Different impacts of native and exotic earthworms on rhizodeposit carbon sequestration in a subtropical soil

Soil Biology & Biochemistry 90 (2015) 152e160

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Different impacts of native and exotic earthworms on rhizodeposit carbon sequestration in a subtropical soil Jinhua Huang a, b, Weixin Zhang a, *, Mengyun Liu a, b, c, d, María J.I. Briones e, Nico Eisenhauer c, d, Yuanhu Shao a, Xi'an Cai a, Shenglei Fu a, Hanping Xia a, ** a

Key Laboratory of Vegetation Restoration and Management of Degraded Ecosystem, South China Botanical Garden, Chinese Academy of Sciences, Guangzhou 510650, China University of Chinese Academy of Sciences, Beijing 100049, China c German Centre for Integrative Biodiversity Research (iDiv) Halle-Jena- Leipzig, Deutscher Platz 5e, 04103 Leipzig, Germany d Institute of Biology, University Leipzig, Johannisallee 21, 04103 Leipzig, Germany e Departamento de Ecología y Biología Animal, Universidad de Vigo, 36310 Vigo, Spain b

a r t i c l e i n f o

a b s t r a c t

Article history: Received 29 May 2015 Received in revised form 3 August 2015 Accepted 5 August 2015 Available online 19 August 2015

Earthworms are known to regulate the sequestration of soil and leaf litter carbon (C). However, their impacts on the more accessible rhizospheric C, which represents a major energy source for soil food webs and an essential factor for C sequestration, are still unclear. Previous studies indicate that earthworms regulate the dynamics of SOC and leaf litter-C by increasing C accessibility to microbiota. However, in the case of labile rhizodeposit-C, microbiota might not require any pre-conditioning by earthworms and may rapidly metabolize most of this root-derived C. Consequently, potential pathways by which earthworms may affect the fate of rhizodeposit-C would be to regulate the biomass and/or activity of rhizosphere microbiota and, further, to mineralize/stabilize microbial products. A 13CO2 labelling experiment was performed to determine the impacts of four different earthworm species on the fate of tree rhizodepositC in a subtropical soil. We hypothesized that endogeic earthworm species, representing primarily geophagous species, would closely interact with soil microbiota and sequester the microbially metabolized rhizodeposit-C more efficiently than epigeic and anecic earthworm species. We found that irrespective of ecological group affiliation, the three native earthworms did not affect rhizodeposit-C sequestration. In contrast, the exotic endogeic species stimulated the immobilization of rhizodeposit-C in the biomass of root-associated bacteria and/or arbuscular mycorrhizal fungi and, further, accessed the microbiotametabolized rhizodeposit-C more efficiently. As a consequence, the exotic endogeic earthworm species transiently tripled rhizodeposit-C retention in soil. We propose that the weak linkages between native earthworms and rhizodeposits-related microbiota limit earthworm impacts on rhizodeposit-C sequestration. However, the exotic endogeic species Pontoscolex corethrurus may potentially alter rhizodepositC dynamics in invaded areas by shifting rhizosphere microbial community composition. This work highlights a distinct mechanism by which earthworms can regulate C dynamics and indicates a significant contribution of invasive earthworm species to belowground processes. © 2015 Elsevier Ltd. All rights reserved.

Keywords: Root-derived carbon Carbon accessibility Earthworm ecological groupings Earthworm invasion Rhizosphere microbiota 13 CO2-labelling

1. Introduction Rhizodeposition is the largest source of labile carbon (C) inputs to the soil, and rhizodeposits are primarily composed of easily decomposable compounds, e.g., sloughed off root caps, border cells

* Corresponding author. Tel./fax: þ86 020 3725 2592. ** Corresponding author. Tel.: þ86 020 3725 2678; fax: þ86 20 37252831. E-mail addresses: [email protected] (W. Zhang), [email protected] (H. Xia). http://dx.doi.org/10.1016/j.soilbio.2015.08.011 0038-0717/© 2015 Elsevier Ltd. All rights reserved.

and root hairs, high molecular weight secretions (mucilages), and low molecular weight organic substances (root exudates) (Nguyen, 2003; Jones et al., 2009). Rhizodeposits account for 5e40% of the net photosynthesized C (Farrar et al., 2003; Nguyen, 2003; Condron et al., 2010) and fuel most processes occurring in the rhizosphere € gberg et al., 2001; Hütsch et al., 2002; Bardgett et al., 2014). (Ho Root-derived C (including both root litter and rhizodeposits) is estimated to account for 50e70% of the C stored in soils of boreal forests (Clemmensen et al., 2013). Also, the contribution of rhizodeposition to soil ecological processes may be enhanced by current

J. Huang et al. / Soil Biology & Biochemistry 90 (2015) 152e160

global environmental changes. For instance, although the rising atmospheric CO2 concentration may simply increase throughput of C (Staddon et al., 2014), it has been suggested to increase the C allocation to root growth and the rhizosphere (Rogers and Runion, 1994; Phillips et al., 2012), which may in turn affect soil C sequestration positively (Wilson et al., 2009) or negatively (Cheng et al., 2012). Also, Boone et al. (1998) concluded that live roots and rhizodeposition are key factors in regulating C emission and C sequestration rates in a warmer world. Soil biota play a determining role in regulating ecosystem functioning and the response of soils to global change (Bardgett and van der Putten, 2014). Earthworms that act as ecosystem engineers have a strong influence on belowground processes through feeding, burrowing, and casting activities (Lavelle et al., 1997; Brown et al., 2000; Scheu, 2003). Earthworms are known to regulate the sequestration of soil organic carbon (SOC) and leaf litter C (Fahey et al., 2013; Lubbers et al., 2013; Zhang et al., 2013). However, their impacts on the more accessible C released into the rhizosphere, which represents an important energy source for soil food webs (Pollierer et al., 2007; Bonkowski et al., 2009; Gilbert et al., 2014) and an essential factor in soil C sequestration (Miltner et al., 2012; Cotrufo et al., 2013), are still unclear. Therefore, a better understanding of the magnitude of earthworm impacts on the fate of rhizodeposit-C is needed. There are five pathways by which earthworms may affect rhizodeposit-C sequestration in soil. Firstly, earthworms may enhance rhizodeposit-C sequestration by facilitating plant growth (Berova and Karanatsidis, 2009; van Groenigen et al., 2014) and the rhizodeposit-C input into soil. Secondly, earthworms may directly access and mineralize and/or stabilize rhizodeposit-C (Brown and Doube, 2004; Gilbert et al., 2014). Thirdly, earthworms may change the fate of rhizodeposit-C by enhancing its accessibility to microbiota as shown for SOC and leaf litter-C (Zhang et al., 2013). Fourthly, earthworms may affect rhizodeposit-C dynamics through regulating the biomass and/or activity of rhizodeposits-related microbiota (Brown and Doube, 2004). Finally, earthworms may indirectly access rhizodeposit-C from microbiota and their residues and, further, affect the retention of rhizodeposit-C through gut processes. Given that most of this root-derived C is readily accessible to rhizosphere microbiota (Todorovic et al., 2001; Kuzyakov and Larionova, 2005; Balasooriya et al., 2014), we consider that the last mechanism is potentially of most relevance. Hence, we hypothesized that endogeic earthworm species, which represents primarily geophagous species, would closely interact with soil microbiota and access microbially metabolized rhizodeposit-C more efficiently, and consequently, have a greater contribution to rhizodeposit-C sequestration than anecic and epigeic earthworm species. Given that the effect of invasive species on the composition and functioning of native ecosystems may be particularly strong when it is functionally dissimilar to the native community (Wardle et al., 2011; Frelich et al., 2012), we compared the effects of three native earthworm species belonging to different ecological groups (epigeic, endogeic, and anecic species) with that of the exotic endogeic species Pontoscolex corethrurus, which is invasive in south subtropical China. This exotic endogeic species is resistant to stress/ disturbance (e.g., drought and infertility) probably due to its behaviour of curled-up quiescence and the presence of welldeveloped calciferous glands (Fig. S1), which may have important roles in regulations of water and acid-base balances (Briones et al., 2008). In fact, P. corethrurus is one of the most widespread pantropical species (Marichal et al., 2012) and the sole exotic earthworm species in the region south of the Nanling mountain ranges of China (Zhang et al., 2005; Hendrix et al., 2008). The proportion of the population density of this invasive species has

153

increased from less than 50% (Liao and Chen, 1990) to more than 95% in some non-natural forests in the last few decades in tropical regions of China, i.e., native earthworm species can only be occasionally sampled (data not shown). In addition, P. corethrurus has been observed to be dominant in some primary forests of South America (Marichal et al., 2010). The presence of this peregrine earthworm species may profoundly alter soil structure, water infiltration, and soil fertility (Jouquet et al., 2006; Briones, 2014). Hence, it is essential to compare the functioning of native and exotic earthworm species to better understand the changes of earthworm contributions to C dynamics in response to the invasion of P. corethrurus. Therefore, a 13CO2 labelling experiment was performed to quantify the impacts of these earthworm species on the belowground sequestration of recently photosynthesized C by a common tree species Schima superba in the subtropical region of China. We determined the incorporation of rhizodeposit-C into bulk soil and different sized aggregates, and earthworm gut contents and tissues as well as the mineralization potential of soil retained rhizodepositC. The effects of different earthworm ecological groupings on soil microbial biomass and activity and their d13C signatures were examined to explore the underlying mechanisms. 2. Materials and methods 2.1. Construction of the labelling chamber In September 2011, a labelling chamber (2 m long  1.6 m wide  2 m high) was built to permit the injection and monitoring of CO2. The chamber was constructed using an aluminium frame, Plexiglas walls (5 mm thick) and roof (2 mm thick). A CO2 probe (GMM220 Series CO2 transmitter modules, Vaisala, Netherland), a quantum sensor (LI190SB, Campbell Scientific, Inc., USA), and a temperature and relative humidity probe (CS215, Campbell Scientific, Inc., USA) were fixed inside the chamber and connected to a control panel and a data logger outside. A 13C-labelled and a nonlabelled CO2 gas cylinder were also connected to the control panel. If the concentration of CO2 inside the chamber did not reach the desired concentration, the valve would automatically open to release more 13CO2 or non-labelled CO2 into the chamber. An airconditioning system was installed to maintain a constant temperature (20  C) inside the chamber. 2.2. Establishment of the microcosms The soil (C%: 6.22 ± 0.10%; N%: 0.41 ± 0.01%; clay%: 19.67 ± 0.88%; pH: 3.80 ± 0.01; d13C27.48 ± 0.03‰; d15N: 0.78 ± 0.07‰) used in the microcosm experiment was collected from the 0e15 cm layer of a 70-year old forest in Dinghushan Biosphere Reserve, Guangdong, China (23120 N, 112 340 E). It is a lateritic red soil with S. superba as one of the dominant tree species. In June 2012, 3 kg of fresh sieved soil (2 mm; soil water content: 33.9 ± 0.2%) was used to fill 30 nylon-bags (20 cm wide  40 cm long, mesh size 0.25 mm), with the bags preventing earthworms from escaping. Each of these soil-filled nylon-bags was introduced into a plastic pot (20 cm in diameter  20 cm in height), and one seedling of S. superba (50e60 cm in height) was planted in each pot. All 30 pots were watered using a drip irrigation system every 3 days, and the watering stopped when excess water started to drain through the holes at the bottom. 2.3. Collection and inoculation of earthworms In September 2012, earthworms belonging to three ecological groups were hand-sorted in the field from the same location as the

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soil. All the collected native earthworm species seemed to be undescribed species, we then classified them to the genus level and recorded as Amynthas sp1. (epigeic species, Epi), Amynthas sp2. (native endogeic species, End), Pontoscolex corethrurus (exotic endogeic species, End e exotic) and Amynthas sp3. (anecic species, Ane). Each earthworm and control (without earthworms) treatment had six replicates. In order to account for their different densities in the field as well as different body sizes and weights, four individuals of the epigeic species, one individual of the native endogeic species, four individuals of the exotic endogeic species, and one individual of the anecic species were inoculated in each treatment pot, resulting in an average fresh biomass of 2.20 ± 0.17, 2.11 ± 0.17, 1.98 ± 0.08, and 4.92 ± 0.8 g pot1, respectively. In addition, 5 g air-dried senescent litter (a mixture of S. superba and Castanopsis chinensis) collected from the same field site was added to the surface of each pot to provide enough food for the earthworms and to mimic the field environment. 2.4.

13

CO2 labelling

The pulse labelling started in late September 2012, 10 days after the inoculation of earthworms (once all the earthworms appeared to adapt to the new environment and no escaping was observed). To ensure maximum photosynthesis efficiency, the 13CO2 (99 atom % 13C, Cambridge Isotope Laboratories, INC., USA) was delivered into  and the chamber between 11:00 am and 1:00 pm (Kastovska   , 2007) for five consecutive days. If it was sunny, a Santr u ckova relatively higher CO2 concentration (500 ppm) was set and the delivery process lasted for 1 h; however, if it was cloudy, a lower concentration (370e400 ppm) and a shorter duration (ca. 30 min) were used. At the end of the 5 days approximately 16 L of 13CO2 had been delivered. During the following 15 days, the pots were kept inside the chamber to maximise the assimilation of 13CO2 by the plants. For the last 11 days, non-labelled CO2 (d13C ¼ 31.49 ± 0.07‰) was delivered into the chamber to maintain the CO2 concentration inside at around 380 ppm, and approximately 120 L of this CO2 was delivered. 2.5. Sampling and analyses Although photosynthate C can be transferred from the plants to the soil food web in a matter of few days, there is a time lag for the invertebrates to get labelled which varies between 6 and 30 days for different groups (e.g. Ostle et al., 2007; Churchland et al., 2012); therefore, after 20 days of labelling exposure, the destructive sampling was conducted. The major reason behind this relatively short-term experiment was to avoid any possible interference derived from the turnover of the less decomposable fractions of root-derived C (Gilbert et al., 2014) so that the impacts of earthworms on the “readily accessible rhizodeposit-C” can be focused. The plants were separated into leaf, stem, coarse roots (1e5 mm), and fine roots (<1 mm). The roots were cleaned with tap water and rinsed in deionised water. All plant materials were oven-dried at 55  C. A subsample (ca. 5e6 mg) of each type of these plant materials was used to determine d13C signature and C content. The soil CO2 flux during the 20 days of incubation was not measured to avoid disturbance of the labelling process. However, the rhizodeposit-C distribution belowground was quantified by tracing the d13C signatures in earthworm tissues and gut contents, bulk soil, and aggregates. Earthworms were hand-sorted and the numbers and fresh weights recorded. Earthworms were euthanized by keeping them in the freezer (20  C) for 2e5 min. Subsequently, the posterior 33% end of the body was sectioned, the gut dissected, and the gut content washed off with deionised water and collected. Both the tissue and gut content were then freeze-dried for d13C

signature and C content analyses. The anterior section of the earthworm body was preserved in 70% ethanol for confirmation of taxonomical identity. One subsample (ca. 200 g) of fresh soil was air-dried to determine soil aggregate composition and the d13C value and C content of each fraction. Soil aggregates were separated by wet sieving 100 g air-dried soil through a series of sieves (with mesh sizes of 2000 mm, 250 mm and 53 mm) (Six et al., 1998). Another subsample (ca. 350 g FW) was passed through a 2 mm sieve and approximately 50 g was freeze-dried for isotopic and C content analyses as well as for the determination of soil microbial biomass (phospholipid fatty acids profile, PLFAs) and d13C value of PLFAs. The remaining 300 g of soil from each microcosm were stored at 4  C. To differentiate the “direct effects” of earthworms (those which remain in the system even after removing the animals; for example due to their burrowing activities) from those result from their interactions with the microbiota (“indirect effects”) we quantified the mineralization potential of the rhizodeposit-C retained in both the control and the earthworm-worked soils (without plant and earthworms) in the longer term. Accordingly, 50 g of those soil samples stored at 4  C were re-incubated and soil basal respiration and the d13C signature of the soil respired CO2 were measured 30 and 540 days after sampling, respectively. The two incubations were continued for 48 and 96 h, respectively, with the temperature and soil water content maintained at 25  C and ca. 26%. The respired CO2 from anecic species-worked soil was not determined on day 540 due to logistic reasons. On each of these two sampling occasions, the CO2 produced was trapped in a NaOH solution (0.1 M) and subsequently measured by titration with HCl (0.1 M) to a phenolphthalein endpoint after adding excess BaCl2. The BaCO3 was washed/centrifuged and oven-dried at 55  C, and the d13C signature of respired CO2 was determined in the newly formed BaCO3. Given that the d13C signature of respired CO2 markedly declined on day 540, additional soil samples stored at 4  C were freeze-dried to determine the C content and d13C signature on day 630 so that the longer-term stability of soil-retained rhizodeposit-C could be assessed. Soil PLFAs were analysed following the methods described by Bossio and Scow (1998). The bacterial PLFA biomarkers (Frostegård and Bååth, 1996) were: (i) 16:1u7c, 15:0 3OH, cy17:0, 16:1 2OH, 18:1u7c, and cy19:0u8c for gram-negative bacteria (G bacteria); (ii) i14:0, i15:0, a15:0, i16:0, i17:0, a17:0 for gram-positive bacteria (Gþ bacteria), and (iii) 14:0, 15:0 and 17:0 as general bacteria biomarkers. The fungal PLFA biomarkers included 18:2u6,9 and 18:1u9c (Kaiser et al., 2010; Frostegård et al., 2011). Arbuscular mycorrhizal fungi (AMF) were indicated by PLFA 16:1u5c; however, as this PLFA may also be of bacterial origin (Olsson, 1999; Ngosong et al., 2012; Staddon et al., 2014), the respective results should be treated with caution. In addition, the ratio of cy17:0 to 16:1u7c was used to indicate bacterial activity (Grogan and Cronan, 1997). The sum of all these PLFAs plus 16:1u5c, 10Me 16:0, 10Me 17:0, and 10Me 18:0 was used as an estimation of the total microbial biomass. Prior to isotopic and C content analyses, all dried solid materials were ground using a ball mill and sieved through a mesh of <0.15 mm, whereas the PLFAs samples were dissolved in 100 ml nhexane. The d13C signatures, d15N signatures, C contents, and N contents were then determined using an Elementar Vario EL Cube or Micro Cube elemental analyzer (Elementar Analysensysteme GmbH, Hanau, Germany) or a PDZ Europa ANCA-GSL elemental analyzer interfaced to a PDZ Europa 20e20 isotope ratio mass spectrometer (Sercon Ltd., Cheshire, UK). Results are expressed as: d13C(‰) ¼ [(13Rsample/13Rstandard)1]  1000, where 13R ¼ 13C/12C, and d15N(‰) ¼ [(15Rsample/15Rstandard)1]  1000, where 15 R ¼ 15N/14N. Pee Dee Belenite (PDB) and atmospheric N2 were used as standard for 13C and 15N, respectively.

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2.6. Isotopic mass balance calculations

fb þ froot þ fSOCt0 ¼ 1

A standard approach to estimate the allocation of recently photosynthesized 13C or recovery of added 13C is to calculate the atom % excess 13C (Svejcar et al., 1990; Chaudhary et al., 2012; Fahey et al., 2013). In this study, however, we tried to quantify earthworm effects on the sequestration of both 13C-labelled and non-labelled rhizodeposit-C in the soil. Hence, we used d value-based mixing models (Schmidt et al., 2004; Staddon, 2004) to calculate the fractions of rhizodeposit-C in the bulk soil, soil aggregates and respired CO2, microbial PLFAs, and earthworm gut contents; whereas the amount of rhizodeposit-C in each component was calculated via multiplying the total C content of each component by its respective fraction. It is difficult to directly determine the amount or d13C signature of rhizodeposits. Nevertheless, fine roots are generally assumed to be responsible for a large fraction of the rhizodeposition process (Jones et al., 2009), and Cheng (1996) reported that the d13C signature of the roots represented the d13C value of root-derived CO2 (including both root respiration and rhizodeposits-related respiration). Therefore, in this study, the d13C value of fine roots was used as an indicator of the 13C signature in rhizodeposits, which consist of both the 13C-labelled and nonlabelled newly formed photosynthate-C. The equations used to calculate the fractions of rhizodeposit-C in different components were:

.

d13 CSOCt0  d13 Croot fsoil ¼ d13 CSOCt0  d13 CSOCt1 (1) .
faggregates ¼ d13 Caggregatest0  d13 Caggregatest1
 d13 Caggregatest0  d13 Croot .

d13 CSOCt0  d13 Croot fCO2 ¼ d13 CSOCt0  d13 CCO2

(2)

(3)

.

d13 CSOCt0  d13 Croot fPLFAs ¼ d13 CSOCt0  d13 CPLFAs (4)

fgut

content

.
¼ d13 CSOCt0  d13 Cgut content 
d13 CSOCt0  d13 Croot

155

(6)

and

i h fworm ¼ d13 Cwormt1  ð1  fb Þ  D  d13 CSOCt0 h i þ d13 CSOCt0  d13 Cwormt0 Þ  fb =  ðd13 Croot  d13 CSOCt0

(7)

where fb is the proportion of earthworm background C, which was assumed to be 0.75 since the experimental duration was relatively short (Martin et al., 1992), and D represents the 13C enrichment factor in the earthworm bodies from their putative diets, which was assumed to be 3.68‰ (Zhang et al., 2010); froot and fSOCt0 are fractions of rhizodeposit-C and SOC in the earthworm tissues at the end of labelling, respectively; d13Cwormt0 and d13Cwormt1 represent d13C of the earthworms before and after labelling, respectively; d13Croot and d13CSOCt0 are the same as in equation (1). 2.7. Statistical analysis One way ANOVA was performed to examine earthworm effects on the allocation of rhizodeposit-C in bulk soil, soil aggregates, earthworm tissues and gut contents, microbial PLFAs, and soil respired CO2, and to compare the growth, survival rate and natural isotopic abundances (13C and 15N) of earthworms belonging to different ecological groups. LSD test was used for comparison of means when homogeneity of variances was met; otherwise, the data was log10-or square-root-transformed. If after transformation the variances were still unequal, Tamhane's T2 was used as a posthoc test for pairwise comparisons. Independent-Samples T Test was used to test the differences of d13C signature of both earthworm tissues and soil microbial community before and after labelling, between fine root and earthworm tissue or soil respired CO2, to compare both the C content and d13C signature of plant tissues in treatments with and without earthworms, and finally assess the rhizodeposit-C retention in earthworm gut contents. All statistical analyses were performed using SPSS15.0 (SPSS Inc., Chicago, IL) and statistical significance was determined at P < 0.05 level. 3. Results

(5)

3.1. Earthworm dietary preferences and assimilation of rhizodeposit-C

In equation (1), fsoil is the fraction of rhizodeposit-C in the soil,

d13Croot is the d13C of the fine roots, d13CSOCt0 and d13CSOCt1 are d13C of soil before and after labelling, respectively. In equation (2),

faggregates is the fraction of rhizodeposit-C in the large macroaggregates (>2000 mm), small macroaggregates (250e2000 mm), and microaggregates (<250 mm), respectively, with d13Caggregatest0 being the d13C of a given size of aggregates before labelling and d13Caggregatest1 being the d13C after labelling. In equation (3), fCO2 is the fraction of rhizodeposit-C in the soil respired CO2, and d13CCO2 is the d13C of the soil respired CO2 after labelling. In equation (4), fPLFAs is the fraction of rhizodeposit-C in the microbial PLFAs, and d13CPLFAs is the d13C of microbial PLFAs after labelling. In equation (5), fgut content is the fraction of rhizodeposit-C in earthworm gut contents, and d13Cgut content is the d13C of earthworm gut contents after labelling. A three-source mixed model was used to calculate the incorporation of rhizodeposit-C into earthworm tissues after labelling (Phillips and Gregg, 2003). The equations were:

The earthworm survival rates were greater than 83% for all ecological groups except for epigeic species, with only 54% of the individuals remaining alive (Table S1). The 15N natural abundances of earthworms differed significantly between ecological groups (F3,16 ¼ 6.39, P ¼ 0.005) and in particular between the native endogeic species and both the epigeic (P ¼ 0.001) and exotic endogeic species (P ¼ 0.004), and between the anecic and epigeic species (P ¼ 0.023) (Fig. 1). The d13C signatures of earthworm tissues significantly increased after labelling (t ¼ 4.09, P ¼ 0.001, df ¼ 36), but they did not differ between ecological groups either before or after labelling (F3,14 ¼ 1.46, P ¼ 0.269 and F3,16 ¼ 1.59, P ¼ 0.231, respectively) (Fig. 2a). In addition, the d13C signatures of earthworm gut contents did not differ between treatments and control soil (F4,22 ¼ 0.32, P ¼ 0.863). Finally, the d13C values of both earthworm tissues and gut contents were much lower than those of fine roots (t ¼ 8.36, P < 0.001, df ¼ 48 and t ¼ 8.39, P < 0.001, df ¼ 49, respectively).

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6 5

although the amount of rhizodeposit-C retained in the exotic endogeic species-worked soil was greater than that of the native endogeic one (P ¼ 0.036) (Fig. 3b). Furthermore, earthworms showed no significant impacts on the d13C signature of any of the soil aggregate fractions. However, on day 0, more rhizodeposit-C was retained in the 250e2000 mm soil aggregates in the treatment containing exotic endogeic species than in the control soil or the soil with native endogeic species (P ¼ 0.04 and P ¼ 0.01, respectively). In addition, more rhizodeposit-C was retained in the >2000 mm soil aggregates in the treatment containing anecic species than in the soil with epigeic species (P ¼ 0.04) (Fig. 3c).

End

4 Ane

3

δ15N (‰)

2 Epi

1

End (exotic)

0 -1

Field soil

-2 -3 -4 -5 -6

3.3. Mineralization potential of soil retained rhizodeposit-C After removing plants and earthworms, the d13C signature of the CO2 respired by soil microbiota did not significantly differ between control and worm-worked soils either at day 30 (F4, 25 ¼ 0.57, P ¼ 0.688) or at day 540 (F3, 20 ¼ 0.14, P ¼ 0.937). Also, the amount of rhizodeposit-C in this microbial-respired CO2 showed no significant differences among treatments at day 0, while it differed significantly at day 540 (F3, 20 ¼ 3.84, P ¼ 0.025), and it was greater for the soil containing exotic endogeic species than for control soil (P ¼ 0.007) or those containing epigeic and native endogeic species (P ¼ 0.024 and P ¼ 0.012, respectively) (Fig. 3d).

Field litter -30 -29 -28 -27 -26 -25 -24 -23 -22 -21 δ13C (‰)

Fig. 1. Natural 15N and 13C abundances of earthworm species belonging to different ecological groups. Mean values are shown ±1 s.e.m. (n ¼ 6).

Notably, the exotic endogeic species showed the highest assimilation of rhizodeposit-C, which was significantly greater than that of anecic species (P ¼ 0.005). The amount of rhizodeposit-C in the gut content of this exotic species was also the highest and significantly greater than that of the native endogeic species and of the control soil (P ¼ 0.044 and P ¼ 0.001, respectively) (Fig. 2b).

3.4. Rhizodeposit-C production: biomass, d13C value and C content of plant tissues There were no differences of the biomasses of plant leaves (F4, ¼ 1.39, P ¼ 0.266), stems (F ¼ 1.82, P ¼ 0.157), fine roots (F4, 25 ¼ 0.79, P ¼ 0.541), and coarse roots (F4, 25 ¼ 0.68, P ¼ 0.613) among treatments with and without earthworms (Table S2). Also, earthworms did not significantly affect the d13C signatures of plant leaves and plant stems (F4, 25 ¼ 1.30, P ¼ 0.297). In contrast, although the d13C signatures of fine roots did not differ significantly across treatments (F4, 25 ¼ 1.14, P ¼ 0.360), they tended to be lower in those soils containing earthworms than the control soil (924.4 ± 115.6‰ vs. 1496.8 ± 358.7‰) (t ¼ 1.98, P ¼ 0.058, df ¼ 28) (Fig. 4a). Furthermore, earthworm presence significantly reduced the d13C signatures of coarse roots (t ¼ 3.02, P ¼ 0.005, df ¼ 28) (Fig. 4b). Notably, earthworms increased the C content of plant stems significantly (t ¼ 2.98, P ¼ 0.028, df ¼ 28), and that of coarse roots, although the differences were not statistically significant (t ¼ 1.82, P ¼ 0.079, df ¼ 28) (Fig. 4c and d). However, they did not 25

3.2. Rhizodeposit-C retained in bulk soil and aggregates The d13C signatures of soils significantly increased after labelling both on day 0 (t ¼ 2.16, P ¼ 0.038, df ¼ 31) and day 630 (t ¼ 9.79, P < 0.001, df ¼ 32) when compared to that of field soil, but they did not differ significantly between treatments neither on day 0 (F4, 24 ¼ 1.01, P ¼ 0.424) nor on day 630 (F4, 25 ¼ 0.57; P ¼ 0.690) (Fig. 3a). However, on day 0, the amount of soil-retained rhizodeposit-C in the treatment with exotic endogeic species was three to five folds greater than those with either control soil (P ¼ 0.007), or the native populations of epigeic, endogeic and anecic species (P ¼ 0.003, P ¼ 0.009 and P ¼ 0.037, respectively). Nevertheless, treatment effects decreased on day 630 (F4, 24 ¼ 1.55, P ¼ 0.221),

Epi

(b)

-8 -16 -24 -32

End(Exotic)

a a

a a a a a t = -4.09; P = 0.001 a

Before labelling After labelling

Rhizodeposit-C in earthwrom gut content (mg kg-1 dry substrate)

δ13C of earthworm tissue (‰)

(a)

End

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Fig. 2. d13C signatures in earthworm tissues (a) and the amount of rhizodeposit-C in earthworm gut contents (b) for species belonging to different ecological groups. Mean values are shown ±1 s.e.m. (n ¼ 6). Different letters indicate significant difference between treatments.

J. Huang et al. / Soil Biology & Biochemistry 90 (2015) 152e160

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Fig. 4. d13C signatures in plant fine and coarse roots (a,b) and C content in plant stem and coarse roots (c,d). In all panels, mean values are shown ±1 s.e.m. (n ¼ 6). Different letters indicate significant difference between treatments.

affect the C content of the actively growing tissues, i.e., plant leaves and fine roots (F4.25 ¼ 0.47, P ¼ 0.76 and F4,25 ¼ 0.07, P ¼ 0.99, respectively). 3.5. Soil microbial community and the assimilation of rhizodeposit-C Both total soil microbial biomass (F4, 25 ¼ 16.10, P < 0.001) and bacterial biomass (F4, 25 ¼ 17.28, P < 0.001) were significantly

increased by earthworms, except in the treatment with native endogeic species (Fig. 5a and b). In addition, the exotic endogeic species significantly increased the biomass of both soil fungi, Gbacteria, and PLFA 16:1u5c relative to that of control soils (P ¼ 0.022, P ¼ 0.005 and P ¼ 0.006, respectively) (Fig. 5cee). In contrast, the earthworm ecological groupings did not significantly change the ratio of cy17:0 to 16:1u7c compared to the control soil (F4, 25 ¼ 1.79, P ¼ 0.162).

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Fig. 5. Earthworm impacts on biomass of major soil microbial groups (aee) and the d13C signature of total microbial PLFAs (f). Mean values are shown ±1 s.e.m. (n ¼ 6). Different letters indicate significant difference between treatments.

The d13C signatures of soil microbial community after labelling were greater than their natural abundances (t ¼ 6.98, P < 0.001, df ¼ 52); however, they were not affected by earthworm presence (F4, 25 ¼ 1.07, P ¼ 0.391; Fig. 5f). The d13C values of the microbial community were significantly lower than those of the microbialrespired CO2 at day 30 (t ¼ 11.73, P < 0.001, df ¼ 58) and of the fine roots (t ¼ 8.71, P < 0.001, df ¼ 58). Notably, the proportion of rhizodeposits-derived C in soil microbial biomass was greater in the treatment containing exotic endogeic species than those containing either the epigeic or native endogeic ones (P ¼ 0.024 and P ¼ 0.019, respectively). 4. Discussion Unexpectedly, all the native earthworm species did not have any significant influence on the critical processes involved in rhizodeposit-C sequestration, which was indicated by the unchanged amounts of rhizodeposit-C in both bulk soil and all aggregate fractions as compared to the control treatment. These findings contrast with the recent observation of enhanced sequestration of SOC and leaf litter C by earthworms (Zhang et al., 2013). Given that earthworms regulate C dynamics primarily through gut processes and protected C in aggregates they egest (Lavelle and Martin, 1992; Bossuyt et al., 2005), more rhizodepositC should have been detected in the posterior gut contents than in the control soil, if rhizodeposit-C was protected by earthworms. However, the amounts of rhizodeposit-C in the gut content of the native species were as low as that in the control soil. In addition,

they did not alter the mineralization potential of rhizodeposit-C, because the amounts of rhizodeposit-C in microbial respired CO2 after labelling also showed no difference compared to control soil. Nevertheless, the increased C contents in both stems and coarse roots in the presence of earthworms suggest that these organisms stimulated plant photosynthesis and root growth. The decrease in the d13C signatures in both coarse and fine roots was probably the result of the stimulation of production of non-labelled photosynthate and hence, the dilution of the label in the root tissues. In other words, earthworms may have induced excess production, transport and/or storage of non-labelled photosynthate-C. Accordingly, the input of rhizodeposit-C into soil could be increased by earthworms, and the d13C signatures in fine roots reflected both the input of 13Clabelled and non-labelled photosynthate-C into soil. Whereas, the observed unchanged d13C signatures in earthworm gut contents, bulk soil, and soil aggregates reflected the balance between rhizodeposit-C inputs and mineralization/stabilization and did not necessarily indicate negligible impacts of earthworms on rhizodeposit-C sequestration. Therefore, in the case of the treatments containing native earthworm species, increased inputs of photosynthate-C may have been promptly mineralized once released into the rhizosphere, i.e., native earthworms may have simply increase throughput of C. The limited effects of native earthworm species on rhizodepositC sequestration in soil may be due to the high accessibility of rhizodeposit-C to microbiota. Given that most rhizodeposit-C would be mineralized and/or stabilized by microbiota before being accessed by earthworms (Kuzyakov and Domanski, 2000)

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irrespective of their ecological group identity, earthworms may primarily access rhizodeposit-C indirectly by processing the microbiota-metabolized rhizodeposit-C. This was partially supported by the low d13C values in the earthworm tissues relative to that of the fine roots (16.68 ± 1.84‰ vs. 1038.90 ± 121.36‰). Even if earthworms had directly assimilated greater amounts of rhizodeposit-C from live labile sub-fractions of fine roots (Gilbert et al., 2014), the earthworm-metabolized rhizodeposit-C may be rapidly consumed by soil microbiota. In contrast, earthworms assimilated greater amounts of the less mineralizable C, such as leaf litter and root litter (Zhang et al., 2010; Gilbert et al., 2014), and improved both the microbial activity by C activation and soil C sequestration through greater C stabilization (Zhang et al., 2013). This suggests that the impacts of earthworms on rhizodeposit-C sequestration rely on their influences on rhizosphere microbiota. Earthworms belonging to different ecological groupings could interact with microbiota differently (Curry and Schmidt, 2007), and hence, exert distinct impacts on the fate of rhizodeposit-C. However, all native earthworms showed weak impacts on soil microbiota; we therefore propose that the weak linkages between native earthworms and rhizodeposits-related microbiota (e. g., G-bacteria and AMF) limited their impacts on rhizodeposit-C sequestration. On the contrary, the retention of rhizodeposit-C was enhanced by the exotic earthworm species P. corethrurus in the present experiment. The presence of P. corethrurus substantially increased the amount of rhizodeposit-C in the gut content, bulk soil, and soil aggregates. Additionally, after earthworms and plant had been removed, larger amounts of rhizodeposit-C were mineralized by soil microbiota from the exotic species-worked soils than in the other treatments indicating that the greater retention of rhizodeposit-C by this exotic endogeic species was only transient. The transient positive effect on rhizodeposit-C sequestration may be sustained unless the population of P. corethrurus declined considerably. This exotic earthworm altered the composition of the microbial community in the rhizosphere and subsequently changed the metabolism of rhizodeposit-C. It was estimated that only 0.5e2.1% of the total microbial community incorporated rhizodeposits (Table S3), which may primarily consist of a few specific responsive microbiota (Marilley et al., 1998; Dennis et al., 2010). The similar ratio of cy17:0 to 16:1u7c across treatments also suggests that earthworms, with the exception of the native endogeic species, did not significantly affect soil bacterial activity. Importantly, the biomass of G-bacteria, one of the dominant microbial groups €derberg et al., metabolizing rhizodeposit-C (Grayston et al., 1998; So 2004; Treonis et al., 2004), was significantly increased by the exotic endogeic species. Also, the PLFA 16:1u5c, which mainly derived from AMF (the other important rhizodeposits-related microbial group) and/or bacteria (Olsson, 1999; Ngosong et al., 2012; Staddon et al., 2014), was stimulated by P. corethrurus. Overall, the presence of exotic P. corethrurus may significantly alter rhizosphere microbial community composition and consequently soil C dynamics of invaded habitats. This finding is of particular significance as this invasive species could potentially expand its territory in pantropical regions according to current climate change simulations. Further studies are needed to investigate the potential adaption of soil microbial communities to new earthworm species, and the functional dissimilarity between exotic and native earthworm species belonging to the same ecological grouping (Wardle et al., 2011) to fully understand why the effects on rhizodepositsrelated microbiota were much stronger when the endogeic exotic species were present and whether this response can be extrapolated to other ecosystems. A combination of stable-isotope probing (SIP) and metagenomic techniques would allow to identify which microbial taxa are responsible for metabolizing rhizodeposit-C and

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