RESEARCH ARTICLE
Off-site impacts of agricultural composting: role of terrestrially
derived organic matter in structuring aquatic microbial
communities and their metabolic potential
Thomas Pommier1, Asmaa Merroune2, Yvan Bettarel3, Patrice Got3, Jean-Louis Janeau2,
Pascal Jouquet4,5, Thuy D. Thu6, Tran D. Toan6 & Emma Rochelle-Newall2
de Lyon, Villeurbanne, France; 2IRD, UMR 242, Institute of
UMR CNRS 5557, USC 1364 INRA, Ecologie Microbienne, Universit
e Lyon1, Universite
Ecology and Environmental Sciences – Paris (iEES-Paris), Ecole Normale Sup
erieure, Paris, France; 3UMR CNRS 5119, Ecology of coastal marine
systems, UMR5119, IRD, Montpellier, France; 4IRD, UMR 242, Institute of Ecology and Environmental Sciences – Paris (iEES-Paris), Centre IRD Ile
de France, Bondy, France; 5IFCWS, Civil engineering Department, Indian Institute of Science, Bangalore, India; and 6Soils and Fertilizers Research
Institute (SFRI), Dong Ngac, Chem, Tu Liem District, Hanoi, Vietnam
1
MICROBIOLOGY ECOLOGY
Received 4 April 2014; revised 2 July 2014;
accepted 29 August 2014. Final version
published online 30 September 2014.
DOI: 10.1111/1574-6941.12421
Editor: Wietse de Boer
Keywords
metabolic potential; lakes; off-site effects;
soil leachates.
Abstract
While considered as sustainable and low-cost agricultural amendments, the
impacts of organic fertilizers on downstream aquatic microbial communities
remain poorly documented. We investigated the quantity and quality of the
dissolved organic matter leaching from agricultural soil amended with compost, vermicompost or biochar and assessed their effects on lake microbial
communities, in terms of viral and bacterial abundances, community structure
and metabolic potential. The addition of compost and vermicompost significantly increased the amount of dissolved organic carbon in the leachate compared with soil alone. Leachates from these additions, either with or without
biochar, were highly bioavailable to aquatic microbial communities, although
reducing the metabolic potential of the community and harbouring more specific communities. Although not affecting bacterial richness or taxonomic distributions, the specific addition of biochar affected the original lake bacterial
communities, resulting in a strongly different community. This could be partly
explained by viral burst and converging bacterial abundances throughout the
samples. These results underline the necessity to include off-site impacts of
agricultural amendments when considering their cascading effect on downstream aquatic ecosystems.
Introduction
Intensive agriculture increases production but is increasingly criticized due to the negative impacts such practices
can have on the environment. Particularly in tropical ecosystems, intensive farming induces soil acidification and
increases the loss of soil organic matter (SOM), leading
to rapid declines in fertility and high erosion rates (Barak
& Laird, 1997; Burle et al., 1997; Valentin et al., 2008).
The erosion and loss of nutrients from soils not only
affects soil quality, it also has important impacts on
nutrient and carbon cycling in downstream lakes, streams
and reservoirs (Quinton et al., 2010; Howarth et al.,
2012), altering lake productivity and organic matter remineralization rates (Ka et al., 2006; Karlsson et al., 2009;
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Guenet et al., 2010). Although these processes have been
examined in temperate systems (e.g. Williams et al.,
2010), the impacts of shifting cultivation practices on soil
and eroded organic matter composition on downstream
microbial diversity are still poorly understood in tropical
and subtropical systems. This is particularly the case in
Vietnam, and South-East Asia as a whole, where the
intensification of agriculture is leading to increasing degradation of both terrestrial and aquatic ecosystems (Sidle
et al., 2006; Le et al., 2008).
Organic fertilizers represent a potential solution to
these cascading problems as their rich organic matter
contents improve soil chemical and physical properties.
The application of organic fertilizers such as compost or
vermicompost (i.e. compost produced in the presence of
FEMS Microbiol Ecol 90 (2014) 622–632
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
Correspondence: Thomas Pommier, UMR
CNRS 5557, Laboratoire d’Ecologie
Microbienne, Universite Lyon1, Universite de
Lyon, USC INRA 1364, b^at. G. Mendel, 43
boulevard du 11 novembre 1918, F-69622
Villeurbanne Cedex, France.
Tel.: +33 4 72 43 13 79;
fax: +33 4 26 23 44 68;
e-mail: thomas.pommier@univ-lyon1.fr
623
Off-site impact of organic fertilizers on aquatic microbes
Material and methods
Leachate preparation and lake water
incubations
Compost, vermicompost and biochar were added in layers to soil microcosms (total volume 500 mL) to provide
semifactorial triplicate combinations of soil, compost or
vermicompost and biochar treatments: soil alone as control (S), compost (SC), vermicompost (SV), biochar (SB),
compost + biochar (SCB) or vermicompost + biochar
(SVB) (Supporting Information, Fig. S1). Then, 500 mL
of reverse osmosis water was added to each soil microcosm, and leachates were collected and filtered (Whatman
GF/F 0.7 lm nominal porosity). This addition was equivalent to a 4.8-mm rainfall event relative to the soil surface
in each microcosm; 100 mL of the resulting filtrate was
collected for the analyses of DOC concentration, absorption and fluorescence. Then, 100 mL of this filtrate was
also added to 900 mL of water from Cau Khau lake, Vietnam. The aquatic microcosms were incubated in triplicate
for 72 h in the dark and at in situ temperature
FEMS Microbiol Ecol 90 (2014) 622–632
(c. 27 °C). Samples were collected during the incubation
for: DOC (at T0, T24, T48 and T72 hours), DOC absorption and fluorescence (T0), bacterial and viral abundance
(at T0 and T72 hours) and bacterial diversity (T0 and T72)
and bacterial catabolic capacity (T72).
DOC concentration, absorption and
fluorescence
DOC concentration and DOC absorption and fluorescence [or chromophoric dissolved organic matter
(CDOM)] were determined on filtered (Whatman GF/F)
samples collected in precombusted (450 °C, overnight)
glass tubes, or amber bottles for CDOM, sealed with a
Teflon-lined cap. Thirty millilitres was collected in duplicate for DOC concentration, preserved with 36 lL 85%
phosphoric acid (H3PO4) and stored at ambient temperature and in the dark until measurement on a Shimadzu
TOC VCPH analyser. DOC absorption (m1) and fluorescence were measured on 125-mL samples stored at
20 °C following the method detailed by Rochelle-Newall
et al. (2014). Briefly, before measurement samples were
left to warm to room temperature and re-filtered at
0.2 lm (Sartorius Minisart NML Syringe filters). Absorption was measured using a spectrophotometer (AnalyticJena) from 200 to 750 nm with a 1- or 10-cm quartz
cuvette to avoid internal quenching at high concentrations of CDOM. Fluorescence was estimated with a
Turner Trilogy fluorometer and values were normalized
to a quinine sulfate standard. The values of fluorescence
are given as normalized fluorescence units (NFlU) and
those of absorption are expressed as m1.
Viral and bacterial abundances
Viral abundance was estimated by determining the abundance of virus-like particles on glutaraldehyde-fixed samples (0.5% final concentration) by epifluorescence
microscopy after staining with SYBR Gold (Molecular
Probes, Eugene, OR) (Noble & Fuhrman, 1998; Chen
et al., 2001). Bacterial abundance was measured by a
FACS CALIBUR flow cytometer after staining with SYBRGreen I (Molecular Probes) and sonication for 10 min
(Ultrasonik 300 Ney; Louati et al., 2012; Marie et al.,
1997). The stained bacterial cells, excited at 488 nm, were
enumerated using right-angle light scatter (RALS) and
green fluorescence (FL1) at 530 nm. Fluorescent beads (1
and 2 lm, Polysciences, Inc., Warrington, PA) were
added to each sample as external standard. True count
beads (Becton Dickinson, San Jose, CA) were added to
determine the volume analysed. Data analyses were carried out with CELLQUEST PRO 5 software obtained from BD
Biosciences (Franklin Lakes, NJ).
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
epigeic earthworms) has been shown to increase nutrient
availability to plants and to increase the quantity and
quality of SOM (Aggelides & Londra, 2000; Rivero et al.,
2004). Biochar, produced by pyrolysis of organic products
such as bamboo, has also been proposed as an inexpensive and effective soil amendment (Marris, 2006). It is
considered to significantly improve soil quality by
increasing soil carbon storage and water retention, reducing nutrient leaching as well as by absorbing and immobilizing pollutants in contaminated soils (Novak et al.,
2009; Laird et al., 2010; Beesley & Dickinson, 2011).
However, although the impacts of these organic amendments on soil properties and on plant yield are fairly well
known, studies presenting the impacts of eroded organic
matter from soils subject to these fertilizers on downstream microbial community structure and metabolic
capacity remain scarce.
In the present study we examined the direct, shortterm (72 h) impact of leachate from different types of
organic fertilizers on downstream aquatic microbial communities in terms of dissolved organic carbon (DOC)
bioavailability, bacterial and viral abundances, and bacterial functional and genetic diversity. The objectives of this
work were to determine (1) if the addition of compost,
vermicompost and biochar to soils leads to a shift in the
quality and quantity of DOC leached from the soil; (2) if
this eroded organic matter was bioavailable to natural
aquatic bacteria; and (3) how this added organic matter
impacted aquatic bacterial community structure and metabolic potential.
T. Pommier et al.
624
Functional diversity assessment and metabolic
potential of the community
DNA extraction and bacterial diversity
assessment
Fifty millilitres of each incubation was filtered onto separate 0.2-lm (ø = 47 mm) Supor filters (PALL Corp.) and
immediately frozen in 1 mL TE buffer (10 mm Tris-HCl,
pH = 8.0, 1 mm EDTA). After an enzyme/phenol-chloroform DNA extraction of each sample collected at the end
of all incubations, a portion of the 16S rRNA gene covering the hypervariable regions V1–V3 was amplified using
the primers 27F (50 -GAGTTTGATCMTGGCTCAG-30 )
and 518R (50 -WTTACCGCGGCTGCTGG-30 ). Prior to
pyrosequencing all PCR amplicons were pooled to equimolar ratio. The forward and reverse primers included
both a 10-bp multiplex identifier (MID) to multiplex the
samples during sequencing. Amplifications were performed in triplicate using the AmpliTaq Gold 360 master
mix (Applied Biosystems), according to the protocol of
the manufacturer. Cycling conditions were as follows: an
initial activation/denaturation step at 95 °C for 10 min;
followed by 25 cycles of 95 °C for 40 s, 55 °C for 40 s
and 72 °C for 1 min; and a final 7-min extension at
72 °C. PCR products were then purified using the QIAquick PCR Purification Kit (Qiagen, Hilden, Germany)
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Sequence processing
The raw data including 446 568 reads were processed
using the MOTHUR v.1.32.1 software package following the
standard operative protocol described by Schloss et al.
(2009). Sequencing error rates were reduced using ‘PyroNoise’ (Quince et al., 2011). Sequences with > 1 mismatch to the MID, with > 2 mismatches to the primer
and including homopolymers longer than 8 bp, and the
resulting short (< 200 bp) sequences were removed
(235 420 remaining sequences). Alignment was performed
against the SILVA SEED database using default Needleman–Wunsch algorithm parameters and the more abundant sequences that were within 2 bp were preclustered
to remove PCR amplification and sequencing errors. Chimeras were removed using ‘Uchime’ (Edgar et al., 2011).
The taxonomy of the remaining 212 640 sequences was
assigned using the Bayesian classifier of the RDP reference
taxonomic outline, and retained when a bootstrap value
was over 80% for 100 bootstrap iterations (Cole et al.,
2005; Wang et al., 2007). All sequences showing close taxonomy to chloroplasts, mitochondria, Eukarya or Archaea
were removed. The cured and aligned sequence dataset
contained 185 321 sequences, which were clustered into
operational taxonomic units (OTUs) according to the
furthest neighbour algorithm after constructing a distance
matrix. Subsequent a-diversity (i.e. richness) and b-diversity [i.e. community structure, nonmetric dimensional
scaling (NMDS)] analyses were performed using R (R
Development Core Team, 2008). To test whether the spatial separation observed in the NMDS plot was statistically significant, i.e. whether the centres of the clouds
representing a group are more separate than the variation
among samples of the same treatment, we performed an
AMOVA. The similarity of the samples to each other was
also described as a dendrogram using the Yue & Clayton
measure of dissimilarity between the structures of two
communities (data not shown). To determine whether
the clustering within the resulting dendrogram was statistically significant, we performed a test of weighted Unifrac significance (Hamady et al., 2009). AMOVA and
Unifrac analyses were performed using MOTHUR v.1.32.1
(Schloss et al., 2009).
FEMS Microbiol Ecol 90 (2014) 622–632
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
The catabolic capacity of each aquatic microbial community impacted by the fertilizer leachates was determined
using Biolog Ecoplate 96-well microplates. The microplates include 31 different carbon-based substrates and
one water control in triplicate. Each well also contains
growth media and a tetrazolium violet dye that becomes
purple when the substrate is oxidized and was inoculated
with 150 lL of sample. The plates were incubated in the
dark at in situ temperature (c. 27 °C) for 96 h as recommended by the manufacturer. Colour development (OD
at 590 nm) was measured using a Bio-Rad Laboratories,
Model 680 Microplate Reader every 24 h. The data from
96 h were used to determine average colour development
for each group of substrates (Supporting Information,
Table S1) after subtraction of the appropriate water
blanks. The relative proportion of substrate utilization for
each biochemical group of substrates (amine, polymers,
etc.) within each triplicate plate was then determined.
These were then summed to provide a total respiration
for each treatment. Metabolic potential of each community was normalized for each substrate relative to the
highest performance in the substrate family (Salles et al.,
2009). Radial charts to visualize metabolic potential of
the community were produced using R (R Development
Core Team, 2008).
after excision of the amplicon from an agarose gel. The
concentration of DNA of each identified PCR product
was determined using Picogreen quantification and then
PCR triplicates of the same site were pooled into equimolar concentrations. Pyrosequencing was then performed
on a 454 GS-FLX Titanium (454 Life Sciences) at the
Molecular Research DNA laboratory (USA). All data are
publicly available at the European Nucleotide Archive
under the Study accession number PRJEB6637.
625
FEMS Microbiol Ecol 90 (2014) 622–632
7.2
52
16.8
1.75
23.6
16
161.21c
578.39a
289.1b
164.4c
522.6a
326.4b
1.7
8.6
1.3
1.9
5.2
3.6
18.8
87.3
44.2
19.7
79.5
45.5
12.6
562
213
8.1
258
282
73.8c
4698a
1774b
115.9c
4329a
2068b
0
11
2
0
2
1
0.29
64.4
13.5
0.37
11
27.3
7.8c
560a
265.9b
6.6c
534.6a
299.4b
Soil
Compost
Vermicompost
Biochar
Compost + biochar
Vermicomp. + biochar
0d
54.5a
13c
0.02d
39.9b
13c
SE
SE
The DOC concentration of the added water (13.7 lM C) was subtracted from the total concentration of DOC in the leachate. The initial value of DOC in the lake water before addition of the
leachates was 151.8 lM C. The same letter indicates no significant difference between the treatments and different letters indicate a significant difference between treatments.
12c
27a
14.8b,c
20b
27.6a
18b,c
5.5
11.2
10.8
2.5
43.4
8.9
SE
SE
SE
D DOC
(lM C)
with T72
Initial
DOC (lM C)
Initial
fluorescence
(NFlU)
DOC
(lM C)
Absorption
(355 m1)
SE
Initial concentrations of DOC varied between 161 lM C in
the control soil incubations to > 578 lM C in the compost
treatment. The amount of DOC removed during the
Treatment
DOC remineralization during 72 h of incubation
Fluorescence
(NFlU)
DOC concentration in the leachates differed by a factor of
more than 80 between treatments (Table 1). Given the low
carbon concentrations in the soil microcosms
(30 mg C g1), DOC concentration in the leachate was
low (73.8 12.6 lM C). DOC concentration was slightly
higher in the leachate from the soil amended with biochar
(115.9 8.1 lM C), reflecting the addition of carbon with
the biochar. However, these values were not significantly
different (P > 0.05). In contrast, the addition of both compost and vermicompost to the soil resulted in a dramatic
and significant (P < 0.05) increase in the amount of DOC
in the leachate. High C concentrations were observed for
the compost and for the vermicompost additions
(4698 562 and 1774 213 lM C, respectively). The
addition of biochar to both compost and vermicompost
did not result in a significant change in DOC concentration
relative to their additions alone.
The optical quality as determined by fluorescence and
absorption spectroscopy of the leachates also differed significantly between treatments, and tended to reflect the
trends observed in the bulk DOC concentration
(P < 0.05). As with bulk DOC concentration, fluorescence
was significantly higher in the compost and compost with
biochar treatments as compared with the other treatments
(P < 0.05). The addition of biochar did not result in a significant change in either optical parameter.
Incubations
Characteristics of the leachates
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
Results
Leachate
Significant differences between the replicate treatments
were tested using JMP v.10 (SAS). ANOVA was used to test
the significance of the differences between treatments
after checking that the assumptions of the ANOVA were
met. When necessary, the data were log transformed to
ensure normality. When a significant difference was
observed, an a posteriori test (Fisher’s LSD or Student’s ttests) was used to determine significant groups of data.
Because bacterial abundance at T0 was not normally distributed, even after various corrections, the non-parametric test of Kruskal–Wallis was applied to this dataset to
assess the treatment effect.
More details of the material and methods are available
in the Supporting Information, Data S1.
Table 1. DOC concentration (lM C) and the fluorescence (normalized fluorescence units – NFlU) and absorption (m1) characteristics of the leachate from each of the six treatments
Statistical analyses
19.33c
156.5a
42.9b,c
32.8b
144.1a
58.8b,c
C removed
after 72 h (%)
Off-site impact of organic fertilizers on aquatic microbes
626
incubation also varied significantly as a function of the
treatment (Fig. 1, Table 1, P < 0.05). In the soil treatment,
12% (19 lM C) of the initial DOC was removed during
the 72-h incubation period. This was in comparison to the
biochar treatment where significantly higher removal rates
(32.8 lM C removed, corresponding to a 20% loss) were
observed (P < 0.05). In the compost and compost + biochar treatments (Fig. 1), the highest DOC removal rates
were observed (27.8% and 27.6%, for compost and compost + biochar, respectively). The addition of vermicompost or vermicompost + biochar leachate also resulted in
significantly higher DOC removal rates (P < 0.05), as
T. Pommier et al.
compared with the soil control. However, these rates were
significantly lower (14.8% and 18.0% for vermicompost
and vermicompost + biochar, respectively; P < 0.05) than
for the compost and compost + biochar treatments. In
summary, the addition of leachate from compost and vermicompost with or without biochar resulted in significantly higher remineralization rates as compared with the
soil-only leachate addition. However, although the addition
of biochar to the soil resulted in significantly (P < 0.05)
higher remineralization rates, relative to the soil alone
incubation, the addition of biochar to the compost or vermicompost treatments did not result in a significant change
in DOC remineralization relative to the compost- or vermicompost-alone treatments (Table 1).
Fig. 1. Removal of DOC during incubations of leachate in lake water.
Error bars indicate standard errors. Lines indicate the best fitting
polynomial or linear model (R2 = 0.90–0.99).
Initial bacterial abundance in the incubation varied
between 2.5 9 105 and 5.4 9 105 mL1 with the soil
treatment having the lowest abundances and the compost
with biochar the highest abundances (Fig. 2, Table S1).
Similar to DOC concentration, there were significant
(Kruskal–Wallis, P < 0.05) differences between treatments
with soil alone and biochar on the one hand with the
lowest abundances, and the four other treatments on the
other, reaching 5.4 9 106 bacterial cells mL1 in the
compost + biochar treatment. Considering the value
observed in the soil treatment as a control, the estimated
amount of added bacteria was highest in the compost +
biochar and vermicompost additions (Table S1). The
Fig. 2. Mean bacterial and viral abundances
at initial (black) and final (grey) sampling times
in the different incubations of leachate in lake
water. Letters indicate significant pairing after
Student’s t-tests.
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
FEMS Microbiol Ecol 90 (2014) 622–632
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
Bacterial and viral abundances before and
after 72 h of incubation
627
Off-site impact of organic fertilizers on aquatic microbes
Effect of leachates on the metabolic potential
of the community
Most of the carbon sources were oxidized in the control
soil incubation. However, the addition of any external
nutrient or carbon source, i.e. compost, vermicompost or
biochar, either decreased (4-hydroxybenzoic acid, Nacetyl-D-glucoseamine) or inhibited C degradation capacities. The latter was especially true for the lignin-phenolic
acid monomer 2-hydroxybenzoic acid. The metabolic
potential of the community, in terms of carbon degradation capacity, may be symbolized as polygons based on
the relative proportion of respiration to the total respiration for each biochemical group and for each fertilizer
treatment (Fig. 3). Compared with the control treatment,
the addition of fertilizer resulted in a significant
(P < 0.05) reduction in carbon utilization both as a
whole, i.e. when all substrates were grouped together, as
well as within each biochemical class (Table S2). The only
exceptions were the amine and the polymer biochemical
groups for which no significant difference in respiration
between fertilizers treatments was observed. The reductions in community metabolic potential were observed
for all of the biochemical groups and all of the fertilizers
with the exception of the phenolic acids in the vermicompost + biochar incubation. However, the reduction was
different among all biochemical groups. Large decreases
in the metabolic potential for carbohydrates, carboxylic
acids and phenolic acids were observed in the fertilizer
additions whereas only small decreases were observed for
the N-containing amines and amino acid biochemical
groups and for the polymers. In other words, all
FEMS Microbiol Ecol 90 (2014) 622–632
fertilizing treatments reduced the metabolic potential of
the aquatic microbial communities relative to that of soil
alone, with a particularly strong effect observed for compost + biochar where the community metabolic potential
was reduced to more than half of the soil community
metabolic potential (CNsoil = 23.14 vs. CNcomp+bioch =
8.44). Therefore, the addition of nutrients and C from
fertilizers reduces the metabolic potential of the communities in all biochemical groups, at least regarding the carbon sources present in the Ecoplates.
Impacts on community richness, taxonomy and
structure
These impacts on relative metabolic activities implied
putative changes on community richness and structures.
The number of OTUs of pyrosequences clustered at a distance of 0.03 substitutions per nucleotide did not differ
significantly (P > 0.05) between treatments (Fig. S2), both
as observed or estimated by catchall (Bunge, 2011) and
when including all sequences or after subsampling to the
lowest number of sequences (i.e. 3125 sequences). At the
genus level, the treatments did not show different proportions of the dominant groups (Fig. S3). Nevertheless, the
community structures showed a significant (AMOVA,
P < 0.001) distinction of all treatments that included biochar (Fig. 3). The weighted Unifrac significance showed
significant (P < 0.001) effects of treatments, except for
the vermicompost compared with the control
(P = 0.227).
Discussion
This study aimed to compare the impact of various
organic amendments on the quantity and quality of dissolved organic matter (DOM) in leachates, to determine
the bioavailability of this DOM to aquatic microbial communities and to determine its impact in terms of genetic
structure and metabolic potential. Although interesting,
the actual variation of the lake microbial communities
during the incubation was not examined here, as the
focus was set on the effect of leachate from soil with or
without fertilizer(s). Three organic amendments that are
often cited as viable, sustainable options for organic fertilization of soils were tested: compost, vermicompost and
biochar along with soil with relatively low organic carbon
and nutrients contents (30 mg C g1; 2.5 mg N g1).
The addition of these three amendments, either alone or
in combination with biochar, resulted in a large increase
in organic carbon concentration in the leachate, particularly in the compost additions. Not only was total organic
carbon concentration higher but the chemical quality of
the organic matter, as determined by fluorescence and
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
initial abundances of virus-like particles (VLP) were also
significantly (P < 0.05) different as compared with the
soil control. All abundances in each treatment differed
significantly from initial values to those measured after
72 h of incubation. However, at the end of the incubation, all bacterial abundances converged to a similar range
(from 7.7 9 105 to 1.2 9 106 mL1 in the biochar and
soil treatment, respectively), therefore showing an increase
in the soil and biochar treatments and a decrease in all of
the other incubations, and this despite the higher concentrations of DOC. VLP abundances significantly increased
after 72 h of incubation, with the highest abundances in
the compost + biochar treatment. These VLP bursts may
partly explain the convergence in bacterial abundances. In
summary, due to the higher DOC liability in compost
and vermicompost, bacterial abundances followed the
same tendency as these compounds in the leaching process even when biochar was added. In parallel, VLP abundances also changed with increases in abundance
observed in all of the incubations after 72 h.
628
T. Pommier et al.
Fig. 3. Impact of the different leachates on
the metabolic potential (as measured by
EcoPlates Biolog) of the incubated microbial
communities. The position on each of the six
axes indicates the highest potential
degradation rate of that group of substrates
by the community after a 96-h incubation.
Bioavailability to aquatic bacteria
DOC is the largest pool of organic carbon in aquatic ecosystems, and bacteria play an important role in the recycling of this organic matter (del Giorgio & Davis, 2003).
Moreover, shifts in the quantity and quality of DOM can
result in changes in bacterial functional and genetic diversity (Bouvier & del Giorgio, 2002; Rochelle-Newall et al.,
2004b; Bonilla-Findji et al., 2009; Paver & Kent, 2010).
The results presented here suggest that the application of
compost or vermicompost to fields would significantly
increase the amount of organic carbon in the leachates
from soil. Consequently, these compost- or vermicompost-derived leachates are exported to the adjacent aquatic systems and impact their ecology. Moreover, the high
absorption and fluorescence values observed in this leachate mean that not only would DOC concentration
increase but that the chemical quality of that DOM
changes. CDOM is generally considered to be recalcitrant
to bacterial degradation (Rochelle-Newall & Fisher, 2002;
Rochelle-Newall et al., 2004a) and so any inputs to the
lake could probably accumulate in the water column with
potentially negative effects on primary production rates
(Karlsson et al., 2009).
It is now generally accepted that terrestrial DOM is
bioavailable to aquatic bacteria (Cole et al., 2007; Jansson
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
et al., 2007). Here we also show that the organic carbon
in the leachate was highly bioavailable to the aquatic bacterial communities. The rates of removal ranged from
almost 12% to over 27%, comparable with that observed
in other freshwaters (14%; Rochelle-Newall et al., 2004b).
However, those observed in the compost and compost +
biochar treatments were almost twice as high, indicating
its high bioavailability to aquatic bacteria. Moreover, the
shape of the degradation curves (Fig. 1) suggests that
there are several different lability fractions in the bulk
DOC pool (Carlson et al., 1994). The degradation curves
for soil, biochar and vermicompost treatments are linear
in form, indicating that remineralization rates were constant over the 72-h incubation period. In contrast, the
degradation curves of the compost and compost + biochar treatments are curved, indicating that the remineralization rates were not constant. This implies that the
lability of the bulk pool varied with a highly bioavailable
part that was rapidly remineralized during the first 24 h
of incubation, followed by a pool with a lower bioavailability. Although biochar has been proposed to lead to
the protection of organic matter in buffalo manure, vermicompost or compost (Ngo et al., 2013), the leachates
from soils that were supplemented with biochar seem to
have stimulated organic carbon mineralization in the
aquatic incubations. While further investigations are
needed to understand the interaction between biochar
and water, such stimulation may in turn accelerate the
role of tropical lakes as a source of CO2 to the atmosphere rather than a sink (Marotta et al., 2009).
Impacts on bacterial and viral communities
Bacterial and VLP abundance in the incubations also
reflected the addition of the different leachates. As previously observed in similar tropical systems, bacterial and
FEMS Microbiol Ecol 90 (2014) 622–632
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
absorption, also dramatically changed (Table 1). The
addition of biochar to the composted materials had no
significant effect on either the quantity of DOM in leachate or the chemical quality, in terms of fluorescence or
absorption. Interestingly, and in contrast to what has
been previously published for biochar (Lehmann, 2007;
Laird et al., 2010), this work did not appear to support
the idea that biochar can significantly increase the retention of organic matter in soils.
Off-site impact of organic fertilizers on aquatic microbes
Fig. 4. Changes in microbial structure of lake water community
exposed to soil leachates only (S), soil and compost leachates (SC),
soil and vermicompost leachates (SV), soil and biochar leachates (SB),
soil and compost and biochar leachates (SCB), and soil and
vermicompost and biochar leachates (SVB). NMDS representation.
Lowest stress value = 0.17 with R2 = 0.88; using three-dimensional
NMDS, the lowest stress value drops to 0.094 with an R2 = 0.958.
The ellipses indicate the groups of samples included in overlapping
points.
FEMS Microbiol Ecol 90 (2014) 622–632
Delaware contained 22–68% inducible prophages
(Williamson et al., 2007). Although the inducible fraction
has not been evaluated in this study, one might suspect
that the brutal physical and chemical shifts experienced
by soil bacteria could have triggered a lytic cycle within
the community of lysogens. Environmental perturbations
have long been recognized for their inducing properties
on marine lysogens (Weinbauer, 2004; Paul, 2008).
Although the addition of allochthonous C sources
could be anticipated to increase the C availability to the
lake, the observed reduction of metabolic potential of all
communities in all incubations underlines instead the offsite impacts of fertilizer-derived DOM on the metabolic
capacity of aquatic bacteria. Moreover, associated with
the shifts in community composition in the treatments,
we show that the aquatic ecosystems responded strongly
to these additions in terms of both metabolic potential
(Fig. 3) and genetic diversity (Fig. 4).
The explanation as to why the metabolic potential
was so reduced may well lie in the relative concentrations of each DOM molecule. Even if a particular bacterial species possesses the metabolic machinery necessary
to utilize specific compounds, the relative concentrations
of the substrate will also influence whether the compound is metabolized (G
omez-Consarnau et al., 2012).
Given the chemical composition of the compost and
vermicompost used here (Ngo et al., 2011), it is reasonable to expect that the leachate from compost and vermicompost contained high concentrations of a relatively
reduced range of DOM molecules of high bioavailability.
Considering the importance of DOM quantity and quality for structuring microbial communities (Paver &
Kent, 2010; Fortunato & Crump, 2011; G
omez-Consarnau et al., 2012), we can hypothesize that perhaps these
high concentrations of DOM are driving the observed
reduction in metabolic potential and genetic diversity
(Kujawinski, 2011). Indeed, the supply of organic matter, in terms of both quality and quantity, exerts a
strong structural role on aquatic bacteria and the addition of organic matter sources can result in the apparent
specialization under certain conditions (G
omez-Consarnau et al., 2012; Sarmento & Gasol, 2012; Sarmento
et al., 2013).
It has been theoretically proposed that along environmental gradients there is a shift in species nutritional
strategies based on the principle of competitive exclusion:
specialist species that have relatively tightly defined niches
and a narrow range of tolerance are replaced by generalist
species that have broad niches and tolerate larger changes
in the environment (Pimm et al., 1991). Thus specialized
species are expected to be found in more ‘simple’ environments, whereas generalists are more likely to appear in
environments containing a more diverse range of
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
viral abundances were higher in the leachates from soils
amended with organic amendments (Doan et al., 2014).
While the addition of biochar to the soil resulted in a
very small increase in bacterial abundance in the leachates
compared with the leachate from the control soil alone,
the addition of leachate from the compost and
vermicompost treatments resulted in a large increase in
bacterial abundance. However, this was not evident at the
end of the experiment when bacterial abundance in the
compost and vermicompost incubations dropped to levels
similar to that observed in the soil and biochar incubations (Fig. 2 and Table S1). Two equally important
aspects may explain these results. First, terrestrial bacteria
washed out with the leachate did not survive in the lake
water conditions. Secondly, lower bacterial biomass may
result from a stimulation of viral lyses, as, and in contrast
to bacterial abundance, viral abundance increased in all
the incubations. Whether this means that the viruses
introduced in the leachate exhibited some flexibility in
their host system cannot be clearly determined here (Chiura, 1997). Nonetheless, the increase in VLP abundance
observed in all treatments may also result from a massive
induction of lysogens caused by the transfer of bacteria
from a terrestrial to aquatic environment. Lysogeny is
common among soil bacteria; for example, it has been
shown that the bacterial assemblage of different soils in
629
630
Acknowledgements
This work formed part of the research requirement of the
MSc degree of A.M. The French EC2CO programme
‘COMPAQUA’, the Institut de Recherche pour le Developpement, the UMR laboratories iEES and EM, the PHC
Hoa Sen Lotus 23970QM and the LOTUS No. 44/2012/
HD-NDT, MOST, financed this research. The Soil and
Fertilizers Research Institute is thanked for hosting E.R.N.
and T.P. during this work. The authors declare no conflict of interest.
References
Aggelides SM & Londra PA (2000) Effects of compost
produced from town wastes and sewage sludge on the
physical properties of a loamy and a clay soil. Bioresour
Technol 71: 253–259.
Barak P & Laird DA (1997) Effect of longterm soil
acidification due to nitrogen fertilizer inputs in Wisconsin.
Plant Soil 1: 61–69.
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Beesley L & Dickinson N (2011) Carbon and trace element
fluxes in the pore water of an urban soil following
greenwaste compost, woody and biochar amendments,
inoculated with the earthworm Lumbricus terrestris. Soil Biol
Biochem 43: 188–196.
Bonilla-Findji O, Rochelle-Newall E, Weinbauer MG, Pizay
MD, Kerros ME & Gattuso JP (2009) Effect of seawater–
freshwater cross-transplantations on viral dynamics and
bacterial diversity and production. Aquat Microb Ecol 54:
1–11.
Bouvier TC & del Giorgio PA (2002) Compositional changes
in free-living bacterial communities along a salinity gradient
in two temperate estuaries. Limnol Oceanogr 47: 453–470.
Bunge J (2011) Estimating the number of species with
CatchAll. Pac Symp Biocomput, 121–130.
Burle ML, Mielniczuk J & Focchi S (1997) Effect of cropping
systems on soil chemical characteristics, with emphasis on
soil acidification. Plant Soil 190: 309–316.
Carlson CA, Ducklow HW & Michaels AF (1994) Annual flux
of dissolved organic carbon from the euphotic zone in the
northwestern Sargasso Sea. Nature 371: 405–408.
Chen F, Lu J-R, Binder BJ, Liu Y-C & Hodson RE (2001)
Application of digital image analysis and flow cytometry to
enumerate marine viruses stained with SYBR Gold. Appl
Environ Microbiol 67: 539–545.
Chiura HX (1997) Generalized gene transfer by virus-like
particles from marine bacteria. Aquat Microb Ecol 13: 75–83.
Cole JR, Chai B, Farris RJ, Wang Q, Kulam SA, McGarrell
DM, Garrity GM & Tiedje JM (2005) The Ribosomal
Database Project (RDP-II): sequences and tools for
high-throughput rRNA analysis. Nucleic Acids Res 33: D294–
D296.
Cole JJ, Prairie YT, Caraco NF et al. (2007) Plumbing the
global carbon cycle: integrating inland waters into the
terrestrial carbon budget. Ecosystems 10: 171–184.
del Giorgio P & Davis J (2003) Patterns in dissolved organic
matter lability and consumption across aquatic ecosystems.
Aquatic Ecosystems: Interactivity of Dissolved Organic Matter
(Findlay S, & Sinsabaugh RL, eds), pp. 399–424. Academic,
Amsterdam.
Doan TT, Bouvier C, Bettarel Y, Bouvier T,
Henry-des-Tureaux T, Janeau JL, Lamballe P, Nguyen BV &
Jouquet P (2014) Influence of buffalo manure, compost,
vermicompost and biochar amendments on bacterial and
viral communities in soil and adjacent aquatic systems. Appl
Soil Ecol 73: 78–86.
Edgar RC, Haas BJ, Clemente JC, Quince C & Knight R
(2011) UCHIME improves sensitivity and speed of chimera
detection. Bioinformatics 27: 2194–2200.
Fortunato CS & Crump BC (2011) Bacterioplankton
community variation across river to ocean environmental
gradients. Microb Ecol 62: 374–382.
G
omez-Consarnau L, Lindh MV, Gasol JM & Pinhassi J
(2012) Structuring of bacterioplankton communities by
specific dissolved organic carbon compounds. Environ
Microbiol 14: 2361–2378.
FEMS Microbiol Ecol 90 (2014) 622–632
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
resources (Kassen, 2002). The results presented here also
seem to support the hypothesis that DOM bioavailability
and relative concentration play an important role in
determining bacterial community structure and catabolic
capacity (Langenheder & Prosser, 2008; G
omez-Consarnau
et al., 2012) resulting in the emergence of a specialized
community with reduced metabolic potential. How these
changes will ultimately impact lake carbon cycling remains
to be determined, but it is clear that the offsite impacts of
organic fertilization need to be taken into account.
It was striking to note the relatively small effect of the
fertilizer leachates on bacterial richness or taxonomic
composition (at the genus level), while the addition of
biochar deeply affected the community structures with a
strong convergence of similar OTUs. Associated with the
reduction of community metabolic potential, this result
underlines the putative specialization of the bacterial
communities towards types that may be more efficient in
reducing compounds related to biochar. The cascading
effects of such specialization on the microbial loop or
resilience capacity of the lake ecosystem to other environmental stresses need to be evaluated.
Our approach addressed the ecological links between
quantitative and qualitative changes of C compounds
imported with allochthonous DOM and the microbial
component of the aquatic ecosystems. Nevertheless, such
allochthonous inputs also affect the broad stoichiometry
of the impacted aquatic system and their consequences
on the microbial communities remain to be addressed
both in controlled experiments such as the one presented
and in in situ conditions.
T. Pommier et al.
Off-site impact of organic fertilizers on aquatic microbes
FEMS Microbiol Ecol 90 (2014) 622–632
or vermicomposting to rehabilitate degraded tropical soils.
Ecol Eng 37: 269–276.
Ngo P-T, Rumpel C, Ngo Q-A, Alexis M, Vargas GV, Mora
Gil Mdl L, Dang DK & Jouquet P (2013) Biological and
chemical reactivity and phosphorus forms of buffalo
manure compost, vermicompost and their mixture with
biochar. Bioresour Technol 148: 401–407.
Noble RT & Fuhrman JA (1998) Use of SYBR Green for rapid
epifluorescence counts of marine viruses and bacteria. Aquat
Microb Ecol 14: 113–118.
Novak JM, Busscher WJ, Laird DL, Ahmedna M, Watts DW &
Niandou MAS (2009) Impact of biochar amendment on
fertility of a southeastern coastal plain soil. Soil Sci 174:
105–112.
Paul JH (2008) Prophages in marine bacteria: dangerous
molecular time bombs or the key to survival in the seas?
ISME J 2: 579–589.
Paver SF & Kent AD (2010) Temporal patterns in
glycolate-utilizing bacterial community composition
correlate with phytoplankton population dynamics in humic
lakes. Microb Ecol 60: 406–418.
Pimm SL, Lawton JH & Cohen JE (1991) Food web patterns
and their consequences. Nature 350: 669–674.
Quince C, Lanzen A, Davenport RJ & Turnbaugh PJ (2011)
Removing noise from pyrosequenced amplicons. BMC
Bioinformatics 12: 38.
Quinton JN, Govers G, Van Oost K & Bardgett RD (2010)
The impact of agricultural soil erosion on biogeochemical
cycling. Nat Geosci 3: 311–314.
R Development Core Team (2008) A Language and
Environment for Statistical Computing. R Foundation for
Statistical Computing, Vienna.
Rivero C, Chirenje T, Ma LQ & Martinez G (2004) Influence
of compost on soil organic matter quality under tropical
conditions. Geoderma 123: 355–361.
Rochelle-Newall EJ & Fisher TR (2002) Production of
chromophoric dissolved organic matter fluorescence
in marine and estuarine environments: an
investigation into the role of phytoplankton. Mar Chem
77: 7–21.
Rochelle-Newall EJ, Delille B, Frankignoulle M, Gattuso J-P,
Jacquet S, Riebesell U, Terbruggen A & Zondervan l (2004a)
Chromophoric dissolved organic matter in experimental
mesocosms maintained under different pCO2 levels. Mar
Ecol Prog Ser 272: 25–31.
Rochelle-Newall EJ, Pizay MD, Middelburg JJ, Boschker HTS
& Gattuso J-P (2004b) Degradation of riverine dissolved
organic matter by seawater bacteria. Aquat Microb Ecol 37:
9–22.
Rochelle-Newall E, Hulot FD, Janeau JL & Merroune A (2014)
CDOM fluorescence as a proxy of DOC concentration in
natural waters: a comparison of four contrasting tropical
systems. Environ Monit Assess 186: 589–596.
Salles JF, Poly F, Schmid B & Le Roux X (2009) Community
niche predicts the functioning of denitrifying bacterial
assemblages. Ecology 90: 3324–3332.
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
Guenet B, Danger M, Abbadie L & Lacroix G (2010) Priming
effect: bridging the gap between terrestrial and aquatic
ecology. Ecology 91: 2850–2861.
Hamady M, Lozupone C & Knight R (2009) Fast UniFrac:
facilitating high-throughput phylogenetic analyses of
microbial communities including analysis of pyrosequencing
and PhyloChip data. ISME J 4: 17–27.
Howarth R, Swaney D, Billen G, Garnier J, Hong BG,
Humborg C, Johnes P, Morth CM & Marino R (2012)
Nitrogen fluxes from the landscape are controlled by net
anthropogenic nitrogen inputs and by climate. Front Ecol
Environ 10: 37–43.
Jansson M, Persson L, De Roos AM, Jones RI & Tranvik LJ
(2007) Terrestrial carbon and intraspecific size-variation
shape lake ecosystems. Trends Ecol Evol 22: 316–322.
Ka S, Pagano M, Ba N et al. (2006) Zooplankton distribution
related to environmental factors and phytoplankton in a
shallow tropical lake (Lake Guiers, Senegal, West Africa). Int
Rev Hydrobiol 91: 389–405.
Karlsson J, Bystrom P, Ask J, Ask P, Persson L & Jansson M
(2009) Light limitation of nutrient-poor lake ecosystems.
Nature 460: 506–509.
Kassen R (2002) The experimental evolution of specialists,
generalists, and the maintenance of diversity. J Evol Biol 15:
173–190.
Kujawinski EB (2011) The impact of microbial metabolism on
marine dissolved organic matter. Ann Rev Mar Sci 3: 567–599.
Laird DA, Fleming P, Wang B, Horton R & Karlen D (2010)
Biochar impact on nutrient leaching from a Midwestern
agricultural soil. Geoderma 158: 436–442.
Langenheder S & Prosser JI (2008) Resource availability
influences the diversity of a functional group of
heterotrophic soil bacteria. Environ Microbiol 10: 2245–2256.
Le TPQ, Nghiem XA, Luu TNM, Duong TT & Dang DK (2008)
Nutrients (N, P) contents in agricultural wastewater in
Day-Nhue river catchment. Vietnam J Sci Technol 46: 54–61.
Lehmann J (2007) Bio-energy in the black. Front Ecol Environ
5: 381–387.
Louati H, Ben Said O, Got P, Soltani A, Mahmoudi E,
Cravo-Laureau C, Duran R, Aissa P & Pringault O (2012)
Microbial community responses to bioremediation
treatments for the mitigation of low dose anthracene in
marine coastal sediments of Bizerte lagoon (Tunisia).
Environ Sci Pollut Res 20: 300–310.
Marie D, Partensky F, Jacquet S & Vaulot D (1997)
Enumeration and cell cycle analysis of natural populations
of marine picoplankton by flow cytometry using the nucleic
acid stain SYBR Green I. Appl Environ Microbiol 63: 186–
193.
Marotta H, Duarte CM, Sobek S & Enrich-Prast A (2009)
Large CO2 disequilibria in tropical lakes. Global Biogeochem
Cycles 23: GB4022.
Marris E (2006) Putting the carbon back: black is the new
green. Nature (London) 442: 624–626.
Ngo PT, Rumpel C, Dignac M-F, Billou D, Duc TT & Jouquet
P (2011) Transformation of buffalo manure by composting
631
T. Pommier et al.
632
ª 2014 Federation of European Microbiological Societies.
Published by John Wiley & Sons Ltd. All rights reserved
Williams CJ, Yamashita Y, Wilson HF, Jaffe R & Xenopoulos
MA (2010) Unraveling the role of land use and microbial
activity in shaping dissolved organic matter characteristics
in stream ecosystems. Limnol Oceanogr 55: 1159–1171.
Williamson KE, Radosevich M, Smith DW & Wommack KE
(2007) Incidence of lysogeny within temperate and extreme
soil environments. Environ Microbiol 9: 2563–2574.
Supporting Information
Additional Supporting Information may be found in the
online version of this article:
Fig. S1. Schematic representation of the experimental setup.
Fig. S2. Observed (black grey) and estimated by catchall
(light grey) richness in OTU (at 0.03 distance subs. per
nucl.).
Fig. S3. Taxonomic distributions at the genus level of all
communities sampled in the various treatments.
Table S1. Bacterial and viral abundance at T0 and T72 of
incubation of leachate with lake water.
Table S2. Comparison of relative oxidation of the different
biochemical substrates within each treatment.
Data S1. Material and methods.
FEMS Microbiol Ecol 90 (2014) 622–632
Downloaded from http://femsec.oxfordjournals.org/ by guest on May 28, 2016
Sarmento H & Gasol JM (2012) Use of phytoplankton-derived
dissolved organic carbon by different types of
bacterioplankton. Environ Microbiol 14: 2348–2360.
Sarmento H, Romera-Castillo C, Lindh M, Pinhassi J, Sala
MM, Gasol JM, Marrase C & Taylor GT (2013)
Phytoplankton species-specific release of dissolved free
amino acids and their selective consumption by bacteria.
Limnol Oceanogr 58: 1123–1135.
Schloss PD, Westcott SL, Ryabin T et al. (2009) Introducing
mothur: open source, platform-independent,
community-supported software for describing and
comparing microbial communities. Appl Environ Microbiol
75: 7537–7541.
Sidle RC, Ziegler AD, Negishi JN, Nik AR, Siew R &
Turkelboom F (2006) Erosion processes in steep terrain –
truths, myths, and uncertainties related to forest management
in Southeast Asia. For Ecol Manage 224: 199–225.
Valentin C, Agus F, Alamban R et al. (2008) Runoff and
sediment losses from 27 upland catchments in Southeast
Asia: impact of rapid land use changes and conservation
practices. Agric Ecosyst Environ 128: 225–238.
Wang Q, Garrity GM, Tiedje JM & Cole JR (2007) Naive
Bayesian classifier for rapid assignment of rRNA sequences
into the new bacterial taxonomy. Appl Environ Microbiol 73:
5261–5267.
Weinbauer MG (2004) Ecology of prokaryotic reviews. FEMS
Microbiol Rev 28: 127–181.