulme early career research fellowship.
Edited by: K. Kalbitz
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