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Unravelling Light and Microbial Activity as Drivers of Organic Matter Transformations in Tropical Headwater Rivers

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Connecting tropical rainforests to larger rivers, tropical headwaters export large quantities of carbon and nutrients as dissolved organic matter (DOM), and are thus a key component of the global carbon cycle. This DOM transport is not passive, however; sunlight and microbial activity alter DOM concentrations and compositions, affecting riverine greenhouse gas emissions and downstream ecosystems. The effects of sunlight and microbial turnover/activity on DOM concentrations and compositions in tropical headwaters are currently poorly understood, but novel Size Exclusion Chromatography (SEC) techniques coupled to suitable detectors can for the first time quantify their influences. Here, we present in-situ incubation experiments from from headwaters of the Essequibo River, in the Iwokrama Rainforest, Guyana, where sunlight oxidised up to 9% of dissolved organic carbon (DOC) over 12 hours, at higher rates than in larger tropical rivers. DOM transformations occurred in both photo-sensitive and supposedly photo-resistant pools. Microbial activity had varying, less clear influences on DOC concentrations over the same time span; compositionally, this appeared to extend beyond known bio-labile components. Biopolymers were particularly reactive to both processes. We show sunlight has the greater potential to mineralise headwater DOM and thus potentially influence degassing. Our approach provides a future template to constrain DOM transformations along river networks, identify biogeochemical activity hotspots, and improve greenhouse gas emissions estimations under changing environmental conditions.
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Unravelling Light and Microbial Activity as Drivers of Organic
Matter Transformations in Tropical Headwater Rivers
James F. Spray1*, Thomas Wagner1, Juliane Bischoff1, Sara Trojahn1, Sevda Norouzi1, Walter Hill1,
Julian Brasche2, Leroy James2, Ryan Pereira1*
1The Lyell Centre, Heriot-Watt University, EH14 4AP, Edinburgh, UK. 5
2Iwokrama International Centre for Rainforest Research and Development, 77 High Street, Kingston Georgetown, Guyana
Correspondence to: James F. Spray ( and Ryan Pereira (
Abstract. Connecting tropical rainforests to larger rivers, tropical headwaters export large quantities of carbon and nutrients
as dissolved organic matter (DOM), and are thus a key component of the global carbon cycle. This DOM transport is not
passive, however; sunlight and microbial activity alter DOM concentrations and compositions, affecting riverine greenhouse 10
gas emissions and downstream ecosystems. The effects of sunlight and microbial turnover/activity on DOM concentrations
and compositions in tropical headwaters are currently poorly understood, but novel Size Exclusion Chromatography (SEC)
techniques coupled to suitable detectors can for the first time quantify their influences. Here, we present in-situ incubation
experiments from from headwaters of the Essequibo River, in the Iwokrama Rainforest, Guyana, where sunlight oxidised up
to 9% of dissolved organic carbon (DOC) over 12 hours, at higher rates than in larger tropical rivers. DOM transformations 15
occurred in both photo-sensitive and supposedly photo-resistant pools. Microbial activity had varying, less clear influences on
DOC concentrations over the same time span; compositionally, this appeared to extend beyond known bio-labile components.
Biopolymers were particularly reactive to both processes. We show sunlight has the greater potential to mineralise headwater
DOM and thus potentially influence degassing. Our approach provides a future template to constrain DOM transformations
along river networks, identify biogeochemical activity hotspots, and improve greenhouse gas emissions estimations under 20
changing environmental conditions.
1 Introduction
Global estimates of terrestrially-derived carbon to inland waters (rivers, lakes, and wetlands) range from ~5.1-5.7 petagrams
of carbon per year (Pg C yr-1) (Drake et al., 2018; Wehrli, 2013). However, only 0.9 Pg C yr-1 is estimated to reach the ocean;
~3.9 Pg C yr-1 is emitted to the atmosphere as carbon dioxide (CO2) (Battin et al., 2008; Cole et al., 2007; Drake et al., 2018). 25
Wetlands (in particular tropical wetlands) represent the largest non-anthropogenic source of methane (CH4); they comprise 20-
30% of total CH4 emissions (Nisbet et al., 2014; Bousquet et al., 2011). In the context of the global carbon cycle, this CO2
evasion is estimated to be on the same order of magnitude as anthropogenic fossil fuel CO2 emissions (7.9 ± 0.5 Pg C yr−1),
deforestation, and changes in land use (1.0 ± 0.7 Pg C yr−1) (Ward et al., 2017). However, there is significant uncertainty over
its sources and controlling mechanisms (Raymond et al., 2013). One key source of this uncertainty is the extent that carbon 30
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and nutrients from organic sources, known as organic matter (OM), flow through inland waters ‘passively’ (without utilization
or transformation) or ‘actively’ (utilized and/or transformed through biotic and abiotic processes). The river continuum
(Vannote et al., 1980) and flood-pulse-shunt concepts (Junk, 1989; Raymond et al., 2016) provide useful frameworks to
encapsulate passive and active land-to-ocean OM transport processes. Both mechanisms likely interchange during key
moments and at hotspot locations along networks of inland waters (McClain et al., 2003), defining focal points of interest and 35
further research. It therefore remains a challenge to constrain the rate of OM transformations at different/sliding temporal and
spatial scales, and the associated release of greenhouse gases.
Rainforests are a major hotspot of the global carbon cycle; they store and release vast quantities of OM into large tropical river
systems, such as the Amazon, Congo, Orinoco, or Essequibo (Battin, 1998; Pereira et al., 2014a; Richey et al., 2002; Spencer
et al., 2012). Indeed, the annual flux of dissolved organic carbon (DOC) from tropical rivers to the ocean alone accounts for 40
~59% of global riverine DOC flux (Huang et al., 2012). However, of the 470 Tg C yr-1 of CO2 emitted to the atmosphere by
the Amazon river alone, up to 75% is estimated to derive from OM of near-channel origin, which has been mobilized to, and
mineralized within, the aquatic zone (Davidson et al., 2010; Richey et al., 2002). Tropical river systems may also be significant
sources of atmospheric CH4 emissions (Upstill-Goddard et al., 2017). Globally, headwaters account for 70-80% of total river
networks (Gomi et al., 2002). Furthermore, first- and second-order streams comprise over 80% of the total channel length of 45
meso-scale Amazon drainage basins (McClain and Elsenbeer, 2001). Therefore, it is necessary to better constrain the rates of
biotic and abiotic transformations of OM in headwaters, where it crosses the terrestrial-aquatic interface.
Headwaters are proximal to terrestrial OM inputs, meaning their components, such as dissolved OM (DOM; which includes
DOC and dissolved organic nitrogen (DON)), exhibit large and variable fluxes that directly relate to their immediate
surroundings and environmental conditions (Pereira et al., 2014a; Voss et al., 2015). The concentration and composition of 50
DOM in such headwaters reflects both hydro-climatic changes in the relevant ecosystem and the compositional diversity of
carbon pools in the surrounding soils and vegetation (Battin, 1998; Mann et al., 2014; Seidel et al., 2015; Spencer et al., 2016).
These factors influence the biotic and abiotic lability or recalcitrance of DOM components entering the rivers, and will likely
control their fate, through degassing, transport, and burial (Amon, 2002).
Two of the main processes that influence, and are influenced by, the concentration and composition of OM in rivers are 55
photodegradation and microbial activity/degradation (biodegradation) (Cory and Kling, 2018; Jones et al., 2016; Mostofa et
al., 2013a; Obernosterer and Benner, 2004; Wiegner and Seitzinger, 2001). Photodegradation refers to the process whereby
ultraviolet (UV) and visible light breaks down OM. Some DOM compounds contain chromophores, defined as functional
groups in organic compounds that can absorb photons with an efficiency sufficient to promote an electron from a ground to an
excited state (Mostofa et al., 2013a). Chromophoric structures that are comprised of conjugated double bonds can be broken 60
down upon absorbing this radiation, leading to the photodegradation of the molecule in question (Jones et al., 2016; Zepp,
1988). This process is thought to mainly effect Coloured DOM (CDOM). However, there is debate over the extent that
photodegradation results in the complete mineralization of DOM to CO2. It may only achieve partial degradation, breaking
down coloured, high molecular weight (HMW) compounds into low molecular weight (LMW) components that are potentially
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more labile (Cory et al., 2014; Jones et al., 2016; Lou and Xie, 2006; Mostofa et al., 2013b; Obernosterer and Benner, 2004; 65
Remington et al., 2011). In addition to CO2 outgassing, the nitrogenous photoproducts ammonium (NH4+), nitrate (NO3-), and
nitrite (NO2-) are also released by the photodegradation of humic substances and DON (Mostofa et al., 2013a); these
compounds are critical for biological productivity (Zepp, 2003). Microbiological cycling of OM represents an
alternative/competing process that also alters the abundance and composition of DOM (Mostofa et al., 2013b). The process of
biodegradation comprises the consumption of DOM compounds by microbes, leading to DOM mineralization and CO2 70
production via respiration (Cory et al., 2014). LMW aliphatic DOM is more easily broken down by microorganisms than
larger, more aromatic and complex molecules (Jones et al., 2016).
It has been posited that these two processes can combine through the mechanism of photo-simulated biodegradation. Here,
larger, more recalcitrant CDOM compounds are partially photodegraded into smaller, more labile compounds. This facilitates
further biodegradation, as the latter are more easily consumed by microorganisms (Cory et al., 2014; Moody and Worrall, 75
2016; Mostofa et al., 2013a).
The quantification and characterization of OM degradation is often limited through commonly used analytical technologies,
which focus on CDOM. The UV-visible absorbance of river water is often used as a proxy for DOC concentrations on global
scales, based on the relationship between CDOM and DOC (Adams et al., 2018; Carter et al., 2012; Griffin et al., 2011; Helms
et al., 2008; Weishaar et al., 2003). This method ignores variations in the concentration of optically invisible DOM (iDOM), 80
however. Fourier-transform ion cyclotron resonance mass spectroscopy (FT-ICR-MS), meanwhile, can reveal compositional
variations in DOM, but is difficult to align with quantitative DOC changes, and excludes biopolymers, a key DOM component.
Novel Size Exclusion Chromatography (SEC) techniques coupled to suitable detectors present a unique alternative to
quantitatively study compositional DOM transformations (Huber et al., 2011), as they can determine the concentrations of
different DOM components. Recently, SEC has been used to reveal that iDOM comprises up to 89% of the DOM fraction in 85
tropical headwaters and varies greatly over short timescales (Pereira et al., 2014a), for example; rapid changes in DOC
composition and flux have also been reported for a large Arctic river (Voss et al., 2015).
Given the importance and apparent compositional variability of DOM in tropical headwaters, it is important to understand the
processes that influence DOM concentration and composition. These processes could have significant implications both
upstream and downstream, for factors including carbon fluxes, bioreactivity, further CO2 evasion, and oxygen availability 90
(Soares et al., 2019). Though several studies have examined the effect of photodegradation on riverine DOC concentrations
(Amon and Benner, 1996; Moody and Worrall, 2016; Mostofa et al., 2005a; Pickard et al., 2017; Pullin et al., 2004; Spencer
et al., 2009; Ward et al., 2013), few have examined compositional changes of DOM (Voss et al., 2015), and to our knowledge
none have quantified the effects of photodegradation on both compositional DOM changes and DOC concentrations in tropical
headwaters. 95
One approach to ascertain the importance of photodegradation and biodegradation of DOM in tropical headwaters is to isolate
and incubate water, and establish how the DOC concentration and DOM composition changes over time, and by what dominant
process (photochemistry versus microbial activity). Here this approach was adopted on water samples collected from three
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sites in the headwaters of the Essequibo River, within the Iwokrama Rainforest, Guyana (Fig. 1). Splitting the initial sample
and only exposing half to sunlight allows us to isolate and quantify photochemical and microbial influences (Fig. 2a). 100
Furthermore, by analysing the resulting samples using SEC, for the first time we can quantify not only changes in DOC
concentrations, but also degradation-influenced compositional changes in DOM.
Figure 1. (a) Location of sample sites within the Iwokrama Rainforest, Guyana (modified from Pereira et al., 2014; inset created
using (b) Blackwater Creek, Site 1 (BC1; headwater stream, shaded conditions); (c) Burro Burro River 105
Site (BBR; open conditions); and (d) Blackwater Creek, Site 2 (BC2; headwater stream, open conditions).
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2 Methods
2.1 Sampling strategy & procedure
In-situ irradiation experiments were conducted at the Iwokrama tropical rainforest in central Guyana, South America (04° 40’ 110
N, 58° 45’ W), based from the Iwokrama International Centre for Rainforest Conservation and Development (IIC). Three
different study sites were identified, one located on the Burro Burro River (BBR), a tributary of the Essequibo connecting the
interior with the tropical Atlantic coastline, and two on the Blackwater Creek (BC1 & BC2), which is a headwater that feeds
into the BBR (see Fig. 1). The DOM compositions of both sites have previously been explored (Pereira et al., 2014a). The
experiments were performed in situ, i.e. within the water column at each location, rather than in a laboratory under controlled 115
light and temperature conditions, for two reasons. First, this approach allows to study real-world conditions – including a full
light spectrum, dappled light, changes in sunlight angle, and temperature fluctuations. Second, this approach is particularly
useful in remote regions such as the Essequibo River in central Guyana, where controlled irradiation experiments using ex situ
approaches are logistically impractical. By their nature, headwaters in the Amazon system will be in remote areas with little
infrastructure, and so the robust, portable, and simplified nature of our approach allows potential for comparison experiments 120
to be run elsewhere.
At each site, experiments were conducted to determine how photochemical and microbial / ‘light’ and ‘dark’ processes affect
the DOC concentration and composition of river water over time. In each case, homogenized river water was collected in pre-
cleaned 20 L carboys rinsed with copious amounts of sample water at 7 am (sunrise), and 180 ml was subsampled in two 60
ml HDPE bottles to provide an initial (t0) sample. The remaining water was then used to rinse and fill two opaque polyethylene 125
containers (6 L in each container), which were modified with floatation devices that permitted free floatation within the water
column at each site, while anchored to a set location. One container was covered with an opaque lid (‘dark’) to act as a dark
control, whereas the other was left exposed to sunlight (‘ambient’); the ‘dark’ container had ventilation holes to allow for free
atmospheric exchange (see Fig. 2a).
Further 180 ml subsamples were taken from each container after 12 hours, by which time the sun had set. Following this, a lid 130
was placed onto the ‘ambient’ container and both containers were left overnight. A further subsample was then taken from
each container the next morning (after 22 hours in total) to address the debate over whether photo-simulated biodegradation is
an important influence in headwaters (Cory et al., 2014; Moody and Worrall, 2016). We consider that photo-stimulated
biodegradation should occur overnight in the ‘ambient’ container, but not the ‘dark’ container, as partial photodegradation
should only have occurred in the former. This process would be evidenced by further decreases in DOC concentration in the 135
‘ambient’ container overnight, relative to the ‘dark’ container. Running the incubation overnight also allows us to calculate
changes in DOC concentration and DOM composition over an approximated day-night cycle.
This incubation procedure was conducted three times at each site (denoted a, b, and c, e.g. BC1a, BC1b, and BC1c) with
incubations being run on successive days (See Table S1 for starting dates of each incubation). The volume of water remaining
in each container was measured after the completion of each experiment to allow subsequent DOC concentrations to be 140
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corrected for evaporation. In the event of rainfall, the lid was temporarily placed upon the ‘ambient’ container until it had
stopped, and the timing of placing and removing the lid was recorded (time lost to rainfall ranged from 0-4 hrs). To allow
comparison, the changes seen in each ambient experiment were corrected for this lost time.
Each subsample was measured for temperature, pH, conductivity, total dissolved solids (TDS), salinity, and conductivity, using
a Hach® Pocket Pro+ Multi 2 Tester, and for oxidation reduction potential (ORP) using a Hach® Pocket Pro+ ORP Tester, 145
which were calibrated daily. Each subsample was then filtered through a pre-combusted (450 °C for 8 hours) GF/F (0.7 µm)
filter. The filtered water was decanted into two 60 ml HDPE bottles, which were then kept in dark conditions to prevent any
further photodegradation. One bottle was kept cool at 4 °C for analysis at a mobile laboratory setup at the IIC and another
bottle was frozen in preparation for transport back to the Lyell Centre, UK, for SEC analysis, following storage protocols by
Heinz and Zak (2018). Over the course of each experiment, UV and visible irradiance at each sample site was measured at 150
hourly intervals using SpectriLight® ILT950 and ILT950UV Spectroradiometers, coupled together to produce a single
spectrum from 200 to 700 nm.
2.2 Analysis
The DOC concentration of each sample was measured at the IIC laboratory using a Sievers M5310C Portable TOC Analyser, 155
with attached GE Autosampler; this instrument has a detection range of 0.03 ppb to 50 mg L-1. The instrument was calibrated
prior to field deployment following the manufacturers specification, and pre-weighed sucrose check standards (Sigma Aldrich,
Supelco #47289) were freshly dissolved with 18.2 M Ohm deionised carbon-free water on site at the IIC and analysed
periodically throughout the field campaign to ensure instrument accuracy and precision; these standards demonstrated
reproducibility better than 3% RSD. Measurements on each water sample were performed in triplicate (from which SD was 160
obtained), and samples were blank corrected with 18.2 M Ohm deionised carbon-free water (one blank was analysed for every
sample run, with ~20 samples per run). Concentrations from the ‘ambient’ container were then corrected for evaporation. UV-
Vis Absorbance was measured using a Biochrom® WPA Lightwave II UV/Visible Spectrophotometer. Measurements were
performed in triplicate, within 24 hrs of sampling in each case. DOM composition was determined at the UK Lyell Centre
laboratories on thawed -20 °C samples using a novel LC-OCD-OND (Liquid chromatography with organic carbon detection 165
and organic nitrogen detection) that allows ~1 ml of whole water to be injected onto a size exclusion column (SEC; 2 ml min-
1; HW50S, Tosoh, Japan) with a phosphate buffer (potassium dihydrogen phosphate 1.2 g L-1 plus 2 g L-1 di-sodium hydrogen
phosphate x 2 H2O, pH 6.58) and separated into five “compound-group specific” DOM fractions. The resulting compound
groups are identified using unique detectors for OC, UV-amenable carbon and ON (Huber, et al. 2011). All peaks were
identified and quantified with bespoke software (Labview, 2013) normalized to International Humic Substances Society humic 170
and fulvic acid standards, potassium hydrogenphthalate and potassium nitrate. The LC-OCD-OND is also able to quantify
organically bound N in the humic substances (HS) and biopolymer compound groups, in addition to inorganically bound NH4+
and NO2-, by passing the sample through a UV-reactor that converts them to NO3- (Huber et al., 2011).
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2.3 Calculations 175
To calculate the photochemical influence on net DOC concentration, the values from the ‘ambient’ container, which
theoretically represent the combined effects of both photochemical and microbial influences, were corrected by subtracting
the values from the ‘dark’ container (microbial influences only; see Fig. 1e).
The LC-OCD-OND data were used to calculate the relative proportions of the five DOC component groups outlined above:
biopolymers, HS, building blocks, LMW neutrals, and LMW acids. The DOC concentrations were then used to calculate 180
absolute concentrations for each component group in the ‘ambient’ and ‘dark’ samples, which were then subtracted as
described above to obtain the apparent photochemical effect.
The LC-OCD-OND data can also be used to further interrogate the properties of the HS compound group. Previous studies
have identified a correlation between the aromaticity (specific absorption coefficient (SAC)/OC, as obtained by the UV-
detector of the LC-OCD-OND) and molecularity (molecular weight) of the HS fraction (Huber et al., 2011; Huber and 185
Frimmel, 1996). Plotted against each other in a HS-diagram, as introduced by Huber and Frimmel (1996), this relationship can
help distinguish between aquagenic and pedogenic origins of HS. Hence, HS aromaticity and HS molecularity were calculated
from the LC-OCD-OND data following the methods of Huber et al. (2011).
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3 Results
3.1 Photochemical and microbial degradation of DOC
Figure 2. (a) Experimental setup. (b) Changes in DOC concentration across the 12 hr study period during the incubation experiments
at Sites BC1, BBR, and BC2 (suffixes a, b, and c denote repeat number of each experiment at each site; note photochemical data for 195
BC1b were lost due to measurement error). Error bars show standard deviation. Change as a percentage of initial DOC
concentration is also shown. (c) Changes in absorbance at 254 nm across the 12 hr study period. Note inverted y axes.
Photochemical mineralization of DOC appears to have occurred in each of the incubations across all three sites, with the
exception of BC2-b (Fig. 2b, blue bars), which exhibited no change exceeding sample SD. Excluding this incubation, 200
measureable (i.e., > 0.3 ppb) and significant (i.e., exceeding both sample SD and 3% RSD from repeated analysis of standard)
decreases of 0.7-1.6 ppm (~4-9% of initial DOC concentrations) were observed over the 12-hour interval in each of the other
incubations. The upper limit of this degradation exceeds those observed in previous studies on river water from the main stems
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of the Rio Negro, which reported photodegradation of ~5% after ten hours of exposure to sunlight (Amon and Benner, 1996),
and the Congo River, which showed ~10% photodegradation over a two-day period (Spencer et al., 2009). 205
The observed photodegradation of DOC was accompanied by decreases in absorbance at 254 nm (a proxy for CDOM) in each
incubation (Fig. 2c, blue bars), with the exception of BC2-b (which did not show any change in DOC within error). These
decreases in DOC and CDOM were positively correlated (R2=0.78, p<.05). Interestingly, however, the specific UV absorbance
at 254 nm (SUVA254) of all but one incubations increased due to photochemical influences (~2-5%, excluding BC1-a; Table
1). It appears then that though photobleaching occurred, non-aromatic compounds were preferentially removed by sunlight 210
over aromatic DOC.
The microbial effect on DOC concentration during the incubations was more varied and less clear than that of photochemistry
(Fig. 2b, orange bars), with biodegradation appearing to occurr in four incubations, and DOC production being apparent in
five of the incubations. All changes were measurable and exceeded sample SD, but only the decreases observed in incubations
BC1-b and BBR-c were large enough to exceed the precision as determined by repeated analysis of standards (3% RSD). Any 215
additional DOC likely either came from (potentially microbial or abiotic) breakdown of POC, or conversion of DIC to DOC
(e.g. primary production). Absorbance at 254 nm (Fig. 2c, orange bars) increased in every incubation. Taken in isolation, this
might suggest that any DOC being produced through microbial action was CDOM, and/or that iDOM was selectively utilized
by biota over CDOM (Amado et al., 2007; Moran et al., 2000). The accompanying increases in SUVA254 during each
incubation (~2-10%; Table 1) likewise suggest the preferential breakdown of non-aromatic substances and/or the production 220
of aromatic compounds; however, as shown below, this situation is more complex.
Table 1. pH and SUVA254 of initial, ‘Ambient’ (A), and ‘Dark’ (D) samples, cumulative irradiation, initial DOC concentration, initial
turbidity, and initial absorbance at 254 nm. Error shown is SD.
pH t0 Turbidity
(MJ m-2)
t0 [DOC]
t0 absorbance
at 254 nm
(L mg-1 m-1)
t0 A D t0 A D
BC1-a 4.88 4.84 4.86 0.83 ± 0.04 0.36 22.4 ± 0.2 1.14 5.10 5.22 5.23
BC1-b 4.48 4.59 4.53 0.84 ± 0.03 0.30 24.6 ± 0.2 1.17 4.75 - 5.05
BC1-c 4.45 4.43 4.52 0.48 ± 0.01 0.25 24.9 ± 0.2 1.18 4.74 5.27 5.15
BBR-a 5.65 5.88 6.02 2.35 ± 0.02 10.99 17.3 ± 0.2 0.83 4.79 5.17 4.93
BBR-b 5.54 5.73 5.62 2.79 ± 0.04 41.97 17.5 ± 0.2 0.87 4.98 5.35 5.19
BBR-c 5.51 - - 2.77 ± 0.01 51.08 17.5 ± 0.2 0.83 4.73 5.34 5.19
BC2-a 4.56 4.44 4.34 0.55 ± 0.02 20.70 17.0 ± 0.1 0.86 5.07 5.48 5.39
BC2-b 4.62 4.24 4.24 3.68 ± 0.09 8.93 18.0 ± 0.1 0.92 5.11 5.48 5.23
BC2-c 5.52 4.37 4.37 0.63 ± 0.01 7.22 19.0 ± 0.1 0.96 5.07 5.40 5.25
Photochemically-induced DOC decreases were broadly similar within and between each site (Fig. 2b), demonstrating that the 225
rate of photodegradation appears to be comparable between smaller upper headwaters (BC) and larger downstream tributaries
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(BBR), despite their hydrological differences (BC is a second-order headwater with a length of ~6 km and a catchment of ~20
km2; BBR is a fifth-order tributary with a length of 90 km and a catchment of ~3,200 km2) (Pereira et al., 2014b). The total
cumulative irradiation at the shaded site BC1 (0.25-0.35 MJ m-2) was approximately two orders of magnitude lower than those
of the more exposed sites BBR and BC2 (~10-50 MJ m-2) (Fig, 1; Table 1). For comparison, six hours of sunlight recorded in 230
a previous study in the Amazon (Amado et al., 2006) equated to ~1 M Jm-2. The similar degradation observed between sites,
despite disparity in the cumulative irradiation, suggests that cumulative irradiation alone did not control the degree of
photodegradation at the three sites. Turbidity and pH may also moderate photodegradation. Turbidity can limit the attenuation
of sunlight into the water column (Helbling and Zagarese, 2007), and thereby photodegradation. Though the turbidity values
of incubations BBR-a & c were higher than that of BC1-a & c, those of BC2-a & c were lower than those of BC1, implying 235
that low turbidity is unlikely to have caused the relatively high yield of photodegradation at BC1 (Table 1). Low pH can
enhance the breakdown of DOC as acidic conditions are more favourable for Fenton’s reactions, a pathway for
photodegradation (Molot et al., 2005; Wu et al., 2005). Conversely, DOC may also be driving the low pH at each site to some
extent – OM has been shown to buffer pH in acidic conditions in a boreal headwater (Köhler et al., 2002). However, incubations
from BC1 and BC2 had similar pH values, but different photodegradation when normalized for irradiation (Table 1). It appears 240
therefore that the amount of irradiation received or initial physical conditions of river water cannot satisfactorily explain the
degree of photodegradation measured in each incubation.
The concentration and nature of DOM has also been suggested to affect the effectiveness of photochemical and microbial
degradation; elevated DOC concentrations have been shown to increase photodegradation (Jones et al., 2016). This is
consistent with our findings; the initial DOC concentrations at Site BC1 were ~3-5 ppm higher than at Sites BBR and BC2 245
(Table 1). From a compositional viewpoint, high CDOM concentrations in aquatic environments may reduce the rate of
photodegradation, as chromophoric structures can have a photo-shielding effect, thereby limiting the attenuation of light
(Artifon et al., 2019; Mostofa et al., 2013a). Increased CDOM concentrations have also been found to be positively correlated
with higher DOC photodegradation, however, as chromophoric structures can be broken down by solar radiation (Jones et al.,
2016). This is also the case in our studies – the initial absorbance at 254 nm for incubations at site BC1 was higher than those 250
conducted at sites BBR and BC2. However, the photochemical breakdown of CDOM may not always lead to the complete
mineralization of DOC. To elucidate the influences of photochemical and microbial degradation, we next explore
compositional changes in DOM.
3.2 Photochemical and microbial driven changes in DOM composition 255
SEC, coupled with suitable detectors, permits the quantification of changes in the composition of DOM. Specifically, five
compound groups can be identified based on their molecular weight and optical properties: (i) biopolymers (likely
hydrophobic, high molecular weight >> 20.000 g mol-1, largely non-UV absorbing extracellular polymers); (ii) HS (higher
molecular weight ~ 1000 g mol-1, UV absorbing); (iii) building blocks (lower molecular weight 300-500 g mol-1, UV absorbing
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humics); (iv) LMW ‘neutrals’ (350 g mol-1, hydro- or amphi-philic, non-UV absorbing); and (v) LMW acids (350 g mol-1). 260
Determining the composition of DOM is important as it can be indicative of both its bioavailability (bio-lability), and its
susceptibility to photodegradation (photo-lability). HMW, chromophoric compounds with relatively high aromaticity are
thought to be biologically recalcitrant but photo-labile (Artifon et al., 2019; Mostofa et al., 2013a; Sharpless et al., 2014),
whereas LMW, optically-invisible molecules with lower aromaticity are bio-labile but may be less susceptible to
photodegradation (Koehler et al., 2012; Stubbins et al., 2010). Some of these bio-labile molecules may themselves be photo-265
products, created by incomplete photodegradation of larger photo-labile DOM (Stubbins et al., 2010).
Figure 3. a) Photochemistry-induced, and b) microbial-induced changes in the concentrations of photo-labile (HS, protein-based
biopolymers and building blocks) and bio-labile compounds (non-protein-based biopolymers, LMW neutrals, and LMW acids) over
12hr incubation period; c) initial proportions of photo-labile and bio-labile compounds. D) Photochemistry-induced, and e) 270
microbial-induced changes in the concentration of each DOM component over 12 hrs of incubation; f) DOM composition of initial
samples for each incubation. Note inverted y axes.
Figure 3a shows the changes in photo- and bio-labile DOM compound groups over 12 hours, as a result of photodegradation.
Photo-labile compound groups (which comprised 80-90% of the initial samples; Fig. 3e) were removed in every incubation,
whereas bio-labile compound groups were removed in some incubations but added in others. This suggests that LMW, 275
optically invisible DOM can be photodegraded (e.g. BC1-a). Furthermore, photodegradation appears to be capable of both
complete mineralization of photo-labile compound groups, and their partial degradation into photo-products (e.g. BC1-c). The
variation seen in bio-labile compound groups (Fig. 3a) may therefore represent a balance between their mineralization and
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addition (through partial photodegradation of photo-labile components). As a percentage of initial concentrations, the
variations in the bio-labile pool were of similar or greater magnitude than in the photo-labile pool (Fig. S2b), implying that the 280
latter is not necessarily more reactive to photodegradation. Therefore, here we show that photodegradation does indeed break
down photo-labile DOM. However, complete mineralization, combined with sunlight’s ability to also remove supposedly
photo-resistant DOM (Stubbins et al., 2010), means that photodegradation in tropical headwaters does not necessarily improve
the bioavailability of DOM transported downstream. It does, however, decrease the total amount of DOM being exported.
These compositional changes can be explored in greater detail – Fig. 3b shows that in all experiments bar BC1-a, the 285
breakdown of photo-labile DOM was largely driven by HS (which was the most abundant compound group in initial samples;
Fig. 3f). Likely HS breakdown products would be building blocks and LMW neutrals. The former; however, is comprised of
UV-absorbing compounds, and so would also be expected to break down upon exposure to UV/visible irradiation. This may
explain why its concentration increased in incubations where large quantities of HS were degraded (e.g., BC1-c, BBR-a),
whereas BC1-a, where there was a (relatively small) increase in HS, showed the largest decrease in the concentration of 290
building blocks. It is interesting to note that the initial proportion of HS in incubation BC-1a (67%) was lower than that of all
other runs (76-79%; Fig. 3f). The initial composition of HS for this incubation may have been relatively resistant to
photodegradation. The photochemical and/or microbial breakdown of POC can recharge DOC, which could also have
recharged the HS compound group (Mayer et al., 2006; Mostofa et al., 2013b; Porcal et al., 2015).
Figure 3c shows that the bio-labile and photo-labile (theoretically bio-recalcitrant) DOM compound groups showed varied 295
responses to microbial action, mirroring the varied response in DOC concentration (Fig. 2b). Despite its relative availability
for microbial respiration, bio-labile DOM was only removed from two incubations (BBR-b, c, of which only BBR-c showed
a change in DOC exceeding instrument RSD). The availability of LMW, bio-labile compound groups in initial samples (Fig.
3f) did not appear to influence the degree of microbial degradation. Further, HMW, supposedly recalcitrant DOM was
apparently susceptible to microbial degradation – indeed, for the two incubations that demonstrated the highest net microbial 300
degradation (~4-5%; BC1-c and BBR-c; Fig. 2), more photo-labile DOM was removed than bio-labile DOM (Fig. 3c).
Therefore, the observed increases in absorbance at 254 nm do not necessarily reflect the preferential removal of non-UV
absorbing DOM. While over longer timescales, previous studies have shown that microbes can process HMW, photo-labile
DOM where bio-labile DOM is limited; this also appears to be the case here (Koehler et al., 2012).
Figure 3d shows that, within the photo-labile pool, HS was removed by microbial activity in five of the incubations, including 305
BC1-c and BBR-c, at a similar magnitude to that of photodegradation. It appears therefore, that even HMW molecules within
this compound group are bio-available, not just smaller molecules (building blocks). LMW acids were only observed to any
significant extent in Experiment BC1-c (Fig. 3d); they were formed through microbial processes, as none were present in the
initial sample (Fig. 3f).
These findings have significant implications for understanding rivers as biogeochemical reactors- on short timescales. 310
Microbial activity does not appear to be dictated by the lability of the DOM pool in tropical headwaters, but instead is capable
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of altering a variety of DOM compositional pools on daily timescales, through both biodegradation and bio-
accumulation/primary production.
Though biopolymers represented a relatively minor proportion of initial samples (~1-6%; Fig. 3f), relative to initial
concentrations, this compound group was highly reactive to both photochemical and microbial influences (Fig. S2a, c). This 315
compound group contains both UV and non-UV containing compounds, which may explain this relatively high reactivity
under both stimuli. Though detectable through SEC, biopolymers are not measured by FT-ICR-MS, further highlighting the
importance of the former. The biopolymer compound group likely contains proteins and amino sugars, which are important
for biota, and polysaccharides, which are a key component in microbial biofilms (Huber et al., 2011; Her et al., 2002; Flemming
et al., 2007). The apparent high reactivity of this compound group may therefore reflect complex microbial processes, and 320
warrants further investigation.
The above conclusions, combined with the revelation that photodegradation does not always yield bio-degradable products,
would suggest that photo-simulated biodegradation (Cory et al., 2014) is unlikely to be an important process in tropical
headwaters on short (daily) timescales in upper headwaters. This was shown to be the case here; there was no consistent change
in the DOC concentrations of the ‘ambient’ incubations overnight, following their exposure to sunlight: five of the incubations 325
showed no change overnight exceeding SD, two incubations (BC1-c and BBR-c) experienced a slight increase in DOC (0.5-
0.6 ppm), and only one incubation (BBR-a) showed further decreases in DOC (-0.5 ppm) that would be expected had photo-
stimulated bacterial degradation occurred overnight (Fig. S1). A previous study also found no evidence of photo-stimulated
bacterial degradation over a similar timescale in a temperate peatland headwater (Moody and Worrall, 2016).
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3.2.1 Revisiting the stability of HS in headwater rivers
Figure 4. HS diagrams showing aromaticity and molecularity of the HS fraction for the initial samples from each incubation and
samples after 12 hrs of exposure to photochemical and microbial influence (b-d). (a) Also plotted are data for standards and
representative lake and river samples from Huber et al. (2011). Dashed red box in (a) shows region of plots (b-d). 335
Figure 3f shows that HS was the most common compound group in riverine DOM at each site. Previous studies have explored
compositional changes and inherent stability of HS by examining the relationship between aromaticity – i.e. the specific
absorption coefficient (SAC) normalized to OC and molecularity (molecular weight) (Huber et al., 2011). This approach
shows that all samples exhibited high aromaticity and molecularity, indicative of a pedogenic origin (Fig. 4). Examining Fig. 340
4 for every incubation, the molecularity of the HS compound group decreased following photodegradation, compared to the
initial sample. This confirms that sunlight breaks down larger organic molecules into smaller components, both within the HS
compound group and between DOM compound groups (Mostofa et al., 2013a). As anticipated, we observe no such clear
change in molecularity in response to microbial action (Fig. 4, red points), in line with the more varied responses of DOM
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composition and DOC concentration to microbial action (Figs. 2b, 3d). While there was no consistent trend, changes in 345
molecular weight during microbial incubation were similar in magnitude to those observed under photodegradation, suggesting
microbes are capable of both the breakdown and self-assembly of DOC within the HS component group on daily timescales
(Xu and Guo, 2018).
Unlike for molecularity, there was no consistent decrease in aromaticity accompanying photodegradation (Fig. 4), nor was
there a consistent change in aromaticity in response to microbial activity across the incubations. The changes in UV absorbance 350
at 254 nm seen in Fig. 2c may therefore have been driven by aromaticity changes outside of the HS compound group.
Photochemically-driven changes in aromaticity did not correspond with observed changes in DOC (Fig. 2), suggesting that the
aromaticity of the HS compound group is relatively unsusceptible to photodegradation. Further examination of the interaction
between aromatic compounds and sunlight within the HS compound group is needed, as previous studies have reported a loss
of DOM aromaticity in response to sunlight (Sharpless et al., 2014). 355
The initial sample of BC1-a, which was the only incubation not to experience photodegradation of HS (Fig. 3b), had the highest
aromaticity and lowest molecularity of any of the samples analysed (Fig. 4b). We can only speculate, but it seems possible
that the lower molecularity, combined with the lower overall proportion of HS in the initial sample (67%, vs. 76-79% for other
samples; Fig. 3f), implies that the HS compound group for BC1a was more resistant to degradation than for the other samples;
possibly it had undergone more photodegradation prior to incubation. If the more labile compounds of the HS compound group 360
in this incubation had already broken down into building blocks, this could explain why photodegradation in BC1-a was largely
achieved through degradation of building blocks, not HS (Fig. 3b).
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3.3 Organic Nitrogen Dynamics of DOM
Figure 5. (a) Initial C/N Ratios for HS and biopolymers (inset shows relationships between components before and after each
incubation); Photochemistry- and microbial-driven changes in C/N ratios in (b) the biopolymer fraction and (c) the HS fraction over
12 hr incubation. Photochemistry- and microbial-driven changes in (d) nitrate and (e) ammonium during incubation.
Addressing changes in DON may help to explain the response of DOM to microbial and photochemical influences. In addition 370
to influencing net DOC concentration and DOM composition, photodegradation can also break down DON, leading to the
formation of Dissolved Inorganic Nitrogen (DIN) in the form of nitrate and ammonium, which are critical for biological
productivity (Zepp, 2003). Furthermore, the availability of DIN has been shown to limit the biological metabolism of DOM,
whereas DON has been shown to be utilized by bacteria at a higher rate than DOC, affecting the rate at which they are
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biologically processed (Wiegner and Seitzinger, 2001) and thereby altering the C/N ratio. Therefore, we tested if 375
photochemistry or microbial activity altered C/N ratios in the biopolymer and HS component groups, and whether this resulted
in an increase in the concentrations of nitrate and ammonium. We also explore whether initial concentrations of DON and DIN
can inform microbial activity.
The initial availability of DIN could not explain the variation in the microbial-driven changes in DOC (Fig. 2a, Table S4), nor
could initial DON concentrations in biopolymers and HS explain the observed changes in concentration for these component 380
groups. In the initial samples from each incubation, the C/N ratios of the HS and biopolymer component groups were highly
correlated (R2 =0.97; Fig. 5a); however, this correlation disappeared following incubation. The influence of terrestrial inputs
in tropical headwaters may maintain the partitioning of DON in HS and biopolymers, but once disconnected from this input
the C/N ratios may diverge via a complex combination of factors that cannot be disentangled via the current approach. For
both biopolymers and HS (Figs. 5b, c), there was no clear trend in C/N ratios during each incubation, for either photochemical 385
or microbial influences. This suggests that DON in these fractions responds differently than non-N containing compounds, but
not in a consistent way across each incubation.
Changes in the C/N ratios from neither fraction correlated with observed changes in ammonium or nitrate (Figs. 5d, e). We
therefore conclude that the concentrations of these components were driven by other processes, such as breakdown of
particulate sources, aquatic biomass, or DON from other DOM components (e.g., building blocks, LMW neutrals, or acids) 390
(Kellerman et al., 2015). Furthermore, we determined that aliphatic DOM, rather than aromatic DOM, was more commonly
associated with N. The lack of a clear trend in microbial-driven changes in nitrate (Fig. 5d, orange bars) highlights the complex
nature of N dynamics presented in this study. Nitrate and ammonium decreased in concentration across many of the incubation
experiments; we envision that (photochemistry-induced) microbial scavenging may have outstripped their production via the
photochemical breakdown of DON (Zepp, 2003); this concept could be tested in future by filtering water prior to incubation 395
to remove bacteria. Though evidently this study was able to capture significant degradation of DOC on daily timescales,
photochemically and/or microbial-driven N cycling between organic and inorganic pools may be occurring at sub-daily
timescales, i.e., too rapid to be characterized by our approach (Kellerman et al., 2015).
4 Discussion 400
In our approach above we investigated effects of photochemical and microbial influences on riverine DOC separately;
however, our analysis also allows for the investigation of their combined effects in ‘real world’ conditions over a day-night
cycle. There was a similar extent of combined degradation in Blackwater Creek (~1-7%), a headwater stream, and the Burro-
Burro River, a larger tributary (~1-7%; Fig. 6a); our results as presented above show that photodegradation was largely
responsible for these decreases in DOC (Fig. 2). This similarity suggests that we may be able to model a relatively constant 405
daily degradation rate of DOC for these scales of river. This approach would be aided by the fact that headwaters comprise
70-80% of total river networks (Gomi et al., 2002). That these rates are similar could suggest either that the rate of
transformation was relatively consistent from one location to the other, or that new material was constantly entering the river
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system away from the headwaters. The latter is unlikely; however, as from a mass balance perspective the addition of new OM
would be less influential in larger tributaries than in headwaters. Our results suggest that photodegradation is not necessarily 410
limited by the availability of aromatic, HMW, and humic DOM, such that the reactivity of the remaining DOM pool may
persist following successive days of exposure. This may explain why the degradation rates obtained for BBR were similar to
those seen in BC.
If combined with a reliable field-based method for measuring CO2, the approach presented here could be used to model DOC
degradation’s contribution to the production of CO2. This would help to quantify the contribution of DOC decomposition to 415
the total CO2 evasion from rivers (Davidson et al., 2010); which is potentially substantial (OM decomposition is estimated to
contribute ~75% of total evasion from the Amazon) but currently poorly constrained (Raymond et al., 2013; Richey et al.,
2002). The finding presented here that photodegradation is a more influential process on the complete mineralization of DOC
than biodegradation suggests that the former is therefore a more important driver of the degassing of CO2 in tropical
headwaters. The upper limit of DOC photodegradation obtained here (up to 9% over 12 hours) exceeded those observed for 420
water samples from the main stems of larger tropical rivers (Amon and Benner, 1996; Spencer et al., 2009), albeit that these
values were obtained with different approaches. This suggests that photodegradation rates, and therefore potentially rates of
CO2 degassing, may decease along the river continuum from source to sea. This could be investigated further by conducting
incubations along the length of a river network – for example analysing the main stem and mouth of the Essequibo in addition
to BC and BBR – to explore how DOM composition and DOC concentrations change along the river continuum (Cole et al., 425
In this study we opted to study daily rates of photo- and biodegradation. This allowed us to study the maximum period of
continuous sunlight that river water will experience in ‘real world’ conditions, and was also in line with estimates of in-stream
residence times (though more data are needed regarding tropical headwaters in this regard) (Worrall et al., 2014). However,
our approach could also easily be adapted to include more frequent sampling steps to study hourly degradation rates, and also 430
to further investigate N cycling. From a compositional point of view, this could also shed further light on the complex microbial
processes affecting the biopolymer compound group, which was shown here to be particularly reactive; biopolymers contain
components important for biota and in particular microbial biofilms (Huber et al., 2011; Her et al., 2002; Flemming et al.,
2007). Conversely, longer, multi-day ‘dark’ incubations may deliver larger, mores discernible and significant changes in DOC
concentrations as a result of microbial processes. This could lead to clearer results regarding the net effects of microbial 435
activity, and could also capture the effects of slower, anaerobic biological processes (e.g., methanogenesis; Stanley et al.,
2016). This approach also better reflect scenarios in which flooding leads to standing water and corresponding slower residence
times, in addition to downstream sections of the river network, where the flow can become lentic (Junk, 1989). The approach
presented here is limited in that it does not quantify POM-DOM interchange during the incubation process – filtering prior to
incubation could circumvent this issue, but would reduce representability, whereas significantly larger samples would be 440
needed to quantify POC.
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Figure 6. Overall changes in (a) DOC concentration, and (b) photo-labile and bio-labile compounds across a day-night cycle (i.e. 445
change seen in ‘ambient’ container over 22 hr incubation); error bars show SD. Note inverted y axes.
The degree of DOC degradation observed in BC is higher than that observed in headwaters in other climate zones like a
boreal headwater (0.2 km2 catchment) in Sweden (~10% decrease in DOC over 48 hrs (Köhler et al., 2002)) – but is lower
than that observed in a temperate peatland headwater (0.2 km2) in the UK (~40% decrease over 24 hrs(Moody and Worrall,
2016)) and two temperate forest streams in Japan (~21-36% over 12-13 hrs (Mostofa et al., 2007, 2005b)). Application of the 450
technique demonstrated here to different climate zones and different river stages could help to quantify how variations in DOC
concentrations, environmental conditions, and DOM composition help to modulate degradation.
This study can also address the relative importance of combined photochemical and microbial influences on DOM composition
and DOC concentration, compared to other potential influences. Hydrological mobilization at Blackwater Creek has been
shown to drive to variations in DOC one-to-two orders of magnitude larger than those measured in this study, over a shorter 455
timescale (Pereira et al., 2014a); this was also accompanied by greater compositional variability than observed here (Fig. 7b).
Further downstream, however, this terrestrial-driven variability may fade, leading to combined photochemical and microbial
influences becoming more important. Additionally, their relative importance in tropical headwaters will also likely increase
outside of the rainy season, when mobilization of DOM from rainstorms will become less frequent.
5 Conclusions
In summary, we reveal that photodegradation is an important process in tropical headwaters – a key component of the global
carbon cycle – regarding both the complete mineralization of DOC (and therefore potentially the generation of greenhouse
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gases) and the alteration of DOM composition. Daily, the degradation of DOC in tropical headwaters and tributaries occurs at
a similar scale (~1-7%) as in larger tropical rivers (Amon and Benner, 1996; Spencer et al., 2009). Photodegradation has a 465
larger influence on DOC transformation than microbial activity, which had a more varied and inconclusive response. Our study
demonstrates that sunlight alone does not drive the extent of photodegradation over daily timescales, rather it appears to be the
product of a complex range of influences, possibly including the initial physical properties of the water column and the initial
concentration and composition of DOM (Jones et al., 2016; Mostofa et al., 2013a), linking with the surrounding environment
and hydro-chemistry of the river headwater. 470
Data availability
All data are available in the supplementary material (see link below).
Supplement Link
Author contributions
JFS, RP, TW, JBi, SN, WH, and ST designed the study; JFS, RP, JBi, ST, SN, JBr, and LJ conducted field sampling; JFS, RP, 475
JBi, ST, and WH conducted laboratory analysis; JFS, RP, and TW wrote the manuscript, with additional contributions from
JBi, ST, SN, JBr, and LJ.
Competing interests
The authors declare that they have no conflict of interest.
Acknowledgements 480
The authors would like to thank Dane Gobin, Raquel Thomas, and the staff of the Iwokrama International Iwokrama
International Centre for Rainforest Research and Development (IIC) for their support, advice, and expertise during the planning
and implementation of the research program. We would also like to thank Alan MacDonald at the British Geological Survey
(BGS) for advice, and for loaning equipment. We would like to dedicate this study to our late friend and colleague, Walter
Hill. 485
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Financial Support
This study was supported by funding from IIC, and RP and SNA acknowledge funding from the BGS BUFI scheme. RP
acknowledges support from the European Research Council (ERC) under the European Union’s Horizon 2020 research and
innovation programme (grant agreement No. 949495).
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... When there are high concentrations of soluble oxygen in the water, a lower concentration of COD, that is to say organic material, is anticipated. The high level of microbial activity occurring and the breaking down of organic material perpetually entering the lake are associated (Spray et al., 2021). Inlet had the maximum mean ammoniacal nitrogen concentration (0.163 ± 0.004 mg/L) in the spring, followed by Trapa abundance site (0.148 ± 0.006 mg/L) and outlet (0.125 ± 0.001 mg/L). ...
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Hokersar wetland is vital for the endurance of migratory birds, but the topography of wetland has been severely altered during the last four decades. This wetland area has reduced from 1875.04 ha in 1969 to 1300 ha in 2008 due to several reasons, viz, encroachments and pollution. Keeping these facts in view, the present study was carried out to determine the impact of anthropogenic activities on the water quality of the wetland. Standard methods were used for analyzing various physico-chemical characteristics of water of a wetland. Among all the four sites, inlet sampling site was having the highest mean value of turbidity (2.06 NTU), water temperature (23.55 °C), electrical conductivity (0.233 dS/m), nitrate-nitrogen (0.046 mg/L), nitrite nitrogen (0.681 mg/L), ammoniacal nitrogen (0.210 mg/L), calcium (48.60 mg/L), phosphate (0.071 mg/L), and potassium (0.456 mg/L). The highest mean value (11.34 mg/L) of magnesium was recorded at the outlet, also having maximum pH (7.10). The lowest mean values of all physico-chemical characteristics of water were observed at the outlet site except for pH and Mg. The highest mean value was found for water temperature (27.25 °C), EC (0.228 dS/m, nitrate�nitrogen (0.485 mg/L), nitrite-nitrogen (0.035 mg/L), ammoniacal nitrogen (0.196 mg/L), Ca (58.08 mg/L), phosphate (0.047 mg/L), and K (0.480 mg/L) during summer season. A decreasing trend of trace metal concentration at diferent site in all seasons was observed in the following order: inlet (polluted)>center>Trapa abundance site>outlet. Similarly, COD and BOD values ranged from 69.69 to 115.89 mg/L and 56.08 to 95.05 mg/L in three diferent seasons, respectively. The wetland’s BOD level (95.05 mg/L) is too high, indicating high pollution level. The results of the present study have indicated that the pollutants and trace elements fnd their entry into the aquatic ecosystem and are a matter of great concern.
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Inland waters transport, transform and retain significant amounts of dissolved organic carbon (DOC) that may be biologically reactive (bioreactive) and thus potentially degraded into atmospheric CO2. Despite its global importance, relatively little is known about environmental controls on bioreactivity of DOC as it moves through river systems with varying water residence time (WRT). Here we determined the influence of WRT and landscape properties on DOC bioreactivity in 15 Swedish catchments spanning a large geographical and environmental gradient. We found that the short-term bioreactive pools (0–6 d of decay experiments) were linked to high aquatic primary productivity that, in turn, was stimulated by phosphorus loading from forested, agricultural and urban areas. Unexpectedly, the percentage of long-term bioreactive DOC (determined in 1-year experiments) increased with WRT, possibly due to photo-transformation of recalcitrant DOC from terrestrial sources into long-term bioreactive DOC with relatively lower aromaticity. Thus, despite overall decreases in DOC during water transit through the inland water continuum, DOC becomes relatively more bioreactive on a long time-scale. This increase in DOC bioreactivity with increasing WRT along the freshwater continuum has previously been overlooked. Further studies are needed to explain the processes and mechanisms behind this pattern on a molecular level.
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CO2 emissions from inland surface waters to the atmosphere are almost as large as the net carbon transfer from the atmosphere to Earth's land surface. This large flux is supported by dissolved organic matter (DOM) from land and its complete oxidation to CO2 in freshwaters. A critical nexus in the global carbon cycle is the fate of DOM, either complete or partial oxidation. Interactions between sunlight and microbes control DOM degradation, but the relative importance of photodegradation vs. degradation by microbes is poorly known. The knowledge gaps required to advance understanding of key interactions between photochemistry and biology influencing DOM degradation include: (1) the efficiencies and products of DOM photodegradation, (2) how do photo-products control microbial metabolism of photo-altered DOM and on what time scales, and (3) how do water and DOM residence times and light exposure interact to determine the fate of DOM moving across the landscape to oceans?
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Dissolved organic matter (DOM) is an important constituent of freshwater that participates in a number of key ecological and biogeochemical processes but can be problematic during water treatment. Thus, the demand for rapid and reliable monitoring is growing, and spectroscopic methods are potentially useful. A model with 3 components—2 that absorb in the ultraviolet (UV) range and are present at variable concentrations and a third that does not absorb light and is present at a low constant concentration—was previously found to yield reliable predictions of dissolved organic carbon concentration [DOC]. The model underestimated [DOC] in shallow eutrophic lakes in the Yangtze Basin, China, however, raising the possibility that DOM derived from algae might be poorly estimated, an idea supported by new data reported here for eutrophic British lakes. We estimated the extinction coefficients in the UV range of algae-derived DOM from published data on algal cultures and from new data from outdoor mesocosm experiments in which high concentrations of DOC were generated under conditions comparable to those in eutrophic freshwaters. The results demonstrate the weak UV absorbance of DOM from algae compared to DOM from terrestrial sources. A modified model, in which the third component represents algae-derived DOM present at variable concentrations, allowed contributions of such DOM to be estimated by combining the spectroscopic data with [DOC] measured by laboratory combustion. Estimated concentrations of algae-derived DOC in 77 surface freshwater samples ranged from 0 to 8.6 mg L⁻¹, and the fraction of algae-derived DOM ranged from 0% to 100%.
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Globally, inland waters receive a significant but ill-defined quantity of terrestrial carbon (C). When summed, the contemporary estimates for the three possible fates of C in inland waters (storage, outgassing, and export) highlight that terrestrial landscapes may deliver upward of 5.1 Pg of C annually. This review of flux estimates over the last decade has revealed an average increase of ∼ 0.3 Pg C yr−1, indicating a historical underestimation of the amount of terrestrial-C exported to inland waters. The continual increase in the estimates also underscores large data gaps and uncertainty. As research continues to refine these aquatic fluxes, especially C outgassed from the humid tropics and other understudied regions, we expect the global estimate of terrestrial-C transferred to inland waters to rise. An important implication of this upward refinement is that terrestrial net ecosystem production may be overestimated with ramifications for modeling of the global C cycle.
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We discuss concentrations of dissolved CH4, N2O, O2, NO⁻3 and NH⁺4 , and emission fluxes of CH4 and N2O for river sites in the western Congo Basin, Republic of Congo (ROC). Savannah, swamp forest and tropical forest samples were collected from the Congo main stem and seven of its tributaries during November 2010 (41 samples; "wet season") and August 2011 (25 samples; "dry season"; CH4 and N2O only). Dissolved inorganic nitrogen (DIN: NH⁺4 C NO⁻3 ; wet season) was dominated by NO⁻3 (63±19% of DIN). Total DIN concentrations (1.5- 45.3 μmol L⁻¹/ were consistent with the near absence of agricultural, domestic and industrial sources for all three land types. Dissolved O2 (wet season) was mostly undersaturated in swamp forest (36±29 %) and tropical forest (77±36 %) rivers but predominantly supersaturated in savannah rivers (100±17 %). The dissolved concentrations of CH4 and N2O were within the range of values reported earlier for sub-Saharan African rivers. Dissolved CH4 was found to be supersaturated (11.2-9553 nmol L⁻¹; 440-354 444 %), whereas N2O ranged from strong undersaturation to supersaturation (3.2-20.6 nmol L⁻¹; 47-205 %). Evidently, rivers of the ROC are persistent local sources of CH4 and can be minor sources or sinks for N2O. During the dry season the mean and range of CH4 and N2O concentrations were quite similar for the three land types. Wet and dry season mean concentrations and ranges were not significant for N2O for any land type or for CH4 in savannah rivers. The latter observation is consistent with seasonal buffering of river discharge by an underlying sandstone aquifer. Significantly higher wet season CH4 concentrations in swamp and forest rivers suggest that CH4 can be derived from floating macrophytes during flooding and/or enhanced methanogenesis in adjacent flooded soils. Swamp rivers also exhibited both low (47 %) and high (205 %) N2O saturation but wet season values were overall significantly lower than in either tropical forest or savannah rivers, which were always supersaturated (103-266 %) and for which the overall means and ranges of N2O were not significantly different. In swamp and forest rivers O2 saturation co-varied inversely with CH4 saturation (log %) and positively with % N2O. A significant positive correlation between N2O and O2 saturation in swamp rivers was coincident with strong N2O and O2 undersaturation, indicating N2O consumption during denitrification in the sediments. In savannah rivers persistent N2O supersaturation and a negative correlation between N2O and O2 suggest N2O production mainly by nitrification. This is consistent with a stronger correlation between N2O and NH⁺4 than between N2O and NO⁻³ . Our ranges of values for CH4 and N2O emission fluxes (33-48 705 μmol CH4 m⁻² d⁻¹; 1-67 μmol N2Om⁻² d⁻¹/ are within the ranges previously estimated for sub-Saharan African rivers but they include uncertainties deriving from our use of "basin-wide" values for CH4 and N2O gas transfer velocities. Even so, because we did not account for any contribution from ebullition, which is quite likely for CH4 (at least 20 %), we consider our emission fluxes for CH4 to be conservative.
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Aquatic systems draining peatland catchments receive a high loading of dissolved organic carbon (DOC) from the surrounding terrestrial environment. Whilst photo-processing is known to be an important process in the transformation of aquatic DOC, the drivers of temporal variability in this pathway are less well understood. In this study, 8 h laboratory irradiation experiments were conducted on water samples collected from two contrasting peatland aquatic systems in Scotland: a peatland stream and a reservoir in a catchment with high percentage peat cover. Samples were collected monthly at both sites from May 2014 to May 2015 and from the stream system during two rainfall events. DOC concentrations, absorbance properties and fluorescence characteristics were measured to investigate characteristics of the photochemically labile fraction of DOC. CO2 and CO produced by irradiation were also measured to determine gaseous photoproduction and intrinsic sample photoreactivity. Significant variation was seen in the photoreactivity of DOC between the two systems, with total irradiation-induced changes typically 2 orders of magnitude greater at the high-DOC stream site. This is attributed to longer water residence times in the reservoir rendering a higher proportion of the DOC recalcitrant to photo-processing. During the experimental irradiation, 7 % of DOC in the stream water samples was photochemically reactive and direct conversion to CO2 accounted for 46 % of the measured DOC loss. Rainfall events were identified as important in replenishing photoreactive material in the stream, with lignin phenol data indicating mobilisation of fresh DOC derived from woody vegetation in the upper catchment. This study shows that peatland catchments produce significant volumes of aromatic DOC and that photoreactivity of this DOC is greatest in headwater streams; however, an improved understanding of water residence times and DOC input–output along the source to sea aquatic pathway is required to determine the fate of peatland carbon.
Organic matter (OM) in aquatic system is originated from autochthonous and allochthonous natural sources as well as anthropogenic inputs, and can be found in dissolved, particulate or colloidal form. According to the type/composition, OM can be divided in non-humic substances (NHS) or humic substances (HS). The present review focuses on the main groups that constitute the NHS (carbohydrates, proteins, lipids, and lignin) and their role as chemical biomarkers, as well as the main characteristics of HS are presented. HS functions, properties and mechanisms are discussed, in addition to their association to the fate, bioavailability, and toxicity of organic microcontaminants in the aquatic systems. Despite the growing diversity and potential impacts of organic microcontaminants to the aquatic environment, limited information is available about their association with OM. A protective effect is, however, normally seen since the presence of OM (HS mainly) may reduce bioavailability and, consequently, the concentration of organic microcontaminants within the organism. It may also affect the toxicity by either absorbing ultraviolet radiation incidence and, then, reducing the formation of phototoxic compounds, or by increasing the oxygen reactive species and, thus, affecting the decomposition of natural and anthropogenic organic compounds. In addition, the outcome data is hard to compare since each study follows unique experimental protocols. The often use of commercial humic acid (Aldrich) as a generic source of OM in studies can also hinder comparisons since differences in composition makes this type of OM not representative of any aquatic environment. Thus, the current challenge is find out how this clear fragmentation can be overcome.
Dissolved organic matter (DOM) is ubiquitous and plays an important role in regulating water quality, ecological function, and the fate and transport of trace elements and pollutants in aquatic environments. Both the colloidal precursors (i.e. <1 kDa) and bulk DOM collected from a freshwater estuary were incubated in the dark for 21 days to examine dynamic changes in molecular size and composition induced by microbial degradation and self-assembly. Results showed that the concentrations of total organic carbon, carbohydrates, and protein-like substances decreased by 11-30% during incubation, while those of humic- and fulvic-like substances remained relatively constant, indicating humic substances are more resistant to microbial utilization compared to carbohydrates and protein-like DOM. Despite the different extents in decline, these DOM components had a similar transformation pathway from the <1 kDa to colloids (1 kDa-0.45 μm) and further to microparticles (>0.45 μm). Overall, carbohydrates and protein-like substances, especially the high molecular weight components, were preferentially decomposed by microorganisms whereas humic- and fulvic-like DOM components significantly coagulated through abiotic self-assembly. The contrasting degradation/transformation pathways between the humic-like and protein-like substances along the size continuum, as also characterized by flow field-flow fractionation analysis, demonstrated that the dynamic transformation and degradation of DOM is regulated by both molecular size and organic composition. This finding provides new insights into the biogeochemical cycling pathways of heterogeneous DOM and its environmental fate and ecological role in aquatic systems.
This study aimed to evaluate the effects of freezing and cold storage at 4 °C on bulk dissolved organic carbon (DOC) and nitrogen (DON) concentration and SEC fractions determined with size exclusion chromatography (SEC), as well as on spectral properties of dissolved organic matter (DOM) analyzed with fluorescence spectroscopy. In order to account for differences in DOM composition and source we analyzed storage effects for three different sample types, including a lake water sample representing freshwater DOM, a leaf litter leachate of Phragmites australis representing a terrestrial, 'fresh' DOM source and peatland porewater samples. According to our findings one week of cold storage can bias DOC and DON determination. Overall, the determination of DOC and DON concentration with SEC analysis for all three sample types were little susceptible to alterations due to freezing. The findings derived for the sampling locations investigated here may not apply for other sampling locations and/or sample types. However, DOC size fractions and DON concentration of formerly frozen samples should be interpreted with caution when sample concentrations are high. Alteration of some optical properties (HIX and SUVA254) due to freezing were evident, and therefore we recommend immediate analysis of samples for spectral analysis.