Journal of Hazardous Materials xxx (xxxx) xxx
Please cite this article as: Martyna Glodowska, Journal of Hazardous Materials, https://doi.org/10.1016/j.jhazmat.2020.124398
Available online 29 October 2020
0304-3894/© 2020 The Authors. Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).
Arsenic behavior in groundwater in Hanoi (Vietnam) inuenced by a
complex biogeochemical network of iron, methane, and sulfur cycling
, Emiliano Stopelli
, Daniel Straub
, Duyen Vu Thi
, Pham T.
, Pham H. Viet
, AdvectAs team members
, Michael Berg
, Andreas Kappler
Geomicrobiology, Center for Applied Geosciences, University of Tübingen, Germany
Microbial Ecology, Center for Applied Geosciences, University of Tübingen, Germany
Eawag, Swiss Federal Institute of Aquatic Science and Technology, Dübendorf, Switzerland
Quantitative Biology Center (QBiC), University of Tübingen, Germany
Key Laboratory of Analytical Technology for Environmental Quality and Food Safety (KLATEFOS), VNU University of Science, Vietnam National University, Hanoi,
AdvectAs members—listed in the SI
UNESCO Chair on Groundwater Arsenic within the 2030 Agenda for Sustainable Development, School of Civil Engineering and Surveying, University of Southern
Editor: Andrew Daugulis
The fate of arsenic (As) in groundwater is determined by multiple interrelated microbial and abiotic processes
that contribute to As (im)mobilization. Most studies to date have investigated individual processes related to As
(im)mobilization rather than the complex networks present in situ. In this study, we used RNA-based microbial
community analysis in combination with groundwater hydrogeochemical measurements to elucidate the
behavior of As along a 2 km transect near Hanoi, Vietnam. The transect stretches from the riverbank across a
strongly reducing and As-contaminated Holocene aquifer, followed by a redox transition zone (RTZ) and a
Pleistocene aquifer, at which As concentrations are low. Our analyses revealed fermentation and methanogenesis
as important processes providing electron donors, fueling the microbially mediated reductive dissolution of As-
bearing Fe(III) minerals and ultimately promoting As mobilization. As a consequence of high CH
methanotrophs thrive across the Holocene aquifer and the redox transition zone. Finally, our results underline
the role of SO
-reducing and putative Fe(II)-/As(III)-oxidizing bacteria as a sink for As, particularly at the RTZ.
Overall, our results suggest that a complex network of microbial and biogeochemical processes has to be
considered to better understand the biogeochemical behavior of As in groundwater.
Arsenic (As) groundwater contamination has been extensively
studied for over two decades, and our knowledge about its mobility and
behavior in the environment has increased substantially in the past
years. Arsenic-bearing sediments deposited in the river delta regions of
South and Southeast Asia have been of particular interest (Postma et al.,
2007; Dowling et al., 2002; Acharyya et al., 2000; Berg et al., 2007).
Reducing conditions in these aquifers led to the mobilization and
enrichment of As in groundwater (Charlet and Polya, 2006; Smedley and
Kinniburgh, 2002). Moreover, these regions are among the most densely
populated areas on the planet (Smedley and Kinniburgh, 2002; Smith
et al., 2000), and many of their inhabitants, mainly situated in rural
areas, still rely on water from shallow wells that is then used either
untreated or ltered through simple sand lters (Nitzsche et al., 2015;
Berg et al., 2006). Because As is “invisible”, it does not affect the taste or
smell of water and food, and, thus, many people have been uncon-
sciously exposed to this toxic metalloid for years. For this reason,
chronical exposure to As has led to massive poisoning throughout local
communities, manifesting as dermal lesions, cardiovascular diseases,
* Corresponding author at: Microbial Ecology, Center for Applied Geosciences, University of Tübingen, Germany.
E-mail address: email@example.com (S. Kleindienst).
Equally contributing authors
Contents lists available at ScienceDirect
Journal of Hazardous Materials
journal homepage: www.elsevier.com/locate/jhazmat
Received 19 May 2020; Received in revised form 30 September 2020; Accepted 24 October 2020
Journal of Hazardous Materials xxx (xxxx) xxx
and various types of cancer (Karagas et al., 2015; Smith et al., 1992).
Many mechanisms for the release of As into groundwater have been
proposed (Islam et al., 2004), yet the most accepted is that As-bearing Fe
(III) (oxyhydr)oxide minerals are reduced and thus dissolved by mi-
croorganisms, a process that is coupled with the oxidation of organic
carbon (Islam et al., 2004, 2005a; Chatain et al., 2005). A large number
of laboratory studies have been conducted in order to underpin micro-
bially mediated Fe(III) mineral reductive dissolution and subsequent As
mobilization. These studies have not only revealed the identity of mi-
croorganisms driving this process, such as Geobacter sp. (Islam et al.,
2005a), Shewanella (Cummings et al., 1999), or Geothrix sp. (Islam et al.,
2005b), but also underline the importance of carbon quantity, quality,
and bioavailability as necessary fuels for Fe(III) mineral reduction
(Chatain et al., 2005; Neumann et al., 2014; Duan et al., 2008; H´
et al., 2010; Glodowska et al., 2020). Nonetheless, additional micro-
bially mediated processes can release As from sediments and contribute
to groundwater contamination. For instance, As(V)-reducing bacteria
can use As(V) as an electron acceptor and reduce it to the more mobile
species of As(III) (Cummings et al., 1999; Islam et al., 2005b; Neumann
et al., 2014; Duan et al., 2008). Amongst others, bacteria belonging to
Geobacter sp. and Sulfurospirillum have been identied to be capable of
dissimilatory As(V) reduction (H´
ery et al., 2008). In addition, isolated
from an As-contaminated aquifer, Enterobacter (ARS-3) has been shown
to release up to 15 µg/L of As from shallow reducing aquifer sediments
via the direct enzymatic reduction of As(V) (Liao et al., 2011).
In contrast, a variety of microbial processes was shown to be a sink
for As in groundwater in laboratory experiments. Iron(II)-oxidizing
bacteria, such as chemoautotrophic Gallionella sp. (Hallbeck and Ped-
ersen, 1990) and heterotrophic Leptothrix sp. (Hashimoto et al., 2007),
were found to catalyze As removal via the precipitation of Fe(III)
(oxyhydr)oxides and the sorption of As onto biogenic Fe minerals
(Katsoyiannis and Zouboulis, 2006; Hohmann et al., 2010). Depending
on the As to Fe ratio, different types of Fe minerals can be produced,
which has been specically shown for nitrate-dependent Fe(II)-oxidizing
Acidovorax (Hohmann et al., 2011). Interestingly, even if Fe(III)
bio-reduction generally leads to As and Fe mobilization into solution,
also different secondary minerals can be formed, such as magnetite,
which can contribute to the re-sequestration of Fe and As (Muehe et al.,
2016). In addition, microbial chemolithoautotrophic or heterotrophic
As(III) oxidation can also be a potential sink for dissolved As, since
microorganisms mediating this process can transform As(III) into the
less mobile As(V), which is more prone to sorption to Fe(III) (oxyhydr)
oxide minerals (Ike et al., 2008; Garcia-Dominguez et al., 2008).
Several studies have been conducted to understand how microbial
reduction affects As mobility. On the one hand, SO
lead to the precipitation of dissolved Fe(II) as iron suldes along with
the incorporation and sorption of As and/or to the direct precipitation of
arsenic trisulde (As
) (Newman et al., 1997; Bostick and Fendorf,
2003). On the other hand, SO
reduction and sulde formation can also
contribute to a reduction in Fe(III) minerals and subsequently either
release some As that was bound to these minerals or directly mobilize As
in the form of thioarsenates (Kumar et al., 2020). As a consequence,
some studies have shown that SO
reduction leads to As removal (Kirk
et al., 2004; Rittle et al., 1995), while others have reported opposite
observations, in which As concentrations in the water increased under
-reducing conditions (Stucker et al., 2014; Kumar et al., 2016; Guo
et al., 2016). The sulde/Fe molar ratio seems to control this behavior
(Kumar et al., 2020). However, a previous study from Red River Delta
aquifers showed that SO
-reducing conditions are associated with net
As immobilization (Sracek et al., 2018), which is likely due to rather low
concentrations and, in consequence, low sulde/Fe ratios.
In addition to microbial mediated Fe, As, and S redox cycles affecting
As (im)mobilization into groundwater, there are additional microbial
processes, such as fermentation, methanogenesis, or methanotrophy,
that have not been the focus in understanding the fate of As. Wang et al.
(Wang et al., 2015) previously suggested an involvement of metha-
nogens in As mobilization. Indeed, in many regions of South and
Southeast Asia, the co-occurrence of high concentrations of As, Fe, and
has been reported (Postma et al., 2007; Harvey et al., 2002; Jessen
et al., 2008; Liu et al., 2009; Polizzotto et al., 2005; Postma et al., 2012).
For example, in southern Bangladesh, CH
, driven from the degradation
of dissolved inorganic carbon (DIC), reached 1.3 mM (21 mg/L), and, at
the same depth, a peak for dissolved As was reported (Harvey et al.,
2002). Similar correlations were found across Bengal, Mekong, and Red
River deltas, where As concentrations were signicantly higher in
methanogenic zones yet signicantly lower in SO
-reducing and Fe
(III)-reducing zones (Buschmann and Berg, 2009). While there seems
to be a link between CH
and As, this relationship still remains elusive.
Moreover, to date, most studies have focused on a single microbial
process under laboratory conditions rather than on the complex network
of several processes co-occurring simultaneously in the eld. Therefore,
here, we used RNA-based 16S rRNA amplicon sequencing for in situ
active microbial taxa in combination with phylogenetic and functional
gene quantication and hydrogeochemical measurements in order to
Fig. 1. Two-dimensional cross-section of Van Phuc aquifers divided by hydrogeochemical zone (A–E), as reported in Stopelli et al. (2020), presenting the distribution
of monitoring wells across the transect. Lighter blue wells are in the proximity to the transect within comparable hydrogeochemical zones. (For interpretation of the
references to color in this gure legend, the reader is referred to the web version of this article.)
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
explain the microbial and biogeochemical processes responsible for the
behavior of As in groundwater at our eld site. In addition, we predicted
the dominating metabolic functions of the in situ active microbial
community using the 16S rRNA amplicon sequencing data to further
corroborate our ndings. The eld site is located in Van Phuc, about
15 km southeast from the capital city Hanoi, Red River delta region,
Vietnam, and it is characterized by zones of distinct hydrogeochemical
conditions accompanied by changing groundwater As concentrations.
Generally, the southeast part of the aquifer consists of strongly reduced
gray Holocene sands and groundwater exceeding the WHO limit of
10 µg/L by a factor of 10–60. The northwest part consists of less reduced
orange Pleistocene aquifer sands, and the groundwater presents As
concentrations below 10 µg/L (Eiche et al., 2008). The transition be-
tween the contaminated and uncontaminated zones is characterized by
changing redox conditions (i.e. a redox transition zone), under which As
mobilization and immobilization seem to co-occur. Several previous
studies investigated the hydrology, geology, lithology and mineralogy of
this site (Eiche et al., 2008; Stopelli et al., 2020; van Geen et al., 2013;
Eiche et al., 2017; Wallis et al., 2020). Moreover, we have recently
provided a detailed description of the site, including the spatial and
temporal evolution of As concentrations in the context of hydro-
geochemical conditions (Stopelli et al., 2020). Yet, none of these studies
assessed the role of microbial communities in As cycling in situ.
With its heterogeneous conditions, this eld site provides ideal
conditions for studying the complex microbial and biogeochemical
network that affects As mobility in groundwater. Therefore, the objec-
tives of the present study are (1) to identify the main active microbial
taxa in situ; (2) to correlate active microbial key taxa with hydro-
geochemical parameters; and (3) to dene the microbial processes and
hydrogeochemical conditions affecting the fate of As in groundwater.
2. Material and methods
2.1. Study area
The study site is in Vietnam and about 15 km southeast from Hanoi in
Van Phuc village, which is situated inside a meander of the Red River
(20◦55’18.7"N, 105◦53’37.9"E). An important feature of the site is an
inversed groundwater ow towards Hanoi city and is caused by a
depression cone due to increased groundwater abstraction in Hanoi. As a
result, water ows in a northwest direction with an estimated velocity of
40 m/year (van Geen et al., 2013). Generally, the studied transect can be
divided into ve main zones (Fig. 1), as described in Stopelli et al.
(2020). Zone A is the riverbank (Red River) in which young sedimentary
deposits are rich in organic matter and As is mobilized (Wallis et al.,
2020). Zone B is located near the riverbank in the Holocene aquifer low
in dissolved and sedimentary organic carbon (OC), in which Fe(III)- and
-reducing conditions are present. In this zone, processes of reduc-
tive As dissolution and sorption/incorporation of As into newly formed
iron- and sulfur phases co-occur simultaneously and seem to be
balanced. Zone C is located further downstream the Holocene aquifer,
where, most likely, an input of OC is occurring. In consequence, meth-
anogenic conditions are present, and As enrichment in the groundwater
is observed. The redox transition zone at which the advection/intrusion
of reduced groundwater from the Holocene aquifer to the Pleistocene
aquifer takes place, is dened as zone D. Here, a decrease of As due to
sorption on and incorporation into Fe(II) and Fe(II)/Fe(III) minerals is
observed, supported by the oxidation of dissolved Fe(II) and the pre-
cipitation of Fe(III) minerals. Finally, zone E is situated in the less
reducing Pleistocene aquifer, in which As concentrations are below
10 µg/L. In total, 18 wells were analyzed for this study; two wells (AMS
15 and 13) were in the proximity of the transect within comparable
hydro(geo)chemical zones and were thus included in the data inter-
pretation (light blue wells in Fig. 1).
2.2. Sample collection and preservation
The groundwater sampling campaign took place in November 2018.
Groundwater samples for hydrogeochemical analyses were collected,
preserved, and analyzed as described previously in detail (Stopelli et al.,
2020). Before sample collection, the groundwater wells were ushed
until the stabilization of O
, and pH and redox potential E
measured using a portable multi-analyzer (WTW 3630). E
normalized to the standard hydrogen electrode (SHE). Trace elements
and cations were determined by Inductively Coupled Plasma Mass
Spectrometry (ICP-MS, Agilent 7500 and 8900), anions by ion chro-
matography (Metrohm 761 Compact IC), dissolved nitrogen (DN) and
dissolved organic carbon (DOC) by a total N and C analyzer (Shimadzu
TOC-L CSH), and NH
by photometry using the indo-
phenol and molybdate methods, respectively. Alkalinity was determined
directly in the eld via titration (Merck Alkalinity Test Kit Mcolortest
11109). Methane was analyzed via gas chromatography (Shimadzu
GC-2014) using the headspace equilibration method (Sø et al., 2018).
For microbial community analysis, samples were also obtained after
the stabilization of O
, pH, and redox potential E
(see above). Water
was collected in 5 L plastic bottles that were ethanol-sterilized and
rinsed with the collected water prior to sampling. Subsequently, the
water was immediately ltered through 0.22 µm pore size sterile
membrane lters (EMD Millipore) using a suction-type lter holder
(Sartorius 16510) connected to a laboratory vacuum pump (Micro-
sart®). In total, 18 wells were sampled, and, from each well, 10 L of
water was ltered. The lters were carefully folded and placed into
sterile Falcon tubes and immersed in LifeGuard Soil Preservation Solu-
tion (Qiagen) in order to stabilize the microbial RNA. Samples were
stored on dry ice during transport and placed in a −80 ◦C freezer upon
arrival at the laboratory.
2.3. DNA and RNA extraction, DNA digestion, reverse transcription, and
DNA and RNA were extracted using a phenol-chloroform method
following a protocol from Lueders et al. (2004). RNA and DNA were
eluted in 50 µL of a 10 mM Tris buffer. DNA and RNA concentrations
were determined using a Qubit® 2.0 Fluorometer with DNA and RNA HS
kits (Life Technologies, Carlsbad, CA, USA). Subsequently, RNA extracts
were digested with the Ambion Turbo DNA-free™ kit, as directed by the
manufacturer (Life Technologies, Carlsbad, CA, USA). Successful DNA
removal was conrmed via 30-cycle PCR using general bacterial primers
(see below). Afterwards, reverse transcription reactions were performed
using a reverse transcriptase (SuperScript™ III), as described by the
manufacturer. Bacterial and archaeal 16S rRNA genes were amplied
using universal primers 515f: GTGYCAGCMGCCGCGGTAA (Parada
et al., 2016) and 806r: GGACTACNVGGGTWTCTAAT (Apprill et al.,
2015) fused to Illumina adapters. The PCR cycling conditions were as
follows: 95 ◦C for 3 min, 25 cycles of 95 ◦C for 30 s, 55 ◦C for 30 s, and
75 ◦C for 30 s. This was followed by a nal elongation step at 72 ◦C for
3 min. The quality and quantity of the puried amplicons were deter-
mined using agarose gel electrophoresis and Nanodrop (NanoDrop
1000, Thermo Scientic, Waltham, MA, USA). Subsequent library
preparation steps and sequencing were performed using Microsynth AG
(Balgach, Switzerland). Sequencing was performed on an Illumina
MiSeq sequencing system (Illumina, San Diego, CA, USA) using the
2×250 bp MiSeq Reagent Kit v2 (500 cycles kit), and between 49,771
and 195,960 read pairs were obtained for each sample. Due to insuf-
cient RNA concentrations in the AMS 12 sample, reverse transcription
was not successful, and, therefore, we performed only DNA-based
analysis for this sample.
2.4. 16S rRNA (gene) sequence analysis
Sequencing data was analyzed with nf-core/ampliseq v1.0.0, which
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
includes all analysis steps and software and is publicly available (Straub
et al., 2020). Primers were trimmed, and untrimmed sequences were
discarded (<4%) with Cutadapt version 1.16 (Martin, 2011). Adapter
and primer-free sequences were imported into QIIME2 version 2018.06
(Bolyen et al., 2018), their quality was checked with demux
(https://github.com/qiime2/q2-demux), and they were processed with
DADA2 version 1.6.0 (Callahan et al., 2016) to eliminate PhiX
contamination, trim reads (before median quality drops below 35; for-
ward reads were trimmed at 230 bp and reverse reads at 207 bp), cor-
rect errors, merge read pairs, and remove PCR chimeras; ultimately, 18,
642 amplicon sequencing variants (ASVs) were obtained across all
samples. Alpha rarefaction curves were produced with the QIIME2 di-
versity alpha-rarefaction plugin, which indicated that the richness of the
samples had been fully observed. A Naive Bayes classier was tted with
16S rRNA (gene) sequences extracted from the SILVA version 132 SSU
Ref NR 99 database (Pruesse et al., 2007), using the PCR primer se-
quences. ASVs were classied by taxon using the tted classier
(https://github.com/qiime2/q2-feature-classier). ASVs classied as
chloroplasts or mitochondria were removed. The number of removed
ASVs was 34, totaling to <0.1% relative abundance per sample, and the
remaining ASVs had their abundances extracted by feature-table
(Pruesse et al., 2007). The abundance table was rareed with a sam-
pling depth of 38,217—the number of minimum counts across sam-
ples—and Bray–Curtis dissimilarities were calculated with q2-diversity
Pathways, i.e. MetaCyc ontology predictions, were inferred with
PICRUSt2 version 2.2.0-b (Phylogenetic Investigation of Communities
by Reconstruction of Unobserved States) (Langille et al., 2013) and
MinPath (Minimal set of Pathways) (Ye et al., 2009) using ASVs and
their abundance counts. Inferring metabolic pathways from 16S rRNA
amplicon sequencing data is certainly not as accurate as measuring
genes by shotgun metagenomics, but it yields helpful approximations to
support hypotheses driven by additional microbiological and biogeo-
chemical analyses (Langille et al., 2013).
The raw sequencing data has been deposited at GenBank under
BioProject accession number PRJNA628856 (https://www.ncbi.nlm.
2.5. Statistical analysis
All statistical analyses were carried out in R (v3.4.4) (R Core Team,
2018). The Bray–Curtis dissimilarity (Sorensen et al., 1948), calculated
via QIIME2, was plotted using Non-metric Multidimensional Scaling
(NMDS) via phyloseq (McMurdie and Holmes, 2013) version 1.22.3.
Environmental variables were tted onto the ordination and denoted by
arrows using vegan version 2.5–1 (Oksanen et al., 2019), and the sig-
nicance of the tted vectors was assessed using 999 permutations of
environmental variables. Spearman rank correlations were determined
between the hydrogeochemical parameters and the relative abundance
of the dominant taxa, and p-values were corrected for multiple testing
using the Benjamini–Hochberg method, yielding a false discovery rate
(FDR) (Benjamini and Hochberg, 1995).
2.6. Quantitative PCR
Quantitative PCRs specic for 16S rRNA genes of bacteria and
archaea, methyl-coenzyme M reductase subunit alpha (mcrA) genes,
particulate methane monooxygenase (pmoA) genes, arsenate reductase
(arrA) genes, and Geobacter sp. genes were performed. The qPCR primer
sequences, gene-specic plasmid standards, and details on the thermal
programs are given in Table S1. Quantitative PCRs on DNA extracts
obtained as described above were performed in triplicate using Sybr-
Green® Supermix (Bio-Rad Laboratories GmbH, Munich, Germany) on
the C1000 Touch thermal cycler (CFX96™ real time system). Each
quantitative PCR assay was repeated three times, with triplicate mea-
surements calculated for each sample per run. Data analysis was done
using the Bio-Rad CFX Maestro 1.1 software version 4.1 (Bio-Rad, 2017).
3. Results and discussion
The microbial diversity based on 16S rRNA (gene) amplicon
sequencing in the groundwater samples correlated signicantly with
groundwater As (p=0.03), CH
(p=0.01) and Mn
(p=0.05). This implies that different concentrations of these
geochemical species were responsible for distinct microbial community
assemblages among the analyzed wells, or, vice versa, the microbial
communities are inuencing the fate of these geochemical species in the
groundwater (Fig. 2). Non-metric Multidimensional Scaling (NMDS)
resulted in the grouping of the wells, which largely reected the
hydrogeochemical zonation proposed by Stopelli et al., (2020). There-
fore, in the forthcoming sections, we follow this zonation and combine
hydrogeochemical data with the analysis of active microbial taxa,
focusing specically on processes that affect As mobilization and
immobilization in situ to explain the behavior of As in each zone.
3.1. Zone A and B: from riverbank sediments to the Holocene aquifer—As
mobilization and transport
River bank sediments (zone A) are a source of dissolved As
(10–508 µg/L As in the porewater, inter-annual average of 100–120 µg/
L (Stopelli et al., 2020)). This zone was also characterized by elevated
concentrations of dissolved organic carbon (DOC) (5.0–6.9 mg C/L) and
dissolved Fe (up to 13 mg/L) (Table 1). These results are in line with the
reactive transport modeling of the Van Phuc aquifers, where the
riverbank-aquifer interface has been identied as a biogeochemical re-
action hotspot and a source of elevated As concentrations (Wallis et al.,
2020; Stahl et al., 2016). This is due to the constant supply of sediments
rich in bioavailable C and reactive Fe(III) (oxyhydr)oxides from the Red
Fig. 2. Non-metric multidimensional scaling (NMDS) plot based on Bray–Curtis
dissimilarity (stress =0.17) to visualize the main biogeochemical groundwater
parameters (arrows for As, Fe, CH
, and Mn) associated with the
microbial community composition. The strength of the interaction is shown by
the length of the arrows; signicant correlations are shown for As (*, p<0.03),
(***, p<0.001), NH
(*, p<0.01) and Mn (*, p<0.05).
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
Hydrogeochemical parameters measured in groundwater wells in the Van Phuc aquifer, Vietnam.
Zone LOQ Riverbank* B C D E
Well ID AMS AMS AMS VPNS VPNS AMS AMS AMS AMS AMS PC AMS PC VPML VPML VPML AMS AMS
12 15 5 5 3 13 11 (25) 11 (32) 11 (47) 32 44 31 43 22 38 54 36 4
Depth m – 0.2–0.5 23–24 23–24 23–24 35–36 25–26 23–24 23–24 30–31 45–46 23–24 36–37 23–24 26–27 20–21 36–37 52–53 23–24 22–23
pH – – 6.42–7.25 6.96 7.04 7.00 7.09 7.22 7.03 7.40 7.09 6.69 7.22 6.96 7.21 7.21 6.74 6.60 6.74 7.04 7.04
mg/L – 2.03–7.05 0.03 0.03 0.05 0.03 0.35 0.62
0.07 0.02 0.08 0.09 0.02 0.05 0.06 0.05 0.04 0.05 0.08 0.03
(SHE) mV – 60–390 34 127 35 21 5 52
18 185 105 8 125 12 18 222 253 62 122 164
mg/L 0.25 1.2–97 28 0.33 <0.25 <0.25 <0.25 <0.25 <0.25 <0.25 0.26 <0.25 4.3 <0.25 <0.25 4.3 6.2 1.7 <0.25 <0.25
mg/L 0.05 1.8–24 5.6 18 4.9 20 32 16 9.8 31 12 13 17 18 26 4.4 4.7 10 28 20
DN mg/L 0.5 1.7–14 0.5 20 61 9.7 5.5 43 22 9.1 0.5 14 0.7 17 14 <0.5 <0.5 0.5 11 11
DOC mg/L 0.5 5.0–6.9 1.2 1.4 8.5 2.6 2.3 7.4 4.4 1.4 1.1 2.6 1.5 3.5 2.5 1.0 1.0 0.9 1.6 1.5
mgN/L 0.01 1.3–15 0.61 23 63 10 5.5 44 25 9.3 0.63 16 0.49 19 15 0.05 0.09 0.60 12 12
mgP/L 0.005 0.005 0.92 0.03 1.8 0.65 0.77 1.4 0.76 0.01 0.32 0.52 0.02 0.52 0.53 0.01 0.01 0.26 0.03 0.02
As µg/L 0.1 15–508 135 22 513 352 337 452 401 0.9 6.2 80 4.3 266 58 1.0 <0.1 6.2 0.6 0.7
As(III) µg/L 0.1 11–450 135 21 489 346 320 416 372 0.3 6.1 76 3.8 262 58 0.3 <0.1 6.0 0.5 0.4
Fe mg/L 0.05 <0.05–14 12 0.54 14 12 20 14 13 <0.05 16 8.9 0.44 10 9.9 0.07 0.07 24 0.75 0.08
Mn mg/L 0.005 3.2–3.9 0.66 1.5 0.15 0.21 0.17 0.16 0.50 1.5 1.0 3.6 2.7 1.0 2.5 2.4 0.28 1.5 1.9 1.1
mg/L 0.02 0.02 1.1 0.09 2.1 0.70 0.89 1.5 0.77 0.06 0.38 0.60 0.06 0.52 0.59 0.03 0.03 0.29 0.06 0.09
mg/L 0.1 0.4–35 11 <0.1 <0.1 <0.1 <0.1 <0.1 <0.1 <0.1 <0.1 <0.1 1.5 <0.1 <0.1 1.5 2.3 0.6 <0.1 <0.1
Si mg/L 1 9–14 14 9 15 11 16 13 10 15 17 8 13 9 9 16 17 18 13 11
Sr µg/L 1 531–537 356 302 478 470 399 397 490 584 231 489 599 475 522 252 198 270 213 360
Br mg/L 0.04 0.16–0.20 <0.04 0.16 0.19 0.08 0.10 0.11 0.18 0.09 0.13 0.10 0.09 0.09 0.11 0.16 0.14 0.19 0.16 0.09
Na mg/L 0.5 5.3–13.4 5.7 19 11 15 13 13 10 14 42 10 17 9.4 9.8 32 34 25 9.4 9.9
K mg/L 0.1 4.7–7.4 2.7 6.6 8.4 3 1.6 6.1 6.0 4.5 3.9 5.0 5.8 5.3 5.0 3.3 2.8 4.2 3.6 4.7
Ca mg/L 0.1 151–181 121 24 92 124 123 64 96 110 30 98 67 100 101 31 21 31 123 108
Mg mg/L 0.01 34–37 27 26 29 30 29 31 33 37 21 26 67 32 34 27 21 32 18 22
Ba µg/L 0.2 269–403 520 469 540 640 352 438 76 342 236 146 110 108 137 102 63 173 48 70
/L 0.1 8.0–12 8.7 5.8 14 11 9.6 11 12 9.7 5.8 9.2 9.8 10 9.3 5.4 4.3 6.3 8.3 8.1
mg/L <0.13 – <0.13 <0.13 53 5 1.9 30 51 <0.13 <0.13 28 <0.13 25 15 <0.13 <0.13 <0.13 <0.13 <0.13
Range from three riverbank samples collected in November 2018.
Parameters affected by sampling: the well was drying quickly and needed several cycles of pumping-rell.
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
River, promoting Fe(III)-reducing conditions. Two wells in zone B
(AMS12 and AMS15) were located in direct proximity to the Red River
in the Holocene aquifer (~200 m downstream of the riverbank; zone B,
Fig. 1). Arsenic (135 and 22 µg/L, respectively) and Fe concentrations
(12 and 0.54 mg/L, respectively) in these two wells differed from each
other. These differences are likely related to the river bank geo-
morphology, as AMS15 lies closer to an erosional meander, while
AMS12 is located on a depositional one, in which more sediments can be
deposited, and, thus, a greater amount of As can subsequently be
released (Stahl et al., 2016). However, the DOC concentrations were
quite similar in both wells (1.2–1.4 mg/L) and lower than in the river-
bank porewater in zone A (5.0–6.9 mg C/L), implying that C consump-
tion was taking place between zones A and B. Wells in zone B present
dissolved As concentrations comparable to the average values in the
riverbank pore water in zone A (100–120 µg/L), suggesting a net
transport of As from the riverbank to the Holocene aquifer. This might
also be related to the co-occurrence of microbial activities leading to a
net balance between As mobilization and immobilization.
Generally, bacterial 16S rRNA gene copy numbers in zone B (Fig. 3)
appeared to be lower (up to 5.0 ×10
/mL), while archaeal
gene copy numbers were relatively high (up to 3.7 ×10
mL) compared to other zones. The observed 16S rRNA gene amplicon
sequencing variants (ASVs) and the alpha diversity indices were highest
in zone B compared to all other zones, suggesting that this zone has the
greatest microbial diversity (Table S2), which might be reected by
diverse microbial processes, which lead to simultaneous As mobilization
and immobilization with the observed limited net change in dissolved As
Among the active microbial community (Fig. 4), fermenters
appeared to be the most abundant group of microorganisms, with the
majority of taxa related to Firmicutes, Chloroexi, and Bacteroidetes
(Kampmann et al., 2012; Gupta et al., 2014). In addition, predicted
fermentation dominated among all environmentally relevant metabolic
pathways, as inferred from 16S rRNA amplicon sequences (Fig. 5). A
variety of organic acids and more bioavailable short-chain fatty acids,
such as acetate, lactate, formate, or propionate, can be produced as a
result of fermentation (McMahon and Chapelle, 1991; Chapelle, 2000).
These fermentation products can fuel reductive dissolution and the
release of As from Fe(III) (oxyhydr)oxides, enhancing reducing (low
redox potential) conditions (Postma et al., 2007; Quicksall et al., 2008).
Low redox potential was shown to play an important role in As dy-
namics, generally favoring the release of As associated with Fe(III)
minerals (Shaheen et al., 2016; LeMonte et al., 2017; Frohne et al.,
2011). Therefore, electron donors provided via fermentation can drive
diverse heterotrophic microbial processes, including methanogenesis,
reduction, and, in particular, Fe(III) reduction.
The riverbank deposits are a source of SO
, impacting the biogeo-
chemistry in zone B. The presence of SO
and saturation indices
pointing toward FeS mineral precipitation (Stopelli et al., 2020) suggest
reduction to sulde (S
) occurs between the riverbank and
the Holocene aquifer, causing S depletion from the solution. Taxa known
to be involved in S-cycling, such as SO
-reducing bacteria which are
afliated with Thermodesulfovibrionia (Maki, 2015), were abundant,
with 4% of the active microbial community in AMS12; these microor-
ganisms may contribute to As immobilization in zone B, where up to
28 mg/L of SO
was measured. Sulde (S
) produced as a result of
reduction can co-precipitate with Fe
greigite, mackinawite, or pyrite, which have a high afnity for As
sorption (Bostick and Fendorf, 2003; Keimowitz et al., 2007;
Huerta-Diaz et al., 1998; Wolthers et al., 2005). Arsenic can also be
precipitated with either S
to form arsenic sulde minerals or with S
to form iron arsenic suldes, such as arsenopyrite (Kirk et al.,
2004; Bostick et al., 2004; O’Day et al., 2004). Thus, microbially driven
reduction in this zone might be a sink for As in groundwater and
diminish its concentration to a certain extent.
Considering the increased concentrations of dissolved Fe and As, it is
not unexpected that Fe(II)-/As(III)- oxidizers thrive in this zone. In both
wells of zone B, Aquabacterium was abundant (6.6% in AMS12 and 12%
Fig. 3. DNA-based quantitative PCR analysis of bacterial 16S rRNA genes, archaeal 16S rRNA genes, Geobacter specic 16S rRNA genes, arsenate reductase genes
(arrA), particulate methane monooxygenase genes (pmoA), and methyl-coenzyme M reductase subunit alpha genes (mcrA) in the groundwater wells of different
zones (B–E) of the Van Phuc aquifer. Error bars show standard deviation from three measurements.
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
Fig. 4. RNA-based 16S rRNA sequence abundance of selected taxa (including potential function) in groundwater samples of Van Phuc. Wells are divided into several
zones according to Stopelli et al. (2020): B: As-transport, C: As-mobilization, D: As-retardation, E: As-pristine/retardation. The wells had a depth between 20 and
27 m, except for those marked otherwise.
Fig. 5. Relative abundance of predicted environmentally relevant microbial metabolic pathways in groundwater wells divided by zone (B–E). Metabolic potential
was inferred from RNA-based 16S rRNA amplicon sequencing data and is based on MetaCyc pathways.
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
in AMS15). Different strains belonging to the genus Aquabacterium were
shown to be capable of Fe(II) oxidation (Zhang et al., 2017; Kalmbach
et al., 1999). Moreover, in many previous studies focusing on
As-contaminated aquifers, Aquabacterium was found abundantly (Sutton
et al., 2009; Li et al., 2014, 2013), implying that this taxon plays an
important role in these reducing aquifers and may contribute to As
immobilization. Arsenic-tolerant Acinetobacter was also highly abundant
in this zone (3% in AMS12 and 6.5% in AMS15). These microorganisms
have been found to be very efcient in As removal from contaminated
soil (Karn and Pan, 2016), and it is therefore possible that these bacteria
also contributed to reducing As concentrations in groundwater sites
from Van Phuc.
was detected in zone B and, in agreement with this nding,
not many microorganisms related to methane cycling were present in
the groundwater samples (Fig. 4). Coherently, mcrA and pmoA genes
related to methane cycling were also less abundant in these wells when
compared to other zones (Fig. 3).
To summarize zone B, carbon degradation and fermentation likely
lead to As mobilization, while SO
reduction, Fe(II) and As(III)
oxidation may promote As immobilization. Consequently, the co-
occurrence of As mobilization and immobilization processes leads to a
net transport of dissolved As from zone A to zone B.
3.2. Zone C: methanogenic Holocene aquifer—further As mobilization
Four wells in the Holocene aquifer in zone C exhibited measurable
concentrations, ranging from 2 mg/L up to 53 mg/L. These wells
were also characterized by low redox potentials (E
+5 to +52 mV,
SHE) and the highest concentrations of both dissolved As (337–513 µg/
L) and dissolved Fe (12–20 mg/L) (Table 1). Moreover, in this zone,
increased DOC values were present (2.3–8.5 mg/L) along with NH
(5.5–63 mg/L). Our previous study showed that the vertical inltration
of aquitard pore water is likely taking place in this zone (Stopelli et al.,
2020). Aquitard sediments contain organic matter intercalations, in
which pore water can be enriched in DOC and percolate into the aquifer
along permeable intercalations (Eiche et al., 2008, 2017). The presence
of additional C can stimulate microbially mediated Fe(III) mineral
reduction and dissolution, thereby contributing to As mobilization. In
fact, the enrichment of putative Fe(III)-reducers, such as Bacillus (up to
4%) and Geobacter (1.5%), was observed in some of the wells in zone C
(Fig. 4). In addition, relatively high Geobacter (up to 9.8 ×10
/mL) and arrA (up to 9.7 ×10
copy numbers were found in most of the wells in zone C (Fig. 3,
Table S3). During Fe(III) reduction, OM-Fe-As complexes within aquifer
sediments may also break up, promoting further C availability and As
enrichment. Supplementary C in this zone likely supports fermentative
processes, which is reected in the high number of sequences afliated
with putative fermenters, with Firmicutes and Bacteroidetes reaching 8%,
and Chloroexi reaching 10% (Fig. 4). Carbon degradation via fermen-
tation probably contributes to the strong reducing conditions as well as
the net As mobilization in this zone. Furthermore, fermentation pro-
cesses provide substrates to fuel methanogenesis, leading to high dis-
concentrations (up to 53 mg/L) (Table 1). The presence of
high concentrations of CH
in zone C is in agreement with a high
abundance of diverse microbial taxa related to the CH
which Methyloparacoccus (20%), Methylomonaceae (13%), Methyl-
omicrobium (9.5%), and Methylomirabilaceae (2.5%) were found to be
dominating (Fig. 4). These results are further supported by high pmoA
and mcrA gene copy numbers in this zone (Fig. 3).
Within the highly reducing conditions of zone C, putative As(III)-
oxidizers, such as Acinetobacter (Karn and Pan, 2016) and Hydro-
genophaga (vanden Hoven and Santini, 1656), were also abundant (up to
10%) (Fig. 4). These microorganisms can cope with high concentrations
of As due to detoxication mechanisms, where As(III) is oxidized by a
periplasmatic enzyme called arsenite oxidase (Chang et al., 2010). This
self-defense mechanism leads to As(III) oxidation and can presumably
decrease As mobility and promote its sorption into sediments. Further-
more, the potential Fe(II)-oxidizer Aquabacterium was highly abundant
(>20%) in well VPNS3, in which the dissolved Fe concentration was
highest (20 mg/L). These microorganisms potentially immobilized part
of the As that was released under highly reducing conditions while
forming As-bearing Fe minerals.
To summarize zone C, the behavior of As is predominantly controlled
by reductive processes that outcompete oxidative ones, since As con-
centrations in Holocene reach 300–500 µg/L. The relation between
and elevated Fe and As in the Holocene aquifer can be
explained by the additional carbon input from aquitard pore water
egression, fueling fermentation and methanogenic metabolisms and
ultimately leading to an increased reduction of As-bearing Fe minerals.
3.3. Zone D: redox transition from Holocene to Pleistocene aquifer—net
The transition between the Holocene and Pleistocene aquifers is
characterized by changing redox conditions, under which both processes
of As mobilization as well as retention have been observed in zone D
(Stopelli et al., 2020). In addition, many abiotic and biotic processes
co-occur simultaneously, which makes this zone hydrochemically and
microbially complex. Generally, shallow wells (23–27 m depth) were
characterized by high dissolved As (58–401 µg/L), Fe (8.9–13 mg/L),
(15–25 mg/L), and CH
(15–51 mg/L) concentrations. Deeper
wells (30–37 m depth), however, had no As, no CH
, rather low NH
(0.49–9.3 mg/L), and almost no dissolved Fe. This is probably due to the
fact that the egression of the aquitard pore water in zone C provides
reducing conditions along the groundwater ow path mainly at
20–30 m, while at, deeper parts of the aquifer in zone D, “Pleistocene”
-like, lesser reducing conditions prevail.
The shallow wells AMS31, AMS32, PC43, and AMS11/25 (23–27 m
deep) showed a similar microbial community composition, which also
corresponded with similar hydrogeochemical conditions. At these
depths, DOC advected with Holocene groundwater from zone C was
more abundant and likely promoted fermentative and methanogenic
processes, as indicated by the high concentrations of CH
are usually related to C degradation. In agreement with the higher C
content, the microbial community composition among all the shallow
wells was generally dominated by fermenters and methanotrophs. All
wells with high CH
concentrations showed an increased abundance of
the microorganisms involved in methanotrophy, which afliated with
Methylomonaceae (10% in AMS31) and Methyloparacoccus (up to 7% in
AMS32 and AMS 11/25; Fig. 4). Although the relative abundance of
methanotrophs was lower in zone D compared to zone C, higher rates of
oxidation was expected in the redox transition of zone D. This is due
to the higher availability of potential electron acceptors at the interface
of the Holocene and Pleistocene aquifers. Several studies have shown
may serve as an electron donor for anaerobic methanotrophs
that couple CH
oxidation with Fe(III) reduction (Ettwig et al., 2016;
Aromokeye et al., 2020; Cai et al., 2018). In fact, our latest study
(Glodowska et al., in press) demonstrated that this process can lead to a
signicant release of As from Fe- and As-bearing Van Phuc sediments, a
process mediated by archaea afliating with Candidatus Methanoper-
edens. However, anaerobic CH
oxidation can also be coupled with SO
reduction, a process driven by the syntrophic consortia of methano-
trophic archaea (ANME-1, ANME-2a,b,c, and ANME-3) and
-reducing bacteria (Boetius et al., 2000; Orphan et al., 2001;
Scheller et al., 2016; Knittel and Boetius, 2009; Milucka et al., 2012),
which can potentially contribute to As immobilization via precipitation.
Finally, a recent study by Leu et al. (2020) revealed that CH
can serve as
an electron donor for members of Methanoperedenaceae, while Mn(IV)
can be used as an electron acceptor. This newly discovered pathway
might be of relevance for As cycling, because Mn(IV) oxides present in
sediments are effective oxidants (Scott and Morgan, 1995) and they can
retain As in sediments; however, once Mn(IV) gets reduced to dissolved
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
, As can be released into the groundwater. Therefore, CH
groundwater owing through Pleistocene sediments containing Fe(III)
and Mn (III/IV) (oxyhydr)oxides as well as the presence of SO
create favorable conditions for anaerobic CH
oxidation in the redox
transition zone. These hydrogeochemical conditions, together with
abundant methanotrophs, strongly imply that complex biogeochemical
interactions between CH
and As are likely taking place in zone D.
Considering the higher abundance of Fe(III) compared to Mn (III/IV)
, Fe(III) is likely preferentially used as an electron acceptor
and, in consequence, contributes to As mobilization. However, it is
important to bear in mind that the reduction of Mn(IV) is thermody-
namically more favorable than the reduction of Fe(III) (Beal et al.,
Furthermore, the wells being screened at different depths allowed for
following the changes in their hydrochemistry and microbial
community across a vertical prole and showed that entirely different
processes seem to occur in deeper parts of zone D. In well AMS11, the
bacterial population increased with depth by one order of magnitude:
from 9.4 ×10
16S rRNA genes per mL at a depth of 25 m
to 3.7 ×10
at a depth of 32 m. At a depth of 25 m, only a
very low relative abundance of microorganisms involved in S-cycling
was identied, whereas, at a depth of 30 m, the groundwater was
dominated by taxa related to Tumebacillus (19.5%). Tumebacillus has
been previously shown to grow chemolithoautotrophically on inorganic
sulfur compounds, such as sodium thiosulfate and sulte, as sole elec-
tron donors (Steven et al., 2008). Furthermore, SO
such as Thermodesulfovibrionia (8%) and Desulfarculaceae (3%), were
abundant (Fig. 4). In agreement, the predicted S-oxidation pathway
appeared to be more abundant at deeper parts of zone D (Fig. 5). The
presence of microorganisms and predicted metabolic functions involved
Fig. 6. Heatmap of Spearman rank correlations of microbial taxa with hydrogeochemical parameters, such as dissolved Fe, As, CH
and DOC (elevated
concentrations more characteristic of Holocene aquifers) and E
and Mn (elevated concentrations more characteristic of Pleistocene aquifers), with the most
abundant taxa clustered by their putative functions. Signicance levels (Benjamini–Hochberg corrected): p<0.1 (*), p<0.05 (**), p<0.001(***).
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
in S-cycling corroborates the hydrogeochemical data, as SO
detected in the deepest wells (0.26 and 4.2 mg/L, respectively)
(Table 1). Additionally, framboidal pyrites and other iron sulde min-
erals such as mackinawite were previously observed in sediments
(Kontny et al., unpublished), further implying that active S-cycling is
taking place at the redox transition zone.
The composition of the in situ active microbial community (Fig. 4)
implies that oxidative processes are also occurring in zone D. An
increased abundance (up to 7%) of putative NO
dizers afliating with Acidovorax was observed. These microorganisms
were previously found in As-contaminated aquifers (Sutton et al., 2009;
Straub et al., 2004), and they likely contribute to As removal through
sorption or incorporation into freshly formed Fe(III) minerals, as was
shown in a laboratory study (Hohmann et al., 2011). Furthermore, in
some of the wells, taxa that afliate with putative NH
cadiaceae (mainly Candidatus Brocadia), (Jetten et al., 2015) were found
abundantly, although NH
was present in all wells (concentrations
ranging from 0.49 to 25 mg/L). In fact, Brocadiaceae accounted for up to
15% of the microbial community in well PC44, in which NH
trations were very low (0.49 mg/L), implying that decreased NH
centrations were inuenced by the activity of these microorganisms.
Generally, a retardation effect on As advection was observed in zone
D (redox transition zone). However, the retardation capacity seems to
depend on several microbial processes that can either contribute to the
release of As from sediments, such as fermentation and methanotrophy,
or retain it in sediments, such as Fe(II) oxidation and SO
Besides biological processes, abiotic processes can also largely inuence
As concentrations in this zone. High dissolved Mn concentrations were
measured in the groundwater samples (Table 1), indicating that abiotic
Fe(II) oxidation by Mn(IV) reduction can contribute to As immobiliza-
tion in zone D. Therefore, the signicant adsorption and incorporation
of dissolved As into Fe(III) minerals—both newly formed and already
present in Pleistocene sediments—are decreasing As groundwater con-
centrations (Stopelli et al., 2020). Furthermore, as an effect of meth-
can be produced, which, in addition to Fe
owing from the Holocene aquifer, could lead to the pre-
cipitation of Fe(II) carbonate and subsequent As sorption (Eiche et al.,
2008; Sø et al., 2018).
3.4. Zone E: Pleistocene pristine aquifer—As immobilization
Dissolved As concentrations generally remained below the WHO
guideline value of 10 µg/L in the Pleistocene aquifer (zone E). The wells
in this zone were hydrogeochemically similar, presenting higher redox
potentials from +122 to +253 mV (SHE), low dissolved As
(<0.1–1µg/L), and Fe (0.07–0.75 mg/L) as well as dissolved CH
below the detection limit (Table 1).
The in situ active microbial community in zone E was abundant in
putative S-oxidizing bacteria mainly associated with Sulfuricurvum
(10%), Sulfuritalea (5%), and Tumebacillus (10%). These bacteria were
particularly abundant in well VPML38, in which elevated SO
centration (6.2 mg/L) were found among samples in zone E. Moreover,
the highest relative abundance of a predicted S oxidation pathway was
inferred from 16S rRNA amplicon sequences in this zone (Fig. 5). These
ndings suggest that active S-cycling is taking place in zone E and,
similarly to the deeper part of zone D, likely contributes to As sorption to
FeS minerals and, thus, probably promotes As removal from
Although no CH
was detected in these wells, AMS36 was dominated
by a methanotrophic taxon related to Methylomirabilaceae (12%), while
VPML22 was dominated by Methylomonaceae (23%). The high abun-
dance of the pmoA gene further suggests that CH
oxidation might take
place in this zone and contributes to a complete consumption of CH
transported from the Holocene aquifer, ultimately representing a cryptic
cycle similar to other cryptic cycles that have been demonstrated
previously (Kappler and Bryce, 2017).
Finally, in well VPML38, potential Fe(II)-oxidizing bacteria, partic-
ularly those afliating with Aquabacterium, represented more than 15%
of the active microbial community. In addition, in the groundwater
samples of other wells, increased abundances of Fe(II)-oxidizers related
to Gallionellaceae (4.6%) and Acidovorax (4.3%) were observed (Fig. 4).
These microorganisms presumably contribute to As immobilization
through the precipitation of Fe(III) minerals and a co-precipitation of As.
To summarize, zone E is generally dominated by oxidative processes,
efciently maintaining low As concentrations in groundwater.
It is important to note that wells AMS11/47 in zone D and well
VPML54 in zone E were screened at further depths (46 m and 53 m,
respectively) compared to other wells. These wells reach a Pleistocene
gravel layer underlying the sandy aquifers (Eiche et al., 2008; van Geen
et al., 2013) and are characterized by high dissolved Fe (16 and
24 mg/L) but low As concentrations (6.2 µg/L; Table 1). At these depths,
generally, processes that can lead to As immobilization seem to prevail.
In AMS11/47, SO
reduction was indicated by a high abundance of
Thermodesulfovibrionia (8%) and Desulfarculaceae (3%) (Fig. 4) as well as
a trace concentration of SO
(0.26 mg/L) (Table 1), while putative Fe
(II)-oxidizers, such as Gallionellaceae (5%) and Aquabacterium (3%),
and As(III)-oxidizers, such as Acinetobacter (3%), were dominating in
3.5. Fermentation, CH
cycling, microbially mediated Fe(III), and SO
reduction dominate aquifer biogeochemistry
Diverse active microbial taxa and predicted metabolic pathways
were identied in groundwater samples across Van Phuc aquifers. Fer-
menting microorganisms were the most abundant and omnipresent
group. The correlation of active taxa with hydrogeochemical parameters
(Fig. 6) indicated that fermenting microorganisms are ubiquitous and
can adapt both to Holocene and Pleistocene aquifer conditions. There-
fore, fermentation seems to be a key biogeochemical process across the
aquifers of Van Phuc. Pyruvate fermentation was the most common
predicted fermentation pathway followed by homolactic, mixed acid,
and heterolactic fermentation (Fig. S1). As a result, a wide range of
short-chain fatty and organic acids are likely produced, including ace-
tate, lactate, formate, or propionate (McMahon and Chapelle, 1991;
Chapelle, 2000). Thus, at our eld sites, fermentation may provide
easily bioavailable C compounds that further fuel diverse heterotrophic
microbial processes, including reductive dissolution and As release from
Fe(III) (oxyhydr)oxides (Postma et al., 2007; Quicksall et al., 2008).
Despite its high abundance in many As-contaminated aquifers
remains largely unexplored when discussing its poten-
tial role in As (im)mobilization. Positive correlations between dissolved
, Fe, and, by consequence, dissolved As have been previously re-
ported (Postma et al., 2007; Dowling et al., 2002; Liu et al., 2009; Sutton
et al., 2009), suggesting that methanogenesis can indirectly promote Fe
(III) mineral reduction and As mobilization by providing CH
electron donor. Methanogenesis is directly fueled by fermentation
products, and, in fact, all substrates necessary for CH
as acetate, CO
, can be produced during fermentation (Alibardi
and Cossu, 2016). Surprisingly, very high concentrations of CH
of the analyzed wells (up to 53 mg/L) did not correspond to the high
relative abundances of methanogens nor mcrA genes. In fact, known
methanogens accounted for as little as <1% in all wells. A recent study
showed that the methanotrophic population is mainly found in sedi-
ments rather than in water (Kuloyo et al., 2020), which is likely also true
for methanogens. For this reason, we decided to explore the presence of
methanogenic microorganisms in sediment samples, where they
accounted for as much as 60% of the microbial community (Glodowska
et al., unpublished). Despite the low abundance of methanogenic
archaea in groundwater, the analysis of the main predicted metabolic
pathways suggested that two types of methanogenesis take place in the
aquifers: rst, acetoclastic methanogenesis, where the main precursor is
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Journal of Hazardous Materials xxx (xxxx) xxx
COOH → CO
and second, hydrogenotrophic methanogenesis:
is reduced to CH
(Goevert and Conrad, 2009). Metabolic
pathways inferred from 16S rRNA amplicon sequences (Fig. 5) suggested
that the predicted acetoclastic methanogenesis is dominating in Van
Phuc aquifers (Fig. S2).
3.5.1. Methane cycling linked to As mobilization
Methane can be used as electron donor to fuel a wide range of
microbially mediated processes, such as reductions of SO
et al., 2020; Cai et al., 2018; Boetius et al., 2000; Orphan et al., 2001;
Scheller et al., 2016), NO
(Haroon et al., 2013), NO
(Ettwig et al.,
2010), and Mn(IV) (Leu et al., 2020). Most importantly, for our eld site,
oxidation can also be coupled with Fe(III) reduction
(Ettwig et al., 2016; Aromokeye et al., 2020; Cai et al., 2018), a process
that we conrmed within Van Phuc sediments and that can lead to
signicant As mobilization (Glodowska et al., in press). Furthermore,
microorganisms with the metabolic potential for CH
abundantly present in most of the sampled wells (Figs. 3 and 4). This
might explain the large variability in the CH
in the groundwater samples in Van Phuc, ranging from <0.13 mg/L to
53 mg/L. Methanotrophic communities at our eld site are related to C
degradation, as we found a strong correlation between methanotrophic
bacteria and C degradation products, such as CH
, DOC, and NH
(Fig. 6). Some of these taxa are also signicantly positively correlated
with As, e.g. Methylococcus (r =0.65), Methylocaldum (r =0.64) or
Beijerinckiaceae (r =0.63) (Fig. 6), which suggests their direct involve-
ment in As mobilization. Thus, the correlation analysis conrmed our
observation that microorganisms mediating CH
cycling were mainly
active under conditions typical of Holocene aquifers and redox transi-
tion zones and less active in Pleistocene Mn-reducing aquifers, where
they were negatively correlated with E
, and dissolved Mn
3.5.2. Low abundance of known Fe(III)-reducers
Active microbial taxa known to be involved in dissimilatory Fe(III)
reduction, such as Bacillus, Deferribacteres, Geobacter, Thermincola, Geo-
thrix, and Magnetospirillum were ubiquitous across the whole eld site,
although at rather low relative abundances (Fig. 4). Many previous
studies have shown the importance of Fe(III)-reducing bacteria in As
mobilization (Islam et al., 2005a, 2005b; H´
ery et al., 2008; Kim et al.,
2012; Ohtsuka et al., 2013). Nonetheless, in most of these studies, the in
situ abundance of known Fe(III)-reducers was generally quite low (H´
et al., 2008; Li et al., 2013; Kim et al., 2012). Van Phuc’s groundwater
samples also showed a lower abundance of known Fe(III)-reducers,
particularly when compared to dominant fermenters and CH
microorganisms. This data suggests that either many unknown micro-
organisms in the community are capable of reducing Fe(III) or various
microorganisms, such as methanotrophs, that have not been previously
considered can actually contribute to Fe(III) reduction and As mobili-
zation to a larger extent than known Fe(III)-reducers. Our previous
study, where natural organic matter was used as an electron donor,
showed that diverse microorganisms contributed to Fe(III) reduction
(Glodowska et al., 2020). However, when easily bioavailable C was
added (acetate and lactate), mainly Geobacter was responsible for the
reductive dissolution of sedimentary Fe(III) minerals. This strongly im-
plies that a much larger microbial community than is currently known is
involved in the reduction of Fe(III) minerals under natural conditions.
3.5.3. Sulfate reduction as a sink for As
The predicted genetic potential for SO
reduction and S oxidation
appeared to be equally present in all wells (Fig. 5). However, RNA-based
16S rRNA amplicon sequencing data showed that taxa identied as
reducers or S oxidizers were particularly abundant only
in some of the wells (Fig. 4), mainly in those where dissolved S species
(dominated by SO
) were detected. Our hydrogeochemical data
showed that the majority of wells for which SO
was reported were also
characterized by low As concentrations, supporting our hypotheses that
microbially mediated S cycling maintains low As concentrations in
groundwater samples and that microbial SO
reduction (leading to
sulde production) is generally a sink for As in Van Phuc. This process
seems to be particularly relevant at the redox transition zone and in the
Pleistocene aquifer, where taxa involved in S-cycling appeared to be
negatively correlated with dissolved Mn (Fig. 6).
3.5.4. Cryptic N-cycle
Interestingly, pathways for predicted nitrate (NO
) reduction and
nitrier denitrication were also inferred in all wells (Fig. 5). At the
same time, active microorganisms involved in the N-cycle were identi-
ed as NH
-oxidizers in some of the wells (Fig. 4), which are mostly
associated with Pleistocene-like moderate reducing conditions (Fig. 6).
Nitrier denitrication is a pathway in which NH
is oxidized to NO
which is subsequently reduced via NO and N
O to N
(Wrage et al.,
2001). This metabolic pathway is found in some autotrophic
-oxidizers as well as in many CH
-oxidizing bacteria (Stein and
Klotz, 2011); therefore, its presence is likely related to methanotrophs,
which were abundant in the microbial community. In fact, while our
hydrogeochemical analyses did not show the presence of any measur-
able N species other than NH
, our microbiological analyses indicated
active N cycling. Such cycling is likely happening rapidly, therefore
hampering the hydrogeochemical measurement of N species on top of
, and could be responsible for the large variability in NH
centrations in the groundwater samples. The implications of the N-cycle
on As (im)mobilization, for instance, via nitrier denitrication and
microbially mediated NH
oxidation coupled with Fe(III) reduction
(Feammox), deserve further investigation. In fact, Xiu et al. (2020)
proposed Feammox to be one of the mechanisms involved in As release
in the western Hetao Basin.
3.5.5. Unexplored role of Mn in As immobilization
Finally, dissolved Mn appeared to be one of the decisive hydro-
geochemical parameters affecting the active microbial community in
Van Phuc aquifers (Fig. 2 and Fig. 6), yet, we could not identify mi-
croorganisms known to be directly involved in the Mn cycle. However,
taxa afliating with Gemmataceae (r =0.65, p<0.1) and Haliangium
(r =0.47, p<0.3) showed a weak positive correlation with Mn (Fig. 6),
suggesting their potential involvement in the Mn cycle or associated
processes. Moreover, a pathway of CH
oxidation coupled with Mn(IV)
reduction has been recently described (Leu et al., 2020). This process
might be relevant for the As mobilization reactions at the redox transi-
tion zone, considering the abundance and diversity of methanotrophs,
the high concentration of dissolved CH
, and the presence of Mn(IV)
oxides within the sediments in this zone.
Understanding the interactions between microbiota and their
hydrogeochemical environment is key in unraveling the biogeochemical
network affecting As (im)mobilization. Our study shows that fermen-
tation and methanogenesis, which have been largely overlooked in this
context, are among the most important microbiological processes,
indirectly favoring As mobilization. In addition, fermentation provides
bioavailable C that further fuels various microbial metabolisms.
Following methanogenesis, CH
is most likely oxidized by the reduction
of various electron acceptors, such as As-bearing Fe(III) minerals that
are especially abundant in the Pleistocene sediments. Methane oxidation
is a process that has not been linked to the As cycle previously; however,
a rich community of active CH
oxidizers was present in all zones of the
studied aquifers, where it may contribute to As mobilization when
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
coupled with Fe(III) and Mn(IV) reduction or to As immobilization
coupled with SO
At the same time, a number of oxidative metabolic processes co-exist
and act as potential sinks for dissolved As, such as Fe(II) and/or As(III)
oxidation. Moreover, the S cycle appears to be closely interlinked with
dissolved As behavior. The presence of active SO
and S metabolizers at
the redox transition and in the Pleistocene aquifer is related with lower
As concentrations in groundwater samples. The formation of FeS, FeAsS,
and AsS minerals is a result of an active microbially mediated S cycle and
is likely responsible for the sorption and incorporation of As into these
These As (im)mobilization processes do not occur separately. In re-
ality, complex biogeochemical interactions co-exist simultaneously and
ultimately inuence the groundwater’s dissolved As concentrations.
This is particularly relevant for the biogeochemistry taking place at
redox transition zones, where fermentation, methanotrophy, SO
reduction, S oxidation, and Fe(II) oxidation may occur together (Fig. 7).
Therefore, only by linking microbial and hydrogeochemical pro-
cesses might we be able to explain the fate of dissolved As concentra-
tions in both contaminated and pristine aquifers and especially its
retardation across redox transition zones between Holocene and Pleis-
tocene aquifers. We believe that our observation can be widely trans-
ferable and helps to understand As behavior in the context of microbially
mediated processes in similar aquifers in South and Southeast Asia,
including the other young Holocene aquifers of the Red River Delta.
CRediT authorship contribution statement
Martyna Glodowska: Conceptualization, Investigation, Writing -
original draft, Writing - reviewing & editing, Visualization. Emiliano
Stopelli: Conceptualization, Investigation, Writing - reviewing & edit-
ing. Daniel Straub: Data curation, Formal analysis. Duyen Vu Thi,
Pham T.K. Trang, Pham H. Viet, AdvectAs team members: Re-
sources, Project administration. Michael Berg: Supervision, Writing -
reviewing & editing, Project administration, Funding acquisition.
Andreas Kappler: Supervision, Writing - reviewing & editing, Funding
acquisition. Sara Kleindienst: Supervision, Writing - reviewing &
editing, Funding acquisition.
Declaration of Competing Interest
The authors declare that they have no known competing nancial
interests or personal relationships that could have appeared to inuence
the work reported in this paper.
We thank the Deutsche Forschungsgemeinschaft (DFG) and the Swiss
National Science Foundation SNF for funding the AdvectAs project
through DACH (grant # 200021E-167821). Sara Kleindienst is funded
by the Emmy-Noether fellowship (grant # 326028733) from the DFG.
Daniel Straub is funded by the Institutional Strategy of the University of
Tübingen (DFG, ZUK 63) and the Collaborative Research Center CAM-
POS (Grant Agreement SFB 1253/1 2017). We are grateful to C. Stengel
and N. Pfenninger for the technical assistance during the ICP-MS ana-
lyses. Thanks are also extended to the entire AuA team, Eawag, for their
analyses and to R. Britt, M. Brennwald, Anh Lang T., and Thanh Nguyen
V. for their help during the eld campaign in November 2018. Our
gratitude also goes to the citizens of Van Phuc for supporting eldwork
in the village. The authors acknowledge support by the High Perfor-
mance and Cloud Computing Group at the Zentrum für Datenver-
arbeitung of the University of Tübingen, the state of Baden-
Württemberg, through bwHPC and the German Research Foundation
(DFG) through grant no INST 37/935-1 FUGG.
Appendix A. Supporting information
Supplementary data associated with this article can be found in the
online version at doi:10.1016/j.jhazmat.2020.124398.
Fig. 7. Two-dimensional cross-section of Van Phuc aquifers divided by hydrogeochemical zone (A–E), as reported in Stopelli et al., 2020. For each zone, the main
microbial processes are indicated together with their net effect on As mobilization (As↑) or immobilization (As↓).
M. Glodowska et al.
Journal of Hazardous Materials xxx (xxxx) xxx
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