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Agriculture, Ecosystems and Environment 337 (2022) 108048
Available online 30 June 2022
0167-8809/© 2022 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).
Environmental drivers for riparian restoration success and ecosystem
services supply in Mediterranean agricultural landscapes
Clara Castellano
a
,
*
, Daniel Bruno
a
, Francisco A. Comín
a
, Jos´
e M. Rey Benayas
b
,
c
, Adri`
a Masip
d
,
Juan J. Jim´
enez
a
a
Instituto Pirenaico de Ecología (IPE), CSIC, Avda. Monta˜
nana 1005, 50059 Zaragoza/ Avda. Ntra Sra Victoria, 16, 22700 Jaca, Spain
b
Life Sciences Department, Alcala University, E-28805 Alcal´
a de Henares, Spain
c
Fundaci´
on Internacional para la Restauraci´
on de Ecosistemas, Princesa 3 dpdo., Apto. 703, E-28008 Madrid, Spain
d
Burgeon EnvironMental Solutions, C. Major 30, 25457 El Vilosell, Spain
ARTICLE INFO
Keywords:
Carbon storage
Habitat provision
Revegetation
Riparian forest
Microclimate regulation
Water purication
ABSTRACT
Riparian forests nestled in agricultural landscapes represent a small proportion in crop-intensive areas, while
contributing remarkably to their biodiversity. This biodiversity supports several ecological processes crucially
involved in the supply of ecosystem services (ES) complementary to that provided by agricultural lands and also
relevant for designing biodiverse and multifunctional landscapes. Riparian forest is one of the most threatened
ecosystems due to land-use intensication and associated water extraction, especially in Mediterranean semi-arid
areas, and proper evaluation of the success of riparian restoration projects is usually lacking. Furthermore, there
is little empirical evidence of the effects of riparian restoration on ES supply. In this study, we rst investigated
the effect of hydrological and soil features on survival and growth of saplings planted in degraded riparian areas
in two Mediterranean watersheds. Then, we evaluated how riparian restoration affected the supply of ES,
comparing nine regulating and supporting ES on these restored areas with other riparian areas spanning a
gradient of conservation status, and with other natural and agricultural land-uses in the same watershed.
We found that restoration success mainly depended on water table depth, soil salinity and soil nutrients
(namely Mg
+2
and Olsen P). Moreover, we detected an antagonistic interaction between the latter two, and a
synergetic interaction between water table depth and soil salinity. Forest patches provided meaningful regulating
and supporting ES in agricultural landscapes. In particular, riparian restoration zones increased the supply of
regulating and supporting ES (water purication, habitat provision, microclimate regulation and soil C storage)
in comparison with degraded natural land-uses and crops. Nevertheless, they were still far from the magnitude
and range of ES provided by mature riparian forests. These results highlight the importance of focusing man-
agement practices on conserving riparian forest patches and restoring the degraded ones to reconcile agricultural
production with the maintenance or enhancement of ES in agricultural Mediterranean landscapes.
1. Introduction
About 40 % of terrestrial natural ecosystems on Earth have been
transformed into agricultural land (FAO, 2018). This vast land-use
change has increased the supply of provisioning ecosystem services
(ES) such as food and bre (Rey Benayas and Bullock, 2012), whereas it
has reduced supporting, regulating and cultural ES (Bullock et al., 2011;
Landis, 2017). In fact, landscape transformation due to agricultural
expansion and intensication is the most important driver of biodiver-
sity loss and ES decline worldwide (Tscharntke et al., 2012).
Importantly, this could be aggravated by the projected population
growth for the next decades and the associated increase in global food
demand (Bruinsma, 2009).
Within land-use changes, agricultural intensication and the de-
mand of land for agriculture is one of the main drivers of habitat loss and
the reduction of spatial heterogeneity (Gavier-Pizarro et al., 2012). This
intensication inuences ES supply due to mismatches between agri-
cultural uses and regulating and supporting ES (Hasan et al., 2020).
Agricultural intensication increases soil degradation, soil erosion, loss
of organic matter content and nutrient leaching, which reduce the
* Corresponding author.
E-mail address: clarac@ipe.csic (C. Castellano).
Contents lists available at ScienceDirect
Agriculture, Ecosystems and Environment
journal homepage: www.elsevier.com/locate/agee
https://doi.org/10.1016/j.agee.2022.108048
Received 15 December 2021; Received in revised form 29 April 2022; Accepted 29 May 2022
Agriculture, Ecosystems and Environment 337 (2022) 108048
2
supply of relevant ES as water supply, water quality, water inltration
and soil water holding capacity (Sharma et al., 2011). Tree removal
during this process affects microclimate regulation by accelerating
evaporation from soil, resulting in a drier and hotter microclimate (Pires
and Costa, 2013; Sampaio et al., 2007). Likewise, deforestation has
detrimental effects on habitat provision, since the capacity to support a
diversity of wildlife is linked to structural vegetation diversity and tree
dominance or coverage (Berg, 1997; Breuste et al., 2013). Although
agricultural practice also inuences soil microbial activity and soil
carbon storage (Janvier et al., 2007; Wu et al., 2003), these changes can
be reverted through restoration actions. For instance, Cheng et al.
(2013) found that the afforestation of agriculture lands increased
considerably soil organic C in response to the increased below ground
biomass.
Agricultural land-use is also a main driver of riparian forest frag-
mentation (Fernandes et al., 2011). Although riparian forest remnants
only represent a small proportion of crop-intensive areas, they
contribute remarkably to the biodiversity of agricultural landscapes
(Sabo et al., 2005). These species-rich ecosystems support numerous
ecological functions and processes related to the supply of ES which are
essential for agriculture production (Cole et al., 2020) and human
welfare (Felipe-Lucia and Comín, 2015; Souza et al., 2013). Riparian
forests supply key provisioning (e.g., biomass and genetic resources),
supporting (e.g., maintenance of nursery population and habitats),
regulating (e.g., carbon sequestration) and cultural (e.g., recreation and
aesthetic value) ES (Riis et al., 2020), which are complementary to those
supplied by agricultural lands. In particular, riparian vegetation can act
as a buffer of diffuse pollution to rivers (Antiguedad et al., 2017; Collins
et al., 2009; Turunen et al., 2019; Welsch, 1991). They contribute to
water purication by ltering sediments, nutrients, and pesticides
entering from upland agricultural elds to the river channel (Anbu-
mozhi et al., 2005), an ES of pivotal importance for the maintenance and
improvement of water quality in agricultural landscapes (Schultz et al.,
2004). Riparian forests also play an important role as refuges for fauna
(Woinarski et al., 2000), promoting biodiversity conservation (Lees and
Peres, 2008). They also provide key regulating ES for agriculture itself,
such as pest control (Landis et al., 2000), pollination and ood protec-
tion (Cole et al., 2020). However, despite their high ecological impor-
tance, riparian forests are one of the most threatened ecosystems due to
land-use intensication and associated water extraction, especially in
Mediterranean semi-arid areas (Bruno et al., 2014; Salinas et al., 2000).
A wide range of riparian restoration projects have attempted to
enhance riparian ES and mitigate impacts of agriculture (Turunen et al.,
2019). Their outcomes depend on the goals, previous conservation sta-
tus, social acceptance and techniques used (Gonz´
alez et al., 2017). Ri-
parian restoration is particularly important in semi-arid areas
(Felipe-Lucia and Comín, 2015) where broadleaf deciduous forests are
usually limited to watercourses (Bruno, 2015). Riparian restoration
focused on revegetating degraded areas allows forest structure to
quickly recover and, although it may not become a genuine riparian
forest, high levels of complexity, diversity and functionality may be
reached (Bourgeois et al., 2016; Holl and Aide, 2011; Vasilopoulos et al.,
2007).
In the Mediterranean context, revegetation success is usually
conditioned by the suitability and adaptation of plant species to survive
under multi-stress conditions due to natural (e.g., drought and high
salinity) and anthropogenic stressors (e.g., agricultural intensication
and ow regime alteration). The identication of the main stressors that
determine plant survival and growth can help managers in the selection
of the most suitable species and river reaches for cost-effective riparian
restoration in semi-arid Mediterranean areas. Among the potential
drivers of restoration success, groundwater depth and soil properties
seem to have a prominent role in plant survival and growth (Lu et al.,
2002; Richardson et al., 2007; Gonz´
alez et al., 2010; Souter et al., 2014;
Solari et al., 2016), which could subsequently inuence the range and
magnitude of ES provided by riparian areas. In addition, we lack a deep
understanding on how these drivers interact in water-limited water-
sheds, which are expected to increase due to current climate change
(Feng and Fu, 2013).
Restoration of degraded riparian areas has experienced a boost in the
last few decades (Szałkiewicz et al., 2018). However, proper evaluation
of restoration success is usually lacking (Wortley et al., 2013; Gonz´
alez
et al., 2015). Furthermore, there is little empirical evidence of the effects
of riparian restoration on ES supply (Chazdon, 2008; Symmank et al.,
2020). Thus, the objectives of this study were: (i) to identify key envi-
ronmental drivers of revegetation success of degraded riparian areas,
and their potential interaction, in Mediterranean watersheds with a
predominant agricultural land-use, and (ii) to evaluate how riparian
restoration affects the supply of regulating and supporting ES compared
to other natural and agricultural land-uses. To accomplish the rst
objective, we investigated the effect of hydrological and soil features on
survival and growth of saplings of ten riparian native species that were
planted in degraded Mediterranean riparian areas. To meet the second
objective, we measured seven regulating (water purication, microcli-
mate regulation, soil C storage, storage of soil organic pollutants, water
holding capacity, soil water inltration rate and mitigation of surface
runoff) and two supporting (habitat provision and soil formation) ES in
riparian areas with different conservation status (mature reference for-
est, restored and degraded areas), and compared them with those pro-
vided by other natural and agricultural land-uses in the same watershed.
We hypothesized that (i) riparian revegetation success in agricultural
Mediterranean areas can be predicted by water table depth and soil
properties (Solari et al., 2016; Zhang et al., 2018) and the interactions
among them, ultimately affecting the ES provided by restored riparian
areas. Then, (ii) mature riparian forest and restored riparian zones could
supply higher levels of regulating and supporting ES than the degraded
ones (Chazdon, 2008; Symmank et al., 2020), therefore complementing
those provided by other natural or semi-natural land-uses and crops at
the landscape scale. We aim to bring out the value of riparian forest for
the supply of supporting and regulating ES in agricultural landscapes, as
well as gaining insight into how to maximize the success of riparian
restoration projects that pursue to achieve a balanced ES supply in
agricultural Mediterranean landscapes.
2. Material and methods
2.1. Study area
The study area comprises two Mediterranean watersheds nestled in
the Flumen River watershed (North Monegros County, NE Iberian
Peninsula; Fig. 1). The smaller watershed (San Juan del Flumen, here-
after San Juan) covers an area of 171.2 ha and the larger one (Lalueza) of
994.8 ha. Both watersheds (<10 km between them) drain to the middle
segment of the Flumen river. They have a semi-arid climate character-
ized by an average annual precipitation of 434 mm and average
maximum and minimum temperatures of 24.7 ºC and 3.8 ºC, respec-
tively (Pedrocchi, 1998). Soils have a pH ranging from slightly alkaline
to very alkaline (Navas, 1998), with materials from the lower-middle
Miocene, and from the Pleistocene to the Holocene, mainly consisting
of conglomerates, sandstones, shales, limestone, marl, gypsum, gravel,
sand, silts and clays (IGME, 2015). Potential riparian communities
correspond to habitat 92A0-Salix alba and Populus alba galleries (Habitat
Directive, 1992), but river oodplains have been intensively trans-
formed for agricultural production. Due to water scarcity, the predom-
inant crops were dryland cereals, mainly wheat and barley, until the
1940s. However, after the construction of the Monegros and Flumen
Channels (1915–1982) there was a wide intensication and trans-
formation into irrigated crops. The current proportion of agricultural
land-use in the study area is 72.6 % (San Juan) and 86.1 % (Lalueza) of
their total watershed areas, mostly irrigated crops (Table A.1).
C. Castellano et al.
Agriculture, Ecosystems and Environment 337 (2022) 108048
3
2.2. Data collection
2.2.1. Monitoring of riparian revegetation
Riparian restoration actions started simultaneously in 2011 in the
framework of project CREAMAgua (LIFE09 ENV/ES/000431). They
comprised of slope re-proling, where feasible, and revegetation in
2011, and replanting in 2012 (restored riparian zone in Fig. 1). We
monitored a total of 436 planted saplings of 10 species (Populus alba,
Populus nigra, Fraxinus angustifolia, Rosa canina, Celtis australis, Salix
alba, Salix atrocinerea, Ulmus minor, Lonicera implexa and Phillyrea
angustifolia) corresponding to plots where water table was monitored
(red boxes in Fig. 1). We measured four indicators of riparian revege-
tation success: survival rate of planted individuals, percentage of living
branches (i.e., with alive leaves or buds), height growth, and diameter
growth (i.e., diameter at breast height). Monitoring of survival and
percentage of living branches was performed twice a year (June and
September-October, i.e., at the beginning and end of plant growth
period, respectively), starting right after the restoration actions and up
to ve years later (2016). Height and diameter growth were incorpo-
rated in 2015, when riparian plants were completely established
(Table A.2).
Additionally, we assessed differences in species composition and
richness among the whole range of conservation status in riparian zones.
For this, we identied spontaneous vegetation (woody and perennial
herbs) and planted species in plots regularly distributed across the
restored riparian zone, the mature riparian forest (red dots in riparian
areas in Fig. 1) and across a degraded riparian area located between the
two studied watersheds (see Appendix B for detailed information on
data collection).
2.2.2. Environmental drivers of revegetation success
To explain plant survival and growth, we measured the water table
depth in riparian zones and a set of seven soil properties: percentage of
clay, soil salinity (measured as electrical conductivity), bulk density and
soil concentration of total N, Olsen P, Mg
+2
and K
+
(Fig. 1 and Table A.3;
see Appendix B for detailed information on data collection and labora-
tory analysis). We performed spatial data interpolation within restored
riparian zones by applying Inverse Distance Weighting (IDW) and
Kriging techniques in ArcGIS 10.3 (ESRI, 2016) based on soil and water
sampling sites, respectively, to assign a value of each variable to every
monitored individual sapling.
2.2.3. Supply of regulating and supporting ES across land-use types
2.2.3.1. Land-use types. We used the Spanish crop and land-use digital
map (MCA, 2009) in ArcGIS 10.3 (ESRI, 2016) and eld observations to
identify nine different land-uses: ve semi-natural (conifer forest,
mature riparian forest, restored riparian zone, shrubland, and wetland)
Fig. 1. Maps of San Juan and Lalueza watersheds showing their location within the Flumen watershed and Iberian Peninsula, their different land-uses, and the
location of sampling plots for environmental measurements (red dots), temperature and humidity sensors (green dots), piezometers (orange stars) and monitoring of
planted species (436 saplings within red boxes). Right panels show riparian areas in detail. (For interpretation of the references to colour in this gure, the reader is
referred to the web version of this article.)
C. Castellano et al.
Agriculture, Ecosystems and Environment 337 (2022) 108048
4
and four agricultural land-uses (rotation crop, rice crop, olive grove, and
fallow crop; Fig. 1 and Table A.1). Rotation crops were irrigated areas
that produce different yields during the same year (especially corn and
barley, but sometimes onion or rye grass) or, in the case of alfalfa, after
3–5 years. The conifer forests were reforested pine groves from the late
70 s, characterized by monospecic plantations of Pinus halepensis, but
also by the presence of some aromatic plants such as Thymus vulgaris.
Species such as Artemisia herba-alba and Salsola vermiculata, often
accompanied by some sub-halophilic species, such as Atriplex halimus,
predominate in the shrubland. Given that we looked for differences in
the ability of different conservation status to supply regulating and
supporting ES, we split the restored riparian zone into (1) restored ri-
parian forest (revegetation that reached >30 % tree cover) and (2)
restored riparian area (revegetation that reached <30 % tree cover with
shrubby dominant strata; Fig. C.1).
2.2.3.2. Ecosystem services. To assess the relationship between land-use
types and ES, we identied the nine regulating and supporting ES (see
above; Table 1) across different land-use types in San Juan and Lalueza
(Table A.4; see Appendix B for detailed information on data collection
and laboratory analysis). To cover the whole range of conservation
status of riparian zones, and therefore complement the mature and
restored ones, we also measured ES in six sampling plots across
degraded riparian areas. Water purication, measured as the retention
of nutrients from upland crops to the river edge, was evaluated only in
riparian zones. For this, we measured groundwater nutrients recurrently
(four times) along a spatial gradient from adjacent croplands to the
riverbank (see Appendix B for detailed methods).
2.3. Data analysis
2.3.1. Environmental drivers of riparian revegetation success
Firstly, to detect differences in the taxonomic composition of mature,
restored and degraded riparian areas, we performed a Non-Metric
Multidimensional Scaling (NMDS) on occurrence data. In addition,
Kruskal-Wallis test was used with post-hoc paired comparisons (Dunn
test) to assess if there were differences in riparian vegetation richness
(woody and perennial herbs) between the different conservation sta-
tuses of riparian zones.
To assess the environmental drivers of riparian revegetation success,
we used Pearson R correlations to identify highly correlated soil vari-
ables (r >0.85) in our dataset, discarding those with greater number of
signicant correlations for subsequent analysis (i.e., bulk density and
K
+
; variables with an asterisk in Table A.3) to avoid redundancy and
collinearity. Thus, the selected set of variables for further analysis were
water table depth, percentage of clay, soil electrical conductivity, and
soil concentration of total N, Olsen P, and Mg
+2
(Fig. C.2).
We used generalized linear mixed models (GLMM) to test the effects
of these six predictor variables on four characteristics of riparian
revegetation success: survival of planted individuals, percentage of
living branches, height growth and diameter growth. Species identity
were included as a random effect in the models to account for species-
dependency. We used the R package glmulti to automate model selec-
tion and identify the best one, based on the Bayesian Information Cri-
terion (BIC) and considering two-way interactions between all variables
(Calcagno and de Mazancourt, 2010). Prior to this analysis, all variables
were logarithmic transformed and standardised to ensure normal error
distributions and homogenous variance. We also used the percentage of
explained deviance to evaluate model performance.
2.3.2. Supply of regulating and supporting ES across land-use types
To evaluate water purication, we used GLMM to assess the temporal
and spatial changes in each nutrient concentration in groundwater from
upland crops to the river edge (i.e. nitrate, phosphate and sulphate
sorption by riparian areas). We considered collection date and distance
to the river channel (and the interaction between them) as xed factors
and site as a random factor. The differences between mature riparian
forest and restored riparian zones were accounted by replicating GLMM
incorporating the type of riparian zone (restored or mature) as another
xed factor.
We assessed the differences in microclimate regulation between
land-uses and associated vegetation structure with ANOVAs and Tukey
post-hoc comparisons. Habitat provision was also evaluated by ANOVAs
and, due to the violation of parametric test assumptions, we performed
the non-parametric Kruskal-Wallis and Dunn post-hoc tests.
Finally, we evaluated land-use differences in the supply of soil
related-ES (i.e. soil C storage, soil formation, storage capacity of organic
pollutants, water holding capacity, water inltration rate, and mitiga-
tion of surface runoff) with GLMM (land-use as xed factor and site as
random factor). Likelihood ratio tests were implemented to compare
these models (xed and random effects) with null models (only random
effects) and detect model signicance for the different ES (Table 1). We
evaluated goodness of t with marginal R
2
associated to xed effects and
conditional R
2
associated to xed and random effects (Nakagawa and
Schielzeth, 2013). If applicable, pairwise post-hoc comparisons were run
to identify meaningful responses.
We visually checked homoscedasticity and normality of residuals in
all GLMM and ANOVAs. Phosphate and sulphate groundwater concen-
trations, organic C, microbial biomass, cation exchange capacity (CEC),
hydraulic conductivity and sorptivity were previously log-transformed
to meet model assumptions. To perform statistical analyses we used
“dunn.test” (function dunn.test; Dinno and Dinno, 2017), “effects”
(alleffects; Fox et al., 2019), “ggplot2” (aes and ggplot; Wickham et al.,
2016), lme4” (lmer; Bates et al., 2015), “lmerTest” (lmer; Kuznetsova
et al., 2015), “multcomp” (glht; Hothorn et al., 2008), “MuMIn” (r.
squaredGLMM; Barton, 2009), “nlme” (lme; Pinheiro and Bates, 2006),
“vegan” (metaMDS, Oksanen et al., 2020) and “stats” (aov, Kruskal.test,
TukeyHSD) libraries in R statistical software version 3.6.1 (R Core Team,
2019).
Table 1
List of the nine studied regulating and supporting ecosystem services. See Ap-
pendix B for detailed methods in data collection and laboratory analysis.
MEA
category
a
Ecosystem service Indicator Unit
Supporting Habitat provision Tree cover %
Number of
vegetation strata
–
Supporting Soil formation Soil microbial
biomass
Kg C m
2
Regulating Water purication NO
3
-
in groundwater mg l
-1
SO
4
2-
in groundwater mg l
-1
PO
4
3-
in groundwater mg l
-1
Regulating Soil C storage Soil organic C Mg ha
-1
Regulating Soil storage capacity of
organic pollutants
Soil CEC cmol
c
dm
-2
Regulating Soil water holding
capacity
Saturated soil water
content
cm
3
cm
-3
Regulating Soil water inltration
rate
Hydraulic
conductivity of
topsoil
mm s
-1
Regulating Mitigation of surface
runoff
Sorptivity of topsoil mm s
-0.5
Regulating Microclimate regulation Inverse DTR adimensional
Inverse DHR adimensional
Mean daily air Tª ◦C
Mean daily air H %
Abbreviations: CEC=cation exchange capacity; DTR=daily temperature range
(mean of the difference between daily maximum and minimum temperature);
DHR=daily humidity range (mean of the difference between daily maximum
and minimum humidity); H=humidity
a
Millennium Ecosystem Assessment category (MEA, 2005)
C. Castellano et al.
Agriculture, Ecosystems and Environment 337 (2022) 108048
5
3. Results
3.1. Environmental drivers of riparian revegetation success
According to NMDS results, mature, restored and degraded riparian
areas differed in the taxonomic composition of riparian assemblages
(stress =0.14; Fig. C.3a). A gradient from mature to degraded riparian
areas in terms of species composition was found. Vegetation richness
also differed between riparian zones (p-value <0.001). Mature and
restored riparian zones showed a higher vegetation richness than
degraded riparian areas (Fig. C.3b, p-value <0.001).
The survival rates of six species were >50 %, namely F. angustifolia,
R. canina, S. alba, S. atrocinerea, U. minor and P. angustifolia (Table A.2).
In particular, F. angustifolia and R. canina were the species with highest
survival rates (83 % and 81.4 % of planted individuals, respectively),
while P. nigra and P. alba showed the lowest ones (13.9 % and 24.4 %,
respectively). F. angustifolia and R. canina exhibited the highest per-
centages of living branches (75.3% and 67.4 %, respectively), whereas
the lowest were observed for P. nigra and P. alba (13.2 % and 20.2 %,
respectively). S. alba and S. atrocinerea were the species with greater
growth in both diameter and height (19.72 cm and 15.29 cm in diam-
eter, and 103.63 cm and 70.63 cm in height, respectively), whereas
P. angustifolia and C. australis presented the lowest growth (0.53 cm and
0.36 cm in diameter, and 3.85 cm and 4.47 cm in height, respectively;
Table A.2).
Regarding the environmental drivers of riparian restoration success,
the best-tting models for the four studied indicators (i.e., survival rate,
percentage of living branches, height growth and diameter growth)
explained between 29 % and 34 % of the total variance (conditional R
2
;
Table 2). The terms included in the best models for explaining the sur-
vival rate of planted individuals, percentage of living branches and
height growth were the same: water table depth (negative relationship;
Fig. C.4), and the antagonistic interaction between Mg
+2
and soil
salinity (positive and negative relationships, respectively; Fig. C.5).
Similarly, the best model explaining diameter growth included a nega-
tive synergistic interaction between water table depth and soil salinity
(both negative relationships; Fig. C.4), and an antagonistic interaction
between Olsen P concentration and soil salinity (positive and negative
relationships, respectively; Fig. C.5). Note that although K
+
and bulk
density were not included as potential predictors in this analysis due to
high correlations with other variables considered, K
+
had a positive
correlation with Mg
+2
(0.56), and bulk density had a negative correla-
tion with Mg
+2
(−0.70; Fig. C.2).
3.2. Supply of regulating and supporting ES across land-use types
3.2.1. Water purication
Nitrate concentration in groundwater was signicantly reduced in
riparian areas (here approached as “Distance to river channel”, p-value
<0.01; Table 3, Fig. C.6). The retention rate from upland to river
channel over time (term “Date: Distance”, p-value <0.01; Fig. 2) ranged
between 51 % in summer 2015, when nitrate inputs were low (mean =
12.83 mg/l), and 79 % in summer 2016, when they were the highest
(mean =33.52 mg/l). Sulphate concentration showed a temporal
decrease (term “Date”, p-value <0.001; Fig. C.7) modulated by riparian
areas (term “Date: Distance”, p-value <0.05; Fig. C.8), which retained up
to 41 % of sulphate inputs at the end of the monitoring period (summer
2016). No signicant temporal or spatial patterns were found for
phosphate concentration. Moreover, no signicant differences were
observed between mature and restored riparian areas (p>0.05).
3.2.2. Microclimate regulation
We found that land-use types signicantly differed (p<0.001) in
terms of daily mean air temperature (R
2
=0.56, Fig. 3a), inverse daily
temperature range (DTR, R
2
=0.53, Fig. 3b), daily mean air humidity (R
2
=0.59, Fig. 3c), and inverse daily humidity range (DHR, R
2
=0.48,
Fig. 3d). Mature riparian forest and restored riparian forest showed the
lowest air temperature values and highest air humidity values. In
addition, increased inverse DTR and DHR were found in mature riparian
forest, conifer forest, restored riparian forest, olive groves and wetland
in comparison with non-arboreal land-uses. Shrubland and fallow crops
showed the highest air temperature and the lowest air humidity, inverse
DTR and DHR. Finally, rotation crops exerted an intermediate buffer
effect in all air temperature and humidity variables (Fig. 4; Table A.4
and A.6).
3.2.3. Habitat provision
As expected, land-use types signicantly differed (p<0.001) in tree
cover (R
2
=0.95, Fig. 5a) and number of vegetation strata (R
2
=0.73,
Fig. 5b). Mature riparian forest, restored riparian forest, and conifer
forest showed the highest tree cover (>70%) and number of vegetation
strata (trees, shrubs, and herbs), whereas rotation and rice crops showed
the lowest values for both variables (0% of tree cover and one single
stratum; Fig. 4; Table A.4 and A.6).
3.2.4. Soil related-ES
We found that land-use types signicantly differed (p<0.001) in
terms of soil formation (R
2
m =0.18, Fig. 6a) and soil C storage (R
2
m =
0.27, Fig. 6b). Mature riparian forest, conifer forest, and rotation crops
exhibited the highest values for both variables (also restored riparian
area for soil C storage), whereas shrubland and fallow crops exhibited
Table 2
Best-tting generalized linear mixed models (GLMMs) selected to explain the four evaluated indicators of riparian revegetation success by environmental drivers.
Explanatory variable Survival rate Living branches Height growth Diameter growth
R
2
c
=0.34
R
2
m
=0.12
R
2
c
=0.33
R
2
m
=0.14
R
2
c
=0.29
R
2
m
=0.18
R
2
c
=0.30
R
2
m
=0.21
Estimate P-value Estimate P-value Estimate P-value Estimate P-value
Water table depth -0.152 <0.001 *** -14.550 <0.001 *** -0.350 <0.001 *** -0.197 <0.001 ***
Soil EC -0.092 0.113 -10.647 0.042 * -0.372 0.001 * * -0.200 <0.001 ***
Soil Mg
2þ
0.149 0.008 * * 17.625 <0.001 *** 0.549 <0.001 *** – –
Soil P Olsen – – – – – – 0.077 0.110
Soil Mg
2
þ
: Soil EC 0.099 <0.001 *** 9.309 <0.001 *** 0.237 <0.001 *** – –
Soil P Olsen: Soil EC – – – – – – 0.185 <0.001 ***
Water table depth: Soil EC – – – – – – -0.159 <0.001 ***
Signicance codes: ‘* ** ’ p<0.001; ‘* *’ p<0.01; ‘* ’ p<0.05; ‘.’ p<0.1
Abbreviations: R
2
c
=conditional R
2
; R
2
m
=marginal R
2
.
Hyphen means that these variables were not selected in the best model for this indicator.
C. Castellano et al.
Agriculture, Ecosystems and Environment 337 (2022) 108048
6
the lowest ones (also degraded riparian area for soil formation;
Table A.4). Water inltration (R
2
m =0.14) and mitigation of surface
runoff (R
2
m =0.25) also differed among land-use types (p<0.001;
Figs. 6c and 6d). The lowest water inltration and mitigation of surface
runoff rates were found in conifer forests, while mature riparian forest,
rotation crops, restored and degraded riparian zones showed the highest
values. Finally, soil water holding capacity (R
2
m =0.11, p=0.002) and
storage capacity of organic pollutants (R
2
m =0.14; p<0.001) showed
also signicant differences among land-use types, displaying mature
riparian forest and shrubland the greatest values, respectively (Fig. 4;
Figs. 6e and 6f; Table A.4 and A.6).
4. Discussion
Our results showed that riparian revegetation was relatively suc-
cessful since most of the planted species had a survival rate over 50 %,
reaching over 80 % in F. angustifolia and R. canina. In addition, mature
and restored riparian zones were similar in terms of taxonomic
composition and vegetation richness, clearly differing from degraded
riparian areas. This success depended mainly on water table depth and
soil properties (i.e., salinity, Mg
+2
and Olsen P), which ultimately
conditioned the range and magnitude of ES provided by restored ri-
parian zones. Riparian restoration increased the supply of a large variety
of regulating and supporting ES (water purication, habitat provision,
microclimate regulation and soil C storage), complementary to that
provided by other natural or semi-natural land-uses and agricultural
crops. However, both the structure and the supply of ES of restored ri-
parian zones were still far from those of mature riparian forest. Diameter
and height growth rates of planted species, except for Salix spp., was
slow (0.80 cm/year and 5.60 cm/year on average for all studied species
except for Salix spp., respectively), as one might expect under the
environmental harshness of these Mediterranean semi-arid areas (Stella
et al., 2013). However, if the observed growth rates continue, the
planted individuals could reach sizes similar to those of mature riparian
forest in the medium term, also getting closer to their level of ES supply.
4.1. Environmental drivers of riparian revegetation success
As expected, water table depth and soil properties (i.e., soil salinity,
Mg
+2
and Olsen P) explained riparian revegetation success (as previ-
ously found by Dreesen et al., 2002, among others). Other studies also
found a negative effect of water table depth (Chen et al., 2008; Meli
et al., 2015; Orellana et al., 2012) and soil salinity (Glenn et al., 1998;
Lymbery et al., 2003) on riparian species survival and growth. However,
we additionally detected that these two stressors had a negative syner-
gistic effect on diameter growth, which had been poorly studied for ri-
parian revegetation. In a global change context where both stressors will
probably experience an increase in magnitude and extent, these ndings
could have great management implications for rivers owing through
water-limited agricultural watersheds.
We observed that Mg
+2
soil concentration showed a strong positive
relationship with most indicators of riparian revegetation success. Mg
+2
and K
+
(which showed a positive correlation with Mg
+2
) are essential
elements for chlorophyll production, photosynthetic performance, and
plant growth (Ding et al., 2006; Lu et al., 2002; Shaul, 2002). In our
study, the effect of Mg
+2
in revegetation success showed an antagonistic
and signicant interaction with soil salinity meaning that even if soil
salinity had a negative effect on revegetation success, this effect was
palliated with a high Mg
+2
concentration. Similarly, we found that the
effect of soil salinity on diameter growth was modulated (antagonistic
interaction) by Olsen P instead of Mg
+2
. These ndings support that
higher nutrient availability may overcome some inhibitory effects of
salinity, ground-truthed in crops (Khaled and Fawy, 2011; Ravikovitvh
and Porath, 1967), but little studied for riparian species. In addition, we
outline the importance of the role of bulk density, negatively related to
nutrient content (Chaudhari et al., 2013) and thus to plant growth
(Houlbrooke et al., 1997). These results provide insight for landscape
managers about the environmental drivers and the relationships among
them that they should take into account for successful riparian resto-
rations in semi-arid Mediterranean areas.
4.2. Supply of regulating and supporting ES across land-use types
4.2.1. Water purication
The studied riparian zone reduced up to 79 % and 41 % of nitrate and
sulphate concentration in groundwater, respectively. These results
demonstrate the positive impact of riparian buffer zones in mitigating
water pollution from agricultural upland elds (Anbumozhi et al., 2005;
Clausen et al., 2000), which seems essential to juggle crop production
with multifunctional landscapes. In the European context, the efciency
of inorganic nitrogen removal in riparian zones seems to be particularly
driven by nitrate inputs, whereas the inuence of vegetation type
(woody or herbaceous) can be negligible (Sabater et al., 2003).
Accordingly, it is not surprising that mature riparian forests, dominated
Table 3
Results of generalized linear mixed models (GLMMs) for water purication.
Water purication Model Distance to river channel Date Distance: Date
P-value R
2
m P-value Spatial trend P-value Temporal
trend
P-value Spatial-temporal trend
Nitrates <0.001 *** 0.43 0.006 * * +0.64 =0.008 ** + + summer 2016
Sulphates <0.001 *** 0.35 0.8 =<0.001 * ** +summer 2016 0.02 * +summer 2016
- autumn 2015
Phosphates <0.001 *** 0.22 0.11 =0.25 =0.17 =
Signicance codes: ‘*** ’ p<0.001; ‘**’ p<0.01; ‘* ’ p<0.05
Signicance symbols: ‘+’ positive, ‘+ +’ highly positive, ’-‘ negative, ‘=’ no signicant trend.
Abbreviations : R
2
m=marginal R
2
Fig. 2. Interaction effects of date and distance to river channel (p <0.01) on
groundwater nitrate concentration tested by generalized linear mixed
models (GLMM).
C. Castellano et al.
Agriculture, Ecosystems and Environment 337 (2022) 108048
7
Fig. 3. Results of generalized linear mixed models (GLMM) and post-hoc comparisons relative to the differences in (a) mean air temperature, (b) inverse daily
temperature range (DTR), (c) mean air humidity, and (d) inverse daily humidity range (DHR) among land-use types. Letters (a, b, c) depict the signicant differences
found among land-use types (Table A.5). The bold horizontal line denotes the median value, the box delimits the interquartile range, and the whisker lines extend to
the observed maximum and minimum, except for the outliers symbolized by black dots. Blue dots represent individual values at each site. (For interpretation of the
references to colour in this gure, the reader is referred to the web version of this article.)
Fig. 4. Normalized ES supply for each land-use type. Yellow gradient: agricultural land-uses; red gradient: semi-natural non-arboreal land-uses; green gradient: semi-
natural arboreal land-uses. Microclimate regulation was calculated by weighted average of inverse daily temperature range (DTR) and inverse daily humidity range
(DHR). Habitat provision was calculated as the weighted average of tree cover and number of vegetation strata. Note that the area covered by each land-use type is
arbitrary (i.e., it depends on the sorting of ecosystem services displayed) and should not be used for comparison (see Table A.6 for normalized ES values).
C. Castellano et al.
Agriculture, Ecosystems and Environment 337 (2022) 108048
8
by trees and woody vegetation, did not show signicant differences with
restored zones, where a greater dominance of herbs can exert an
equivalent effect on nutrient removal. Contrary to the result obtained by
Sabater et al. (2003), in our study the efciency of nutrient ltering was
independent of the nitrate inputs, as the output values (5–10 mg/l) were
equally low irrespective of nitrate inputs (12–35 mg/l).
4.2.2. Microclimate regulation
All arboreal land-uses exerted a high microclimate regulation effect
with equivalent temperature and humidity ranges, irrespective if they
were semi-natural (mature riparian forest, restored riparian forest,
conifer forest) or agricultural (olive groves). However, mature and
restored riparian forests showed the lower mean air temperature and
higher humidity. These results support that arboreal land-uses tend to
reduce the daily air temperature range compared with large open areas
(Dan Moore et al., 2005; Felipe-Lucía et al., 2014). In particular, riparian
gallery forests provide cooler and more stable conditions within highly
disturbed agricultural watersheds (Bertrand et al., 2011; Decocq et al.,
2016), which seems especially relevant in semi-arid zones, where harsh
temperature and dryness prevail. These fragments could ultimately act
as biodiversity shelters in the current context of global change (De
Frenne et al., 2013; Keppel et al., 2012). Wooded riparian zones also
provide shade to the river channel, moderating uctuations in water
temperatures and dissolved oxygen. Such an effect contributes to buffer
eutrophication and, consequently, improves water quality in agricul-
tural watersheds (Broadmeadow et al., 2011).
4.2.3. Habitat provision
As expected, the mature riparian forest, the restored riparian forest
and the conifer forest were the land-uses supplying highest levels of
habitat provision. In this regard, it has been shown that small forest
patches in agricultural landscapes can act as “lifeboat” habitats,
particularly in fragmented landscapes (Decocq et al., 2016). Addition-
ally, microhabitat conditions within these forest patches offer short-term
protection from anthropogenic disturbances dominating in the sur-
rounding matrix and thus act as “transit shelters” or “temporal refuges”
for species (Heroldov´
a et al., 2007) that could deliver other comple-
mentary ES. For example, riparian forests provide habitat for benecial
insects such as ground beetles and insect pollinators, which can enhance
ES that regulate and support agricultural production (e.g. pollination
and pest control), achieving greater crop yields (Cole et al., 2020).
Further, it has been found that woody riparian zones supply higher
habitat provision for all biological dispersal groups (including ants,
pollinators, birds and small mammals) than non-woody riparian zones
(Fonseca et al., 2021).
4.2.4. Soil C storage and soil formation
Although differences among land-uses in soil related-ES were lower
than for other ES, mature riparian forests, conifer forests and rotation
crops displayed the highest capacity of soil C storage and soil formation.
In forests, higher amounts of plant litter enhance C storage through
increased microbial necromass accumulation (Liang et al., 2011).
Additionally, elevated soil C storage at sites with high plant diversity is
directly related to the soil microbial functional community (i.e., soil
biodiversity; Lange et al., 2015). Thus, we could expect much higher
values of soil C storage and soil formation in natural forests than in
rotation crops. However, rotation crops also had high values, which can
be explained by two reasons. On the one hand, most farmers in our study
area adopted conservation tillage practices, based on minimal tillage
systems by reducing the number of interventions and leaving plant
residues on the soil surface. The plant material remaining on the soil
surface or supercially incorporated enhances biological activity and
thus constitutes an important resource of organic matter (Lee and
Stewart, 1983). On the other hand, in agricultural lands, fertilization
and nitrogen xation increase soil C in the majority of cases and
represent an opportunity for soil C storage (Johnson, 1992). However,
excessive fertilization may lead to nutrient leaching from the soil into
aquatic systems causing eutrophication, water quality decline and
further ecological damage (Groffman, 2000).
It must be noted that we did not detect an improvement of soil C
storage and soil formation with the riparian restoration because it is too
early to register the effects of the restoration on these services. Other
studies found that revegetation of riparian zones increased C storage in
soils (Dybala et al., 2019) as a function of restoration age (Matzek et al.,
2020). Therefore, in the mid- and long-term, C storage could suppose a
strong co-benet of riparian restoration in a context of climate change
Fig. 5. Results of Kruskal-Wallis test and Dunn post-hoc comparisons relative to the differences in (a) tree cover and (b) number of vegetation strata among land-use
types. Letters (a, b, c, d) depict the signicant differences found among them (Table A.7). The bold horizontal line denotes the median value, the box delimits the
interquartile range, and the whisker lines extend to the observed maximum and minimum, except for the outliers symbolized by black dots. Blue dots represent
individual values at each site. (For interpretation of the references to colour in this gure, the reader is referred to the web version of this article.)
C. Castellano et al.
Agriculture, Ecosystems and Environment 337 (2022) 108048
9
mitigation.
4.2.5. Soil water holding capacity
The mature riparian forest was the land-use with greater soil water
holding capacity. This result supports that, as high-water-retention
ecosystems, riparian forests decrease both ood peaks and low ows
(Sikka et al., 2003). Due to their water holding capacity, riparian forests
regulate water ows on the ground, streams, and rivers (Maes et al.,
2009), providing substantial benets to human societies (Decocq et al.,
2016). This is especially relevant in Mediterranean agricultural water-
sheds, where episodes of torrential rains and long drought periods
during autumn and summer can cause both destructive ash oods and
river desiccation, respectively, with detrimental effects on crop pro-
duction. These extreme events are expected to increase in the short-term
due to current global change (Llasat et al., 2014).
4.2.6. Water inltration rates and mitigation of surface runoff
All riparian areas and rotation crops displayed the highest rates of
Fig. 6. Results of generalized linear mixed models (GLMM) and post-hoc paired comparisons relative to the differences in (a) microbial biomass, (b) organic C, (c)
water inltration, (d) sorptivity, (d) saturated soil water content, and (f) cation exchange capacity (CEC) among land-use types. Letters (a, b, c) depict the signicant
differences found among them (see Table A.8). The bold horizontal line denotes the median value, the box delimits the interquartile range, and the whisker lines
extend to the observed maximum and minimum, except for the outliers symbolized by black dots. Blue dots represent individual values at each site.(For inter-
pretation of the references to colour in this gure, the reader is referred to the web version of this article.)
C. Castellano et al.
Agriculture, Ecosystems and Environment 337 (2022) 108048
10
water inltration and mitigation of surface runoff. Riparian forests are
acknowledged to promote the inltration of surface water to the
groundwater table, which reduces ood peaks (Brauman et al., 2007).
Contrary to the results of other studies (e.g., Yimer et al., 2008), we did
not nd any signicant differences in water inltration rates and miti-
gation of surface runoff between riparian areas and rotation crops. This
could be because, as mentioned above, many farmers in this area
adopted conservation tillage practices. One of the immediate benets of
this agricultural practices is to increase water inltration through the
soil and reduced surface water runoff compared to conventional agri-
culture (Thierfelder and Wall, 2009). On the other hand, conifer forest
supplied the lowest water inltration rates and less mitigation of surface
runoff. This could be due to the soil water repellency that can occur in
Mediterranean pine forests, especially when soils are dry, declining and
eventually disappearing as soils become progressively wet (Alagna et al.,
2017). Resins, waxes, aromatic oils, and other organic substances from
pines can cause organic coating on soil particles during dry periods,
which is responsible of soil hydrophobicity (Doerr et al., 2000).
4.2.7. Storage capacity of organic pollutants
Finally, shrubland had the greatest storage capacity of organic pol-
lutants, as indicated by cation exchange capacity (CEC). Given that CEC
is positively related to soil organic matter (Paz Ferreiro et al., 2016), we
could expect the highest values of CEC also in mature riparian forests,
conifer forests, and rotation crops. However, CEC can be inuenced by
other physical (e.g., soil texture) and chemical (e.g., pH) soil charac-
teristics (Khaledian et al., 2017). In fact, soil salinity is positively
correlated with CEC (Corwin and Lesch, 2003; Shainberg et al., 1980).
Accordingly, shrubland was the land-use with highest soil salinity
(measured as EC; mean in shrubland =1965.15 µs cm
-1
; mean in all land
uses =594.12 µs cm
-1
), which could explain this elevated storage ca-
pacity of organic pollutants.
5. Conclusions
Our results highlight the main environmental drivers (water table
depth, soil salinity, and soil nutrients) affecting riparian revegetation
success in semi-arid Mediterranean areas. In particular, we provide new
insights for landscape managers about the interactive effects of water
and soil variables driving the survival and growth of planted vegetation,
which should be taken into account to optimize the design of riparian
restorations in Mediterranean agricultural landscapes. This approach
will help to increase the success of further riparian restorations while
progressively recovering ES, contributing to build multifunctional
landscapes.
This study reveals that riparian restorations in agricultural land-
scapes can be effective in terms of enhancing ES supply. Restored ri-
parian zones increased the supply of regulating and supporting ES
(water purication, habitat provision, microclimate regulation and soil
C storage) in comparison with degraded riparian areas and agricultural
crops. However, they did not reach the full magnitude and range of ES
provided by mature riparian forests. Even so, if the observed growth of
planted individuals continues at the same rate, riparian restored areas
can get closer to the level of ES supplied by mature riparian forests in the
medium term.
In semi-arid areas, global change is expected to intensify environ-
mental stress so management measures seems necessary to reach a
balanced supply of provisioning, supporting and regulating ES in water-
limited watersheds. In this respect, our ndings outline that conserving
forest patches and restoring degraded riparian areas is crucial to
reconcile food production with the enhancement of key regulating and
supporting ES in agricultural Mediterranean landscapes.
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.
Acknowledgments
This research was supported by project AMUSE (CGL2014-53017 C2-
1-R, Spanish Ministry of Economy and Competitiveness) led by J.J.J.
and F.A.C. We gratefully acknowledge to Dr. F´
elix Picazo (University of
Granada) for constructive suggestions on earlier versions of the manu-
script. We thank Dr. Brendan Fisher (University of Vermont) for kindly
revising English language. We thank the students (especially Nadia
Marras) and technical staff from IPE-CSIC (especially Alberto Barcos and
Mercedes García) for their help with eld and laboratory work, and
Ricardo Sorando for his help in delimiting the two subwatersheds. We
are also grateful to the farmers interviewed, “Comarca de los Mone-
gros’’, and ”AISECO’’ for their collaboration. This is a contribution of
Ecological Restoration Research Group E40–17R. C.C. was supported by
the Spanish Ministry of Education, Culture and Sport under FPU contract
(FPU14/01682). DB was supported by CSIC Interdisciplinary Thematic
Platform (PTI) Síntesis de Datos de Ecosistemas y Biodiversidad (PTI-
ECOBIODIV) through the "Vicepresidencia Adjunta de ´
Areas Cientíco-
T´
ecnicas (VAACT-CSIC) and “Juan de la Cierva” research contract
(Spanish MINECO FJCI-2016-29856). JMRB acknowledges input from
the REMEDINAL project (TE-CM S2018/EMT-4338) funded by the
Madrid Autonomous Government.
Supporting information
Supplementary data associated with this article can be found in the
online version at doi:10.1016/j.agee.2022.108048. Supplementary data
are organised in:
Appendix A: supplementary tables
Appendix B: supplementary methods
Appendix C: supplementary gures.
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