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\begin{document}

\title{Transformation of phosphorus forms and its regulation on phosphorus
availability across differently degraded marsh soils}



\author[1]{Shirong Zhang}%
\author[1]{Yulin Pu}%
\author[1]{Yuanjun Luo}%
\author[1]{Ting Li}%
\author[1]{Dagang Yuan}%
\author[1]{Xiaoxun Xu}%
\author[1]{Hongyu Qian}%
\author[1]{Guiyin Wang}%
\author[1]{Yong-Xia  Jia}%
\author[1]{Yun Li}%
\affil[1]{Sichuan Agricultural University - Chengdu Campus}%


\vspace{-1em}



  
  \date{May 25, 2024}


\begingroup
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\selectlanguage{english}
\begin{abstract}
Soil phosphorus (P) is an essential nutrient that controls wetland
productivity and ecological functions. However, the effects of soil P
forms on P availability during wetland degradation are relatively
unknown. Soil samples from differently degraded marshes, including
relatively pristine marsh (RPM), lightly degraded marsh (LDM),
moderately degraded marsh (MDM), and heavily degraded marsh (HDM), were
collected to investigate the changes in soil P forms and its regulation
on P availability in the Zoige Plateau, China. We observed that compared
with RPM, the main changes in total P concentration were a significant
increase of 31.6\%--44.2\% in the 0--30 cm soil layers of LDM and MDM,
and the available P concentration increased in LDM and MDM but decreased
in HDM with a lower P activation coefficient. Marsh degradation
increased the concentration and proportion of dicalcium phosphates, P
occluded in iron hydroxides, and organic P but decreased those of iron
oxide surfaces adsorbed P and apatite P. Soil available P was mainly
related to organic P and P non-occluded in iron oxide minerals that
might also be non-negligible direct source of available P. The
transformation from apatite P to organic P was an important regulation
mechanism of P availability in soils during marsh degradation. This
study revealed the risk of P limitation in heavily degraded marsh soils
and established the mechanism by which marsh degradation significantly
influences soil P availability. Therefore, some measure of on improving
P availability should be implement for the ecological restoration of
heavily degraded marsh in the future, such as grazing exclusion and the
application of organic fertiliser.%
\end{abstract}%



\sloppy


\textbf{Highlights}

Marsh degradation induced transformation from soil stable P to organic
and labile P.

Risk of P limitation occurred in heavily degraded marsh soil with low
available P.

Available P in alpine soils was mainly related to organic P and
non-occluded P.

Marsh degradation regulated soil P availability via transformation of
apatite P.

\textbf{Transformation of phosphorus forms and its regulation on
phosphorus availability across differently degraded marsh soils}

Yulin Pu \textsuperscript{a, \#}, Yuanjun Luo \textsuperscript{a, \#},
Shirong Zhang \textsuperscript{b,*}, Ting Li \textsuperscript{a, *},
Dagang Yuan\textsuperscript{a}, Xiaoxun Xu \textsuperscript{b}, Hongyu
Qian\textsuperscript{a}, Guiying Wang \textsuperscript{a}, Yongxia
Jia\textsuperscript{a}, Yun Li \textsuperscript{a}

\textsuperscript{a} \emph{College of Resource Science, Sichuan
Agricultural University, Chengdu 611130, China}

\textsuperscript{b} \emph{College of Environmental Science, Sichuan
Agricultural University, Chengdu 611130, China}

\textsuperscript{\#} \emph{These authors contributed equally to this
work}

\textsuperscript{*}Corresponding author:

Prof. Dr. Shirong Zhang

\emph{College of Environmental Science, Sichuan Agricultural University,
Chengdu 611130, China}

\emph{E-mail address} : rsz01@163.com

Associate Prof. Dr. Ting Li

\emph{College of Resource Science, Sichuan Agricultural University,
Chengdu 611130, China}

\emph{E-mail address} : LT\_SICAU@163.com

\par\null

\textbf{ORCID}\selectlanguage{english}
\begin{figure}[H]
\begin{center}
\includegraphics[width=0.70\columnwidth]{figures/image1/image1}

\end{center}
\end{figure}

Shirong Zhang

\emph{https://orcid.org/0000-0002-1895-2494}

\emph{} \textbf{Abstract:} Soil phosphorus (P) is an essential nutrient
that controls wetland productivity and ecological functions. However,
the effects of soil P forms on P availability during wetland degradation
are relatively unknown. Soil samples from differently degraded marshes,
including relatively pristine marsh (RPM), lightly degraded marsh (LDM),
moderately degraded marsh (MDM), and heavily degraded marsh (HDM), were
collected to investigate the changes in soil P forms and its regulation
on P availability in the Zoige Plateau, China. We observed that compared
with RPM, the main changes in total P concentration were a significant
increase of 31.6\%--44.2\% in the 0--30 cm soil layers of LDM and MDM,
and the available P concentration increased in LDM and MDM but decreased
in HDM with a lower P activation coefficient. Marsh degradation
increased the concentration and proportion of dicalcium phosphates, P
occluded in iron hydroxides, and organic P but decreased those of iron
oxide surfaces adsorbed P and apatite P. Soil available P was mainly
related to organic P and P non-occluded in iron oxide minerals that
might also be non-negligible direct source of available P. The
transformation from apatite P to organic P was an important regulation
mechanism of P availability in soils during marsh degradation. This
study revealed the risk of P limitation in heavily degraded marsh soils
and established the mechanism by which marsh degradation significantly
influences soil P availability. Therefore, some measure of on improving
P availability should be implement for the ecological restoration of
heavily degraded marsh in the future, such as grazing exclusion and the
application of organic fertiliser.

\textbf{Keywords:} Zoige Plateau; Marsh degradation; Phosphorus forms;
Phosphorus transformation; Phosphorus availability

\textbf{1. Introduction}

Phosphorus (P) plays a crucial role in maintaining microbial activity
and plant growth (Cui et al., 2019a; Fan et al., 2021; Huang et al.,
2015a; Oelmann et al., 2017; Sitters et al., 2019) and is anticipated to
replace nitrogen as a limiting nutrient in natural ecosystems (e.g.
wetland ecosystems) (Chen et al., 2020; Du et al., 2020; Turner et al.,
2018; Vitousek et al., 2010). However, elevated P concentrations in
soils and sediments are critical environmental problems (e.g.
eutrophication) (Cheesman et al., 2012; Zhang et al., 2020; Zhou et al.,
2019). In wetland ecosystems, P limitation risk and eutrophication
primarily depend on labile P concentrations (e.g. soluble phosphate)
that can be transformed into other P forms and then back again via
biological or geochemical processes (Gao et al., 2019; Hu et al., 2022;
Vitousek et al., 2010). Moreover, soil P transformation is strongly
linked to carbon (C) cycles via the effect of the microbial community
(Luo et al., 2021; Wang et al., 2021a; Zhai et al., 2022). The
concentrations of P and the composition of P forms ultimately affect
wetland ecological functions such as water conservation,
CO\textsubscript{2} fixation, and climate regulation and health (Bai et
al., 2020; Cui et al., 2019a; Liu et al., 2020; Sorrell et al., 2011).
Therefore, soil P forms have received considerable attention in wetland
ecosystems (Cheesman et al.; 2012; Dunne et al., 2011; Hamdan et al.,
2012; Luo et al., 2021; Qu et al., 2021; Zhang et al., 2020).

The forms of P in soils generally exhibit evident variations owing to
the differences and changes in the pedogenic environments of wetlands,
such as parent material, hydrothermal, and vegetation conditions
(Cheesman et al., 2010, 2012; Cui et al., 2019b; Negassa et al., 2020;
Qu et al., 2021; Wang et al., 2016a). Presently, climate change, rodent
harm, and human disturbances such as drainage, overgrazing, and
aquaculture have led to the degradation of more than half of wetlands
worldwide to some degree (Huisman et al., 2017; Nguyen et al., 2016; Ren
et al., 2019; Zuquette et al., 2020), which has completely altered the
conditions of hydrology, salinity, vegetation, and soil characteristics
in some wetland ecosystems (Cheng et al., 2020; Li et al., 2022; Zeng et
al., 2021; Zhao et al., 2017a). This has further affected soil P
accumulation and its forms via P transformation processes such as
sorption/desorption, precipitation/dissolution,
immobilisation/mineralisation, and weathering (Augusto et al., 2017;
Barrow, 2015; Khosa et al., 2021; Qu et al., 2021; Smith et al., 2021).
For example, cultivation and drainage stimulated the transformation of
organic P to inorganic P and decreased organic P in freshwater wetlands
of the Sanjiang Plain region, China (Wang et al., 2006), and in peatland
wetlands of Mecklenburg-West Pomerania and Saxony-Anhalt, Germany
(Negassa et al., 2020; Schlichting et al., 2002), resulting in reduced P
accumulation. Grazing increased soil P accumulation in wetlands within
the dairy pasture of the Okeechobee Basin wetlands in the United States
(Dunne et al., 2011). However, existing studies have primarily focused
on the varying effects of P forms in wetlands with low elevations, such
as estuaries and coastal areas (Cheesman et al.; 2012; Dunne et al.,
2011; Hu et al., 2021; Zhang et al., 2020), and minimal studies
regarding this varying effect have been conducted in alpine wetlands
where environmental factors are complex and shifting (Li et al., 2022;
Wang et al., 2022; Wu et al., 2021).

The changes in P accumulation and forms would further alter P
availability in wetland soils (Hu et al., 2021; Huang et al., 2015a;
Wang et al., 2021b). For example, Huang et al. (2015a) observed that the
exotic invasive plant \emph{Spartina alterniflora} significantly
increased the concentrations of P extracted by 1.0 mol
L\textsuperscript{-1} NH\textsubscript{4}Cl,
bicarbonate/dithionite-extracted P, and P extracted by 0.5 mol
L\textsuperscript{-1} NaOH in the Yancheng wetland of eastern China,
leading to an increase in available P. Similarly, Hu et al. (2021)
observed that an increase in wetland salinity might significantly
enhance labile P release owing to Fe-bound P reduction in the Min River
estuary wetland, China. Furthermore, the primary forms of P in soils
that regulate P availability differ owing to the differences in
environmental conditions such as vegetation (Gama-Rodrigues et al.,
2014; Hou et al., 2016). For example, some studies have revealed that
the major sources of the available P pool are organic P,
oxalate-extractable P, and iron-bound P in tropical forest, tropical
acid farmland, and temperate meadow soils, respectively (Gama-Rodrigues
et al., 2014; Melese et al., 2015; Yang et al., 2013). In addition,
wetland ecosystems have unique environmental characteristics, such as
perennial waterlogging and anaerobic conditions, and multiple functions
compared with terrestrial ecosystems (Ouyang and Lee, 2020; Shen et al.,
2019). However, limited information is available on the main P forms
that regulate available P in different wetland soils, particularly in
alpine wetlands.

Marsh wetlands on the Zoige Plateau with alpine and fragile environments
are located on the eastern edge of the Qinghai--Tibet Plateau, China,
and have undergone degradation to different degrees because of natural
threats and human activities over the past 60 years (Li et al., 2015,
2019; Shen et al., 2019). Several studies have revealed that alpine
marsh degradation has decreased soil C sink function by increasing C
emission fluxes in the Zoige Plateau (Ma et al., 2016; Pu et al., 2020;
Zhou et al., 2020). Furthermore, total P and available P significantly
affect the concentration of organic C in the Zoige peatland soils (Luo
et al., 2021). However, it remains unclear whether marsh degradation
impacts P accumulation and availability in soils, further inducing P
limitation of primary productivity of alpine wetland ecosystems. Solving
this question would contribute to effectively assessing soil P supply
and further implementing the measures of soil P regulation to improve
the C neutrality potential and promote the ecological restoration of
degraded alpine marshes. Therefore, we hypothesised that marsh
degradation would have a pronounce effect on the accumulation and
transformation of soil P owing to the desiccation accompanied with plant
community and overgrazing, further influencing soil P availability. To
test this hypothesis, this study selected Zoige marsh wetlands with
different degradation degrees and aimed to (1) quantify the changes in
soil P and its forms for exploring the characteristics of P accumulation
and transformation during marsh degradation, (2) determine how marsh
degradation influences soil P availability, and (3) elucidate the
regulation of soil P forms on available P under marsh degradation.

\textbf{2.} \textbf{Materials and methods}

\textbf{2.1 Study area and marsh degradation gradient}

The study area is located in the Huahu Science Spot, Haqingqiao, and
Heihe Village on the Zoige Plateau, China, northeast edge of the
Qinghai-Tibet Plateau (between 102\selectlanguage{ngerman}°25\selectlanguage{english}'--103deg0\selectlanguage{english}' E and 33deg30\selectlanguage{english}'--33deg55\selectlanguage{english}' N)
(Figure 1). The area is at an average altitude of \textasciitilde{}3450
m and subjected to an alpine cold temperate monsoon humid climate, the
wet season between May and September, and the dry season between October
and April. The mean annual temperature is 0.4--1.1 degC, with a minimum
mean monthly temperature of \selectlanguage{english}-10.8 degC in January and a maximum of 10.9 degC
in July. The annual precipitation is 647--753 mm, with a relative
humidity range of 64\textasciitilde{}73\% and a potential
evapotranspiration of 1100--1274 mm. The landform is characterized by a
combination of broad valleys, low mounds, lake depressions, terraces and
river flood plains that are associated with the Black River system (a
tributary of the Yellow River).

In this region, large marsh areas distribute on lowlands, terraces, and
some underground stream gully zones due to the flat, low-lying ground
and poor drainage of surface water. These marsh soils have the similar
pedogenic conditions, such as soil parent materials of the Quaternary
loose sediments from rivers and lakes, perennial waterlogging with water
table depth greater than 10 cm, and the dominant hygrophytes
of\emph{Carex L} . (\emph{Carex muliensis} , \emph{Carex lasiocarpa}
and\emph{Carex meyeriana} ). However, since the early 1960s, these
marshes have occurred a regressive succession from original marsh to
marsh meadow, meadow, desertified meadow with ``black beach'' due to
climate change, rodents and human activities (Li et al., 2015, 2019; Huo
et al. 2013; Zhao et al., 2017b). Taking comprehensively into account
the diagnostic indices of the hydrologic conditions, dominant plant
community and external disturbance (Huo et al., 2013; Wu et al., 2015),
we divided degraded marshes into lightly degraded marsh (LDM),
moderately degraded marsh (MDM) and heavily degraded marsh (HDM)
according to the grading criteria of the degradation of relatively
pristine marsh (RPM) (Table 1). Furthermore, the moderately and heavily
degraded marshes are eroded by wind and easy to be sandy due to an
increase in the area of surface soil exposed to sunlight. The soils of
RPM, LDM, MDM and HDM are classified as Histi-Orthic Gleyosols,
Molli-Orthic Gleyosols, Dark-Aquic Cambosols, Ochri-Aquic Cambosols in
the Chinese soil taxonomy, respectively (Cooperative research group on
Chinese soil taxonomy, 2001).

\textbf{2.2 Soil sampling}

To ensure degraded marshes from the degradation of RPM, we compared the
remote sensing images recorded by the 1977 Landsat 2 MSS and 2015 Google
Earth, and established sampling plots (100 m x 100 m) from differently
degraded marshes following grading criteria of marsh degradation (Table
1) within the Huahu Science Spot, Haqingqiao, and Heihe Village in
August 2017. After three subplots (1 m x 1 m) in each sampling plot were
randomly selected, we investigated the main habitat conditions (Table 2)
and then clipped the aboveground parts of all standing plants (dead and
alive) to measure the aboveground biomass. Subsequently, soil samples
were taken at depths of 0--10, 10--20, 20--30, 30--40, 40--60, 60--80,
and 80--100 cm in each subplot using a soil sampler (Eijkelkamp 07.53.
SE, Netherlands) to determine soil bulk density and other properties.
Three soil cores from the same soil depth in each plot were pooled to
form a composite sample. 84 mixed soil samples were collected from 12
plots. Root biomass estimated from soil bulk density samples at a depth
of 0--60 cm was washed through a 0.5 mm sieve to remove the soil, and
all roots with a diameter of \textgreater{} 0.5 mm were collected and
oven-dried at 65 degC (Table 2).

\textbf{2.3 Soil basic physiochemical properties}

The soil bulk density was calculated as the ratio of the oven-dried (105
degC) soil mass to the volume of the cylinder. The soil pH was determined
in a soil-to-water mass ratio of 1:2.5, and particle size by the United
States Department of Agriculture (USDA) were measured using the
gravimetric method. The soil organic matter was analysed using the
Walkley-Black potassium dichromate sulfuric acid oxidation method
(Nelson and Sommers, 1982). The total N was measured using the
semimicro-Kjeldahl method after complete digestion in
CuSO\textsubscript{4}-K\textsubscript{2}SO\textsubscript{4}-Se
(100:10:1) at 400 degC (Bremner and Mulvaney, 1982), and available N was
determined using NaOH-hydrolysation reduction-diffusion (Lu, 2000). The
total P and available P were measured using the colorimetric molybdenum
blue method via HNO\textsubscript{3}-HClO\textsubscript{4}-HF digestion
and 0.5 mol L\selectlanguage{english}\textsuperscript{-1} NaHCO\textsubscript{3} (pH = 8.5)
extraction, respectively (Murphy and Riley 1962; Olsen et al. 1954;
Thomas et al. 1967). The basic soil physiochemical properties are listed
in Table 3.

\textbf{2.4 Sequentially extracted phosphorus forms}

Given that the soils are derived from the calcareous parent material of
lacustrine sediment, and the development of soils is weak in this study,
we assumed that most of soil P was combined with primary minerals,
existing in the form of ``Calcium (Ca)-bound P''. Furthermore, the forms
of Ca-bound P in the calcareous soils generally have a series of calcium
phosphates with the complex transformation as follows: from dicalcium
phosphate (i e. CaHPO\textsubscript{4}*2H\textsubscript{2}O,
Ca\textsubscript{2}-P) to octacalcium phosphate (i e.
Ca\textsubscript{8}H\textsubscript{2}(PO\textsubscript{4})\textsubscript{6}*5H\textsubscript{2}O,
Ca\textsubscript{8}-P) and apatite (i e.
Ca\textsubscript{10}(PO\textsubscript{4})\textsubscript{6}F\textsubscript{2},
Ca\textsubscript{10}-P) (Jiang and Gu, 1989). Therefore, a sequential
extraction procedure, provided by Jiang and Gu (1989) and modified by
Adhami et al. (2006), was applied to determine P forms instead of the
sequential extraction schemes according to Hedley et al. (1982) for
distinguishing the series of Ca-bound P. Briefly, the
NaHCO\textsubscript{3}-soluble,
CH\textsubscript{3}COONH\textsubscript{4}-soluble,
NH\textsubscript{4}F-soluble,
NaOH-Na\textsubscript{2}CO\textsubscript{3}-soluble, citrate
dithionite-soluble, and H\textsubscript{2}SO\textsubscript{4}-soluble
phosphate was successively extracted with 0.25 mol
L\selectlanguage{english}\textsuperscript{-1}NaHCO\textsubscript{3} (pH = 7.5), 0.5 mol
L\selectlanguage{english}\textsuperscript{-1}CH\textsubscript{3}COONH\textsubscript{4} (pH =
4.2), 0.5 mol L\selectlanguage{english}\textsuperscript{-1} NH\textsubscript{4}F (pH = 8.2), 0.1
mol L\selectlanguage{english}\textsuperscript{-1} NaOH-Na\textsubscript{2}CO\textsubscript{3},
0.3 mol
L\selectlanguage{english}\textsuperscript{-1}C\textsubscript{6}H\textsubscript{5}Na\textsubscript{3}O\textsubscript{7}plus
1.0 g Na\textsubscript{2}S\textsubscript{2}O\textsubscript{4} plus 1.0
mol L\selectlanguage{english}\textsuperscript{-1} NaOH (citrate dithionite) and 0.5 mol
L\selectlanguage{english}\textsuperscript{-1} H\textsubscript{2}SO\textsubscript{4},
respectively. The above soluble phosphate was mainly composed of the
form of Ca\textsubscript{2}-P, Ca\textsubscript{8}-P, aluminium oxide
surfaces adsorbed P (Al-P), iron oxide surfaces adsorbed P (Fe-P),
occluded P in iron oxide minerals (occluded P), and
Ca\textsubscript{10}-P, respectively (Jiang and Gu, 1989; Adhami et al.,
2006), and was termed as Ca\textsubscript{2}-P, Ca\textsubscript{8}-P,
Al-P, Fe-P, occluded P, and Ca\textsubscript{10}-P respectively. Among
these forms, Ca\textsubscript{2}-P, Ca\textsubscript{10}-P and occluded
P are equivalent to NaHCO\textsubscript{3}-extracted P, diluted
HCl-extracted P and concentrated HCl-extracted P, respectively, in
Hedley et al. (1982) scheme. The Ca\textsubscript{8}-P, Al-P and Fe-P
are inorganic P non-occluded in iron oxide minerals, and named as
non-occluded P.

According to the results of Jiang and Gu (1989) and Adhami et al.
(2006), the solubility and bioavailability of Ca\textsubscript{2}-P are
higher than those of non-occluded P and thus Ca\textsubscript{2}-P and
non-occluded P was named as labile P and moderate labile P,
respectively, whereas occluded P and Ca\textsubscript{10}-P with a
lowest solubility and bioavailability was termed as slowly cycling P
(recalcitrant P or stable P). Total organic P was obtained as the
difference in P through 0.5 mol
L\selectlanguage{english}\textsuperscript{-1}H\textsubscript{2}SO\textsubscript{4} extraction
between ignition and non-ignition soils (Saunders and Williams, 1955).
The correlation coefficient (\emph{r} = 0.990) between total P
determined by
H\textsubscript{2}SO\textsubscript{4}-HClO\textsubscript{4} digestion
and the sum of various P forms (including organic P) was significant
(\emph{p} \textless{} 0.001), which showed that the values of various P
forms were reliable in our study.

\textbf{2.5 Data calculations and statistical analyses}

The soil P activation coefficient and available P were characterized by
P availability, and the P activation coefficient was the proportion of
available P content in total P content (Li and Zhang, 1994; Sharma and
Chowdhury, 2021). All statistical tests were performed or produced using
SPSS 18.0 software (IBM Corp., Armonk, NY, USA). All data were tested
using a z-test for normality before analysis. Data meet normality if
absolute z-scores for skewness or kurtosis (skew or kurtosis values
dividing their standard errors) are less than 1.96 and 3.29 in small (n
\textless{} 50) and medium-sized (50\textless{} n \textless{} 300)
samples, respectively. Thus, some data were square root transformed
(total P, P activation coefficient) or log10 transformed (P forms
excluding Ca10P), exponent transformed (pH) to meet normality.
Two-factor analysis of variance was conducted to determine the effects
of marsh degradation and soil sampling depth on soil P and its forms;
one-way analysis of variance (ANOVA) followed by Duncan's test was used
to compare the differences in the concentrations of available P, P
activation coefficient and other basic physiochemical properties among
differently degraded marsh soils. Difference was considered to be
significant when \emph{p} \textless{} 0.05. The figures except for the
structural equation model were drawn in Origin 9.0 (OriginLab Corp.,
Northampton, MA, USA). Structural equation model, combining factor
analysis and path analysis, was broadly employed to analyse the
relationship between variables (including unobserved variables) based on
the covariance matrix of variables in natural ecosystems (Bai et al.,
2020; Hou et al., 2016; Lefcheck, 2016; Melese et al., 2015). Therefore,
we used a structural equation model to further identify the effects of
different P forms in regulating available P variations and the complex
interactions among P forms in soils.

Theoretical and experimental studies demonstrated that soluble P
(available P) was regulated by the direct and indirect effect of various
P forms (Gama-Rodrigues et al., 2014; Hou et al., 2016; Liu et al.,
2019; Melese et al., 2015), and the changes in soil environments (e.g.
plant community and hydrography) also influenced the transformation of P
forms to available P (Hu et al., 2021; Wang et al., 2021b). Thus, a
conceptual structural equation models of the direct and indirect effects
of soil P forms on available P under marsh degradation are constructed
in Figure 2 based the following hypothesis: (1) marsh degradation has an
indirect effect on available P by the mobilisation of stable and
moderate labile inorganic P, and organic P; (2) the mobilisation of soil
P in alpine wetlands mainly follow a process from organic P to inorganic
P, and from stable P to moderate labile P and labile P in turn (Hou et
al., 2016). The model was fitted using a combined data from the three
sites, four types of marsh degradation, and seven depths, totalling 84
observations, because P forms and available P from the same profile have
the significant differences (Table 4; available P of one-way analysis of
variance: F statistic value 5.185, \emph{p} \textless{} 0.05). Thus, the
minimum sample size can meet a general rule that the ratio of the total
number of samples to the number of variables is 5 times (Lefcheck,
2016). The categorical variables of marsh degradation were indicated by
the wetland degradation index that was calculated by Qinghai Provincial
Standards ``Degradation classification of alpine marsh wetlands''
(DB63-T-1794-2020). The wetland degradation index was equal to the
average of water area ratio and importance value of hygrophytes.
Structural equation model was performed using the SPSS Amos 24.0
software package (IBM SPSS Inc., Chicago, IL, USA) after all variables
excluding marsh degradation were log10 transformed. The best-fit model
was derived using the maximum likelihood based on the relative/normed
chi-square test (\selectlanguage{greek}χ\selectlanguage{english}2/df), normed fit index (NFI), comparative fit index
(CFI), and root square mean error of approximation (RMSEA) of model fit
(Hooper et al., 2008; Grace and Bollen, 2005). Model fits with
\selectlanguage{greek}χ\selectlanguage{english}\textsuperscript{2}/df ranges from 2.0 to 5.0, NFI \textgreater{} 0.95,
CFI \textgreater{} 0.95, and RMSEA \textless{} 0.08 can be accepted as a
good structural equation model (Hooper et al., 2008; Grace and Bollen,
2005).

\textbf{3 Results}

\textbf{3.1 Soil total phosphorus concentration}

The total P concentration was significantly affected by soil depth and
marsh degradation degree (\emph{p} \textless{} 0.05, Table 4). However,
no significant differences were observed between the interaction effect
of soil depth and marsh degradation degree (\emph{p} \textgreater{}
0.05, Table 4). As the soil depth increased, soil total P concentration
generally decreased (Figure 3a--d). However, marsh degradation generally
increased soil total P concentrations, and their significant changes in
soil total P concentration were primarily embodied in the shallow soils
at 0--30 cm depth (Figure 3a--d). The total P concentrations of shallow
soils significantly increased by 31.6\%--38.5\% and 32.9\%--44.2\% in
LDM and MDM (\emph{p} \textless{} 0.05), respectively, but slightly
increased in HDM compared with RPM (Figure 3a--d).

\textbf{3.2 Soil phosphorus forms and their proportions}

Soil depth and marsh degradation degree also significantly affected the
concentrations of other P forms, except for Fe-P (\emph{p} \textless{}
0.05, Table 4). As the soil depth increased, soil organic P,
Ca\textsubscript{2}-P, Ca\textsubscript{8}-P, Al-P, and occluded P
concentrations generally decreased, whereas soil Ca\textsubscript{10}-P
concentration slightly increased. The Fe-P concentration increased in
LDM soil but exhibited no apparent trend in other soils (Figure 3a--d).

As illustrated in Figure 3a--d, the organic P concentration generally
increased in degraded marshes. Over the entire soil profile, the average
organic P concentrations in the MDM, HDM, and SDM were significantly
higher (79.9, 78.1, and 82.1\%, respectively) than those in the RPM
(\emph{P} \textless{}0.05). Along the marsh degradation gradient, the
Ca\textsubscript{2}-P and occluded P concentrations increased, reaching
a maximum in the MDM (Ca\textsubscript{2}-P) and LDM (occluded P) soils
and subsequently decreased; Fe-P concentration decreased, reaching a
minimum in the MDM soil and subsequently increased;
Ca\textsubscript{10}-P concentration decreased gradually, whereas no
apparent changes in Ca\textsubscript{8}-P and Al-P concentrations were
observed. Compared with those in the RPM soil, the Ca\textsubscript{2}-P
and occluded P concentrations from the entire profile of degraded marsh
soils increased by 46.7\%--137.7\% and 31.7\%--49.3\%, respectively,
whereas Fe-P and Ca\textsubscript{10}-P concentrations decreased by
17.2\%--32.2\% and 15.3\%--27.8\%, respectively.

Based on the proportion of each P form (ratio of various P forms to
total P), Ca-P (including Ca\textsubscript{2}-P, Ca\textsubscript{8}-P,
and Ca\textsubscript{10}-P) was the dominant P form for the whole soil
profile. However, marsh degradation altered the proportion of the P
form. Compared with those in the RPM soil, Ca\textsubscript{2}-P,
occluded P, and organic P proportions in the 0--30 cm soil layer from
degraded marshes increased by 21.9\%--70.7\%, 8.1\%--15.8\%, and
35.2\%--64.0\%, respectively, whereas Al-P, Fe-P, and
Ca\textsubscript{10}-P proportions decreased by 10.7\%--43.4\%,
19.9\%--60.5\%, and 32.2\%--50.7\%, respectively (Figure 4a). The
changes in soil Ca\textsubscript{2}-P, occluded P, organic P, Fe-P, and
Ca\textsubscript{10}-P proportions at 30--100 cm depth were similar to
those at 0--30 cm because of marsh degradation, whereas Al-P and
Ca\textsubscript{8}-P proportions at 30--100 cm in degraded marsh soils
irregularly changed compared with those in RPM (Figure 4b). Moreover,
organic P in soils changed to the first dominant form at 0--30 cm depth,
accounting for 41.2\%--50.0\% (Figure 4a).

\textbf{3.3 Soil phosphorus availability}

The available P concentration and P activation coefficient first
decreased and then slightly increased in the LDM soil but gradually
decreased in other soils as the soil depth increased (Figure 5a--b).
Compared with RPM soil, the available P concentration was
17.0\%--234.0\% and 0.0\%--85.1\% higher in all soil layers from the LDM
and MDM, respectively, with a significant difference at the 10--30 cm
and 60--100 cm layers in the LDM and the 0--20 cm layers in the MDM
(\emph{p} \textless{} 0.05) (Figure 5a); it decreased by 4.0\%--30.7\%
at the 0--40 cm depth of HDM but hardly changed at the 60--100 cm depth
(Figure 5a). The variations in P activation coefficient between the RPM
and degraded marsh soils exhibited a trend similar to that of available
P (Figure 5a--b). Furthermore, the P activation coefficient in the
surface (0--10 cm) soil was lower than the threshold of soil P
deficiency of 2.0\% (Li and Zhang, 1994).

\textbf{3.4} \textbf{Soil available phosphorus related to phosphorus
forms}

The conceptual structural equation model in Figure 2 was successfully
fitted to our data, and the goodness-of-fit metrics were satisfied
(\selectlanguage{greek}χ\selectlanguage{english}\textsuperscript{2}/df was approximately 1, NFI \textgreater{} 0.95,
CFI \textgreater{} 0.99, and RMSEA \textless{} 0.08; Figure 6a). All P
forms explained 79\% of the variations in soil available P. The
non-occluded Ca\textsubscript{8}-P and Al-P exhibited a strong effect on
the variations in soil available P, whereas the effect of non-occluded
Fe-P was so minor that it did not exhibit in the fitting structural
equation mode (Figure 6a).

Soil organic P had a strong total effect (0.784) on the variations in
available P and was the most important positive factor influencing the
variations in available P, followed by non-occluded P (0.578), occluded
P (0.447), Ca\textsubscript{2}-P (0.409) and Ca\textsubscript{10}-P
(-0.093) (Figure 6b). This indicated that the variations of available P
were mainly related to organic and non-occluded P in alpine wetland
soils. The direct effect of soil non-occluded and organic P on available
P was slightly less than that of Ca\textsubscript{2}-P (Figure 6b),
suggesting that soil non-occluded and organic P also might be
non-negligible direct source of available P. Marsh degradation exhibited
a high direct effect (0.455) on the variations in
Ca\textsubscript{10}-P. Meanwhile, Ca\textsubscript{10}-P presented a
high indirect effect (-0.402) on the variations in available P through
the effect of organic and non-occluded P in soils (Figure 6a--b). These
revealed that marsh degradation significantly influenced soil P
availability via the transformation from Ca\textsubscript{10}-P to
organic and non-occluded P.

\textbf{4. Discussion}

\textbf{4.1 Marsh degradation increases soil phosphorus accumulation}

In this study, the P accumulation (total P concentration) in all soil
profiles ranged from 183.0 to 1433.8 mg kg\textsuperscript{-1}, with an
average value of 683.3 mg kg\textsuperscript{-1}, indicating slight P
enrichment according to the threshold of 500 mg
kg\textsuperscript{-1}(DeBusk et al., 2001). Similar to estuaries or
coastal, depressional, and riparian wetlands (Huang et al., 2015a; Wang
et al., 2006, 2013; Zhang et al., 2015), the P accumulation in all soils
exhibited a decreasing trend with increasing soil depth (Figure 3a--d)
because plant detritus was deposited, which led to P enrichment in
surface soils (Meyerson et al. 2000; Reddy et al. 1999); however, no
significant difference was observed in vertical variations among
differently degraded marshes.

Marsh degradation primarily increased soil P accumulation at 0--30 cm
depth (Figure 3 a-d) owing to the comprehensive effects of hydrothermal
environments, input of organic materials, and human activities. In RPM,
long-term floods and highly anaerobic environments may stimulate the
reductive dissolution of binding partners of phosphate (Kjaergaard et
al., 2021) and increase soluble P concentrations in the overlying water
(Dunne et al., 2011; Kr\selectlanguage{ngerman}öger et al., 2012; Qu et al., 2019; Zhou et al.,
2019), further resulting in a significant loss of soluble P via runoff
and leakage. However, marsh desiccation can greatly decrease the loss of
soluble P, even under light or mediated grazing. Hence, the accumulation
of shallow soil P in LDM and MDM significantly increased owing to marsh
desiccation and a higher litter (approximate to aboveground biomass)
input compared with RPM (\emph{p} \textless{} 0.05, Figure 3. a--c),
which was inconsistent with the results that wetland degradation due to
cultivation activities decreased soil P accumulation in the Sanjiang
Plain, China (Wang et al., 2006) and Saxong-Anhalt, Germany (Schlichting
et al., 2002). This might be because cultivation decreases the amount of
soil P returned owing to crop absorption (Schlichting et al., 2002; Wang
et al., 2006).

As meadows continued to degrade into desertified meadows, soil P
accumulation in HDM was significantly lower than that in LDM and MDM.
Two facts could explain this: (1) decrease in litter input and root
biomass (Table 2) reduced the amount of soil P returned owing to
overgrazing and severe rodent damage; and (2) decrease in vegetation
coverage, soil clay, and organic matter (Tables 1 and 2) weakened the
ability of soil fixation and increased P loss of dissolved P via runoff
and leakage (Cui et al., 2018). Similarly, previous results demonstrated
that the replacement of the native mangrove community by the invasive
plant \emph{Phragmites australis} reduced the ability of soil to adsorb
immobilised P in the Minjiang River estuarine wetland, China (Wang et
al., 2016a), and cultivations reduced P accumulation in the Sanjiang
Plain wetlands, China (Wang et al., 2006). However, desiccation and
excreta from bovines and sheep may supplement a certain amount of P to
offset the partial loss of P. Thus, soil P accumulation in HDM was
slightly higher than that in RPM, which is consistent with the result of
Dunne et al. (2011) that soil P accumulation increased because of the
input of organic fertiliser (cattle manure) in the Okeechobee Basin
wetlands in the United States.

\textbf{4.2 Marsh degradation influences the transformation of soil
phosphorus forms}

The most remarkable changes in P forms due to marsh degradation was an
increase in organic P concentration and proportion, and organic P became
the predominant P form in surface soils (Figures. 3a--d and 4a--b) in
Zoige, which is consistent with the results of the Min River Estuary,
China (Zhang et al., 2015). However, some studies have revealed that the
concentration and proportion of organic P in wetland soils decreased
because cultivation and drainage stimulated the transformation from
organic P to inorganic P and increased organic P loss (Negassa et al.,
2020; Schlichting et al., 2002; Wang et al., 2006). Negassa et al.
(2020) demonstrated that the proportion of organic P was higher in
long-term rewetted peatlands than in drained peatlands. This result was
closely associated with marsh desiccation, combined with the changes in
the plant community or overgrazing. The alpine marsh experienced light
or moderate degradation, and the dense plant communities of hygrophytes
and mesophytes, such as \emph{Carex} , \emph{Blysmus sinocompressus,}
and\emph{Kobresia tibetica} (Table 2), increased plant coverage,
aboveground biomass (Table 2), and biodiversity and richness owing to
water table drawdown (Zeng et al., 2021) and drying-rewetting (Wang et
al., 2021b). Consequently, plant roots might take up more bioavailable
inorganic P and then return more organic P (Huang et al., 2015b) via a
high litter input compared to RPM, particularly in surface soil (Table
2). Although livestock foraged some plant shoots during grazing, this
effect was very minor and hardly altered the above increase in soil
organic P because of the limitation of flooding or excess water during
the outdoor grazing season (wet season) in LDM and MDM. Alpine marshes
are heavily degraded, and more livestock excreta can compensate for the
decrease in organic P owing to livestock foraging and an increase in the
bioaccumulation of organic P in surface soils during overgrazing.
Furthermore, the improvement in aerobic environments in degraded marshes
can enhance the microbial activity and decomposition rate of litter and
livestock excreta (Gyaneshwar et al., 2002; Jones et al., 2018; Prado
and Airoldi, 1999; Qian et al., 2010), which can induce labile organic P
migration from the upper to lower layers of natural soil profiles
through soil porewater to further increase the organic P concentration
in subsurface soils (Pan et al., 2021).

Ca-P was not only the dominant inorganic P form but also the majority of
total P for the whole soil profile in this study (Figures 3a--d and
4a--b) because the soil was derived from the calcareous parent material
of lacustrine sediment, where the concentration of calcium carbonate
equivalent ranged from 61.3 to 240.6\selectlanguage{english} g kg\selectlanguage{english}\textsuperscript{-1} (data from
an investigation of soil series survey in Sichuan Province, China). For
Ca-P forms, a remarkable change was observed in the concentration and
proportion of available Ca\textsubscript{2}-P increased under marsh
degradation, whereas those of unavailable Ca\textsubscript{10}-P
decreased (Figures 3a--d and 4a--b). This is because marsh desiccation
accompanied by the transition of the plant community or overgrazing may
enhance the activity and/or quantity of phosphate-dissolving bacteria,
such as \emph{Pseudomonas} , \emph{Bacillus,} and \emph{Erwini} (Liu et
al., 2015; Susilowati et al., 2019) and increase the organic acids from
plants, microbes, and/or livestock excreta (Almeida et al., 2020;
Gyaneshwar et al., 2002; Miao et al., 2013; Nuryana et al., 2019).
Improved phosphate-dissolving bacteria and organic acids can further
stimulate the transformation of Ca\textsubscript{10}-P to other labile
inorganic P forms such as Ca\textsubscript{2}-P, Ca\textsubscript{8}-P,
and Al-P (Almeida et al., 2020; Susilowati et al., 2019). Similarly,
Jamil et al. (2016) revealed that Calcisols with greater biological
activity under sugarcane crop cover had a high concentration of
Ca\textsubscript{2}-P in the estuary plains of Pakistan. However, the
input of external P fertiliser from surrounding pasture uplands resulted
in an increasing trend of exchange P (primarily Ca\textsubscript{2}-P)
with increasing hydroperiod in isolated Lake Okeechobee in Florida, USA
(Cheesman et al. 2010), which was contrasting with our result, owing to
a significant loss of soluble P (e.g. Ca\textsubscript{2}-P) and lack of
external P in RPM with perennial flooding. Another remarkable change in
inorganic P forms was that marsh degradation decreased the concentration
and proportion of soil Fe-P but increased those of soil occluded P
similar to the changes in Al-P or Fe-Al-P during soil drying (Bai et
al., 2019; Zhang et al., 2015). This is primarily because soil
Fe\textsuperscript{2+} is oxidised and transformed into colloidal ferric
hydroxide and amorphous hydrated iron oxide (Ding et al., 2016; Wang et
al., 2021b), which can adsorb and immobilise more phosphate during soil
desiccation, resulting in the transformation from Fe-P to occluded P
(Sah and Mikkelsen, 1986, 1989).

In addition, marsh degradation did not alter the vertical variation of
all forms of P. However, the vertical variation differed among the
different forms (Figures 3a--d and 4a--b). The concentration and
proportion of Ca\textsubscript{10}-P evidently increased with increasing
soil depth because of the downward trend of soil minerals from the
surface to the bottom soil, whereas the concentrations of other
inorganic P forms and organic P generally decreased with depth because
of `surface aggregation' from plants and microbes (Meyerson et al. 2000;
Reddy et al. 1999; Wu et a., 2020). These results are supported by those
of previous studies on wetland soils (Luo et al., 2021; Wang et al.,
2008; Yuan et al., 2015; Zhang et al., 2020).

In summary, the differences in P forms among marsh soils with different
degrees of degradation likely indicates that marsh degradation primarily
induced the transformation from recalcitrant Ca\textsubscript{10}-P to
organic P and labile Ca\textsubscript{2}-P via litter return or
livestock excreta combined with marsh desiccation.

\textbf{4.3 Marsh degradation alters soil phosphorus availability}

Our results showed that P availability in alpine wetland soils decreased
with soil depth owing to organic matter accumulation in shallow soils
and `surface aggregation' of soil nutrients, particularly surface soils
(Ockenden et al., 2014; Wu et a., 2020). This result conforms with those
of previous studies on freshwater and salt wetland soils (Negassa et
al., 2020; Qu et al., 2019; Wang et al., 2008; Xu et al., 2012). Soil P
availability hardly had significant differences in the vertical
variations among differently degraded marsh soils. However, it generally
increased along the marsh degradation gradient from the RPM, reaching a
maximum in the LDM or MDM soil and subsequently decreasing to the lowest
in HDM (Figure 5a--b). This pattern can be primarily attributed to the
changes in factors affecting soil P availability, such as the
hydrological regime, vegetation presence, and external disturbance
(Duhamel et al., 2017; Ganjegunte et al., 2018; Kroger et al., 2012; Qu
et al., 2018; Wang et al., 2021a, b). As RPM evolved into marsh meadow,
the drying-rewetting and changes in vegetation community promoted the
dissolution of slowly cycling P, decomposition of extracellular polymers
and dead microorganisms containing labile monoester and diester P and
plant debris containing phytate P, and the release of colloid-bound P
through the breakdown of soil aggregates (Stevens and Tullos, 2011; Wang
et al., 2021b). Thus, soil P availability was significantly higher in
LDM than in RPM (Figure 5a--b), although the released soluble P might
have been slightly lost because of inundation (Kroger et al., 2012; Zhou
et al., 2019) in wet season. This was similar to the results previously
observed in the water-level fluctuation zones of the Three Gorges
Reservoir area, China (Wang et al., 2021b), estuarine wetlands with
different vegetation in the Yellow River Delta, China (Qu et al., 2021,
2018; Xu et al., 2012), and managed wetland cells of Mississippi, USA
(Kroger et al., 2012).

As marsh meadow continuously degraded into the meadow, the comprehensive
effects of desiccation and vegetation resulted in a significant increase
in soil P availability (Figure 5a--b) through the decrease in labile P
loss, according to Kroger et al. (2012), and transformation from slow
inorganic P and organic P to bioavailable P owing to the activation of
organic acids from microbes and plant roots and debris (Hallama et al.,
2019; Qualls and Richardson, 2000; Schelfhout et al., 2021; Wang et al.,
2021a). However, when the marsh was heavily degraded, soil microbes
significantly decreased by 39.3\%--94.1\% compared with the other three
types of wetlands (Pu et al., 2022), and the activities of
phosphomonoesterase, phosphodiesterase, and phytase were lower than
those in the other wetlands by 50.5\%--55.9\%, 44.3\%--52.6\%, and
34.2\%--70.5\% (data from unpublished results), respectively. These
results indicate that the mobilisation of soil in slowly cycling P was
inhibited by the low activities of microbes and phosphatase through
overgrazing and rodent damage stress. Therefore, P availability in the
HDM soils decreased compared with that in RPM (Figure 5a--b), and even P
activation coefficient in surface soil was lower than the threshold of
soil P deficiency (Figure 5b). This decrease indicated a risk of P
limitation via plant biomass removal in HDM soils. Similarly, Shang et
al. (2016), Teng et al. (2020) and Zhai et al. (2022) revealed that P
limited plant growth in degraded alpine meadow and saline wetland
ecosystems.

\textbf{4.4 Regulation of soil phosphorus forms on available phosphorus
under marsh degradation}

Organic P can directly or indirectly transform into available P by
mineralization, and some labile organic P\textsubscript{,} such as
nucleic acids, phospholipids, and mononucleotides, might be directly
available forms of P for plant uptake (Gatiboni et al., 2021; Tiessen et
al., 1984). Thus, soil organic P not only acted as a dominant P form of
regulating available P, but also was an important direct source of
available P (Figures 6 a--b), because Zoige wetland soils had higher
organic P concentration compared with other wetlands (e.g. estuarine
wetland) soils (Cheesman et al., 2010; Li et al., 2018; Shao et al.,
2019; Zhang et al., 2015). The moderate labile non-occluded P, similar
to labile Ca\textsubscript{2}-P, had a high regulation on available P,
which was accorded to the reports from Gama-Rodrigues et al. (2014) and
Hou et al. (2016), and also supported by a previous result exhibiting
the significant correlations between available P and Al-P in
albic-bleached meadow soils (Yang et al., 2013). Meanwhile, non-occluded
P is easy to be mobilized into labile P (e.g. Ca\textsubscript{2}-P) due
to the activation of organic acids and phosphate-dissolving bacteria
(Almeida et al., 2020; Susilowati et al., 2019), and indirectly regulate
available P (Hou et al., 2016). Hence, soil non-occluded P was the
second P form of regulating available P (Figures 6 a--b). Soil available
P was primarily related to organic and non-occluded P that might also be
non-negligible direct source of available P in alpine wetland ecosystem.
However, to confirm organic and non-occluded P considered as the direct
source of available P, further examination of the contribution of soil
organic and non-occluded P to plant P uptake is required in future
studies.

Marsh degradation significantly influenced soil available P through the
transformation from soil Ca\textsubscript{10}-P to organic and
non-occluded P, especially organic P (Figure 6a--b). This can be
ascribed to the differences in hydrography, vegetation, and grazing
between RPM and degraded marshes. For degraded marshes, soil organic P
increased (Figures 3a--d and 4a--b) via plant uptake of bioavailable
inorganic P and a high litter input in LDM and MDM with drying-rewetting
and/or dense vegetation (Table 1 and 2); an increase in input of
livestock excreta in HDM with overgrazing. Moreover, the risk of P
limitation occurred with low available P in HDM. Some studies confirmed
that the frequent drying-rewetting and organic acids from litter and/or
organic fertiliser (e.g., livestock excreta) might induce direct and
indirect transformation from apatite and organic P to labile and
moderately labile inorganic P (Hallama et al., 2019; Schelfhout et al.,
2021; Wang et al., 2021b), organic acids could dissolve occluded P into
non-occluded P (Gatiboni et al., 2021; Touhami et al., 2020; Wang et
al., 2016b), and organic and occluded P might be potential sources of
available P in P-deficient soils (Turner et al., 2014; Yu et al., 2019).
Additionally, the loss of carbon pools owing to marsh degradation may
also result in the release of moderate labile inorganic P such as Fe-P
and Al-P from organic matter (Yuan et al., 2015; Zhang et al., 2015). In
this study, an increase in organic acids associated with a mobilisation
of phosphate-dissolving bacteria and a loss of organic carbon (Gatiboni
et al., 2021, Almeida et al., 2020; Susilowati et al., 2019; Pu et al.,
2020) can improve the activation from soil Ca\textsubscript{10}-P to
organic P, non-occluded P and Ca\textsubscript{2}-P under marsh
degradation. Therefore, the transformation from Ca\textsubscript{10}-P
to organic P was an important regulation mechanism of P availability in
soils during marsh degradation on the Zoige plateau.

\textbf{5. Conclusions}

Overall, our results verified our primary hypothesis that marsh
degradation significantly influenced the accumulation and transformation
of soil P owing to the desiccation accompanied with plant community and
overgrazing, further altering soil P availability. Lightly and
moderately degraded marsh soils had significantly higher total P and P
availability than relatively pristine marsh soils. However, soil P
availability in the heavily degraded marsh was lower than that in the
relatively pristine marsh. Marsh degradation primarily induced the
transformation from soil slowing P to organic and labile P via litter
return or livestock excreta combined with marsh desiccation. For alpine
wetland ecosystem, soil available P was mainly related to organic and
non-occluded P that might also be non-negligible direct source of
available P. Marsh degradation regulated soil P availability via the
transformation of apatite P to organic P. Our work illustrates that the
risk of P limitation occurred in heavily degraded marsh soils with low
available P. Future research should focus on measures to improve P
availability for the ecological restoration of desertified meadows, such
as grazing exclusion and the application of organic fertiliser. In
addition, the variations of organic P fractions in differently degraded
marsh soils also should be investigated to understand the transformation
of organic P fractions and their effects on soil P supply, which has
great significance for the ecological restoration of alpine degraded
marshes.

\textbf{Acknowledgements}

This research was financially supported by the grant from Applied Basic
Research Programs of Science \& Technology Department of Sichuan
Province, China {[} grant number 2021YJ0341{]}, the National Natural
Science Foundation of China {[}grant number 41401328{]} and the Key
Research and Development Program of Sichuan Province, China
(2019YFN0020).

\textbf{Declaration of competing interest}

The authors declare that they have no known competing financial
interests or personal relationships that could have appeared to
influence the work reported in this paper.

\textbf{Author contributions}

Yulin Pu: Funding acquisition, Methodology, Supervision, Writing -
Original Draft, Review \& Editing. Yuanjun Luo: Conceptualization,
Investigation, Visualization, Writing - Original Draft, Review \&
Editing. Shirong Zhang: Funding acquisition, Methodology, Supervision,
Writing - Review \& Editing. Ting Li: Validation, Visualization, Writing
- Review \& Editing. Dagang Yuan and Xiaoxun Xu: Validation, Writing -
Review \& Editing. Hongyu Qian: Sampling and Resources. Guiying Wang,
Yongxia Jia, and Yun Li: Formal analysis, Validation.

\textbf{Data availability statement}

Data available on request from the authors

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