1. 10484463
BSc (Hons) Applied Zoology (CORN310)
Project Report
O2 to be, or not to be? A Comparison of the Oxygenating
Differences of Invasive Non-Native Lagarosiphon major
(Ridley) Moss and Native Ceratophyllum demersum L.
Supervisor: Dr Peter McGregor
Lagarosiphon major (left) and Ceratophyllumdemersum (right) (Mitchell-Holland, 2016).
2. 1
Abstract
Lagarosiphon major is a submerged macrophyte that is recognised as a problematic,
invasive non-native species (INNS) in many countries including the UK. It is widely
sold and promoted through the aquarium and water garden industry as an “efficient
oxygenator” for freshwater systems, irrespective of evidence of its adverse ecological
and economic impacts, and an absence of evidence to support its statement. A key
concern, relating to its rapid growth rate and high biomass density is that L. major can
impose self-limitation of photosynthetic and respiratory activity, causing it to
consume more oxygen than it produces. Low dissolved oxygen (DO) conditions
typify diminished water quality and seriously limits oxygen-dependent organisms.
Established in small pond conditions in buckets over several months, the DO,
biomass, and associated pond life abundances of L. major and a comparable native
macrophyte, Ceratophyllum demersum, were assessed experimentally to determine
which species is the most efficient at maintaining a healthy freshwater environment.
Both the establishment time and the species had highly significant effects on DO
concentrations and pond-life abundance; L. major produced the least amount of
oxygen over time and had significantly less associated pond life compared to the
native plant. L. major also increased significantly in overall biomass compared to C.
demersum, indicating the higher invasive ability of the non-native species. In
conclusion, I suggest that invasive L. major is detrimental to freshwater ecosystems,
causing DO depletions and creating unfavorable living conditions for pond life, which
deteriorates over time. These detriments are likely to be exacerbated during the usual
growth season of L. major, and in the future as a result of global warming increases,
indicating that this highly invasive species should be withdrawn from sale in the UK.
It is recommended that native C. demersum be promoted through the trade as an
efficient oxygenator that improves water quality and habitat conditions over time.
Key words: invasive non-native species, macrophytes, dissolved oxygen, pond life,
biomass, water quality.
3. 2
1. Introduction
As a result of worldwide travel and trade, numerous invasive species have been
introduced, both intentionally and inadvertently, into areas beyond their natural range
(Westphal et al., 2008; Stiers et al., 2011). Aquatic plant species invasion has been
recognised as one of the largest threats to freshwater ecosystems and biodiversity,
with detrimental consequences for ecology and the economy (Pimentel et al., 2000;
Dudgeon et al., 2006; Strayer, 2010; Riis et al., 2011). In cases where such plants
have outcompeted and replaced native submerged vegetation (Rattray et al., 1994:
Keenan et al., 2009), severe depletion of dissolved oxygen (DO), quantity of primary
production, and diminished water quality has been the result (Caraco et al., 2006;
Leppi et al., 2016). Organisms, such as freshwater fish, invertebrates, plants and
bacteria, rely on DO (the level of free, non-compound oxygen present in water) for
survival, thus cannot withstand anoxic (a total depletion of oxygen) or even hypoxic
(low-oxygen) conditions for extended periods of time (Gray et al., 2002; Caraco et
al., 2006; Lenntech, 2015). Such depletions also lead to cascading impacts on nutrient
and trace gas chemistry, altering the suitability of the environment as habitat and
further threatening most aquatic life forms (Fenchel et al., 1998; Wetzel, 2001; Baird
et al., 2004; Morris et al., 2004; Dybas, 2005; Kemker, 2013). In the same way DO
has an effect on the environment and organisms, the presence and abundance of
organisms and organic matter (and their associated biological processes) can greatly
influence DO concentrations in a body of water (Caraco et al., 2006; Desmet et al.,
2011; Kemker, 2013; Ribaudo et al., 2014). For example, while photosynthesis
contributes to an increase in DO (Desmet et al., 2011), the process of respiration by
organisms, and decomposition of organic matter by microorganisms, can severely
deplete the available DO for aerobic species (Caraco et al., 2006: Desmet et al., 2011;
EPA, 2012; Annis, 2014). Despite high productivity during the day (light induced
photosynthesis) (Carrillo et al., 2006; Ribaudo et al., 2014), dense clusters of aquatic
plants can cause water hypoxia at night, particularly in slow-moving water bodies, as
the rate of oxygen consumption at the lower level of the bed cannot be replenished by
diffusion from the atmosphere (Mazzeo et al., 2003; Loverde-Oliveira et al., 2009).
Quantifying the impact of abundant macrophytes on basic water quality (oxygen
dynamics, nitrogen retention, and nutrient concentrations), Desmet et al (2011) found
that diurnal and seasonal fluctuations of DO were strongly correlated with plant
4. 3
growth/biomass density, temperature, and solar irradiances, observing water hypoxia
during the summer. Since light and temperature are triggers for biological processes,
these diurnal and seasonal shifts can be assigned to climatological conditions; such
findings match literature knowledge about the impacts of dominating macrophytes on
DO dynamics (Natural Heritage Trust, 2003; Tadesse et al., 2004; Hussner et al.,
2011; Riis et al., 2012).
An invasive non-native species (INNS) of particular concern is Lagarosiphon major
(Ridley) Moss. L (Hydrocharitaceae). Native to South Africa, this submerged
macrophyte is a considerable potential threat to static water bodies in many countries,
including the UK, where it is now well established (Figure 1) (Bowmer et al., 1995;
NNSS, 2011; J. Newman, pers comm). In addition to the aforementioned problems
associated with aquatic plant species invasion, L. major has been observed forming
dense canopies that often occupy entire water volumes of slow-moving water bodies
(Stiers et al., 2011). These thick mats block light penetration to other flora
(eliminating their growth), restrict water movement, and interfere with recreation
activities, ultimately exacerbating flood risks (Schwarz and Howard-Williams, 1993;
McGregor and Gourlay, 2002; Stiers, 2011).
Figure 1. GB and Ireland 2011 Ordnance Survey; Grid map for Lagarosiphon major (Ridl.)
Moss ex V.A. Wager [Curly Waterweed] (Botanical Society of Britain and Ireland, 2012; NBN
Gateway, 2013)
5. 4
As with other aquatic invasive species, L. major outcompetes native aquatic
vegetation and affects associated populations of species as it has a rapid growth rate
and is effectively perennial, surviving through the winter (Keenan et al., 2009; NNSS,
2011; J. Newman, pers comm). Furthermore, it is effective at removing CO₂ and
HCO3
− from water via photosynthesis, resulting in very high pH values that create
further complications for many aquatic vertebrate and invertebrate species (Sand-
Jensen, 1989; Hussner et al., 2014). Due to the decidedly invasive nature of L. major,
which has already been banned in New Zealand and Australia (Natural Heritage
Trust, 2003) it is an offence to plant or otherwise allow this species to grow in the
wild, under Schedule 9 of the Wildlife and Countryside Act (1981) regarding
England, Wales and Scotland (NNSS, 2016). However, natural checks on the growth
of L. major in the UK are insufficient, and control/eradication of the species is
extremely costly and often ineffective (Caffrey 1993b; Caffrey and Monahan 2006;
Stiers et al 2011; European Parliament, 2014).
Despite its environmental, ecological and economic impacts, L. major is a popular
water garden/aquarium plant, often mis-sold as Egeria or Elodea densa through the
aquatics industry (NNSS, 2011; J. Newman, pers comm). The UK population of L.
major has been intentionally planted as an ‘oxygenator’ and is often promoted
through the trade as one of the best (Natural Heritage Trust, 2003; Nault and
Mikulyuk, 2009; CBD, 2011; CABI, 2016; Royal Horticultural Society, 2016). Its
English common name ‘oxygen weed’- referring to the species’ ability to add oxygen
to the water as a result of its high photosynthetic rate (Rattray et al., 1994; CABI,
2016)- is the likely reason behind the industry’s promotion of the plant. However, not
only is this assumption unsupported by scientific evidence, but the high biomass
densities that are characteristic of this macrophyte species are likely to lead to a
higher consumption than production of oxygen, seriously limiting other aquatic
species (Natural Heritage Trust, 2003; Nault and Mikulyuk, 2009). This aspect of L.
major strongly opposes its designation as an oxygenating plant in the aquarium and
water garden industry and indicates that it should not be promoted as such.
Furthermore, the trade of this plant as an ornamental through the Internet and mail
order greatly increases its obtainability and ease of spread to new locations (Kay and
Hoyle, 2001; Australia Natural Heritage Trust, 2003; CABI, 2016).
6. 5
In 2014, The Wildlife and Countryside Act 1981 (prohibition on Sale etc. of Invasive
Non-native Plants) prohibited a number of invasive plants from sale in England due to
their adverse impacts on biodiversity and the economy (NNSS, 2016). Although L.
major was one of the species considered for prohibition of trade, the industry strongly
opposed, claiming that significant sacrifices had been made by not selling the five
banned species through a voluntary code. However, as stated by the UK’s leading
expert on research into aquatic invasive species (J. Newman, pers comm), there was
minimal countrywide financial loss from the ban (c. £10 - 25,000 per annum.) and the
exact value of L. major is still unknown. As a result of limited scientific evidence, L.
major escaped the ban, thus is still widely sold in the UK, despite there also being no
scientific reason not to ban it (J. Newman, pers comm).
Much of the existing literature regarding L. major focus on the core factors affecting
aquatic plant growth and morphology, such as temperature/light conditions (Desmet
et al., 2011; Riis et al., 2012), availability of carbon and nutrients (Hussner et al.,
2014), or competitive abilities (Stiers et al., 2011; Martin and Coetzee, 2014).
However, little attention has been paid primarily to the oxygenating abilities of
invasive L. major, with seemingly no existing scientific evidence that justifies the sale
of it as a designated “oxygenator”. Thus, in addition to filling this gap in the
literature, the study aims determine the efficacy of L. major as an oxygenator,
allowing the assessment of its fitness for sale in the UK. As a comparison species, the
UK native macrophyte, Ceratophyllum demersum (ridgid hornwort) was chosen based
on its similar morphological and growth characteristics to L. major. C. demersum is
also widely sold as an oxygenating plant and is recognised as invasive outside of it’s
natural range, although it doesn’t share as much of the inherent risks as L. major
(McGregor and Gourlay, 2002). The experiment will also measure and compare the
differences in growth (biomass) and associated pond life abundance and diversity of
the native and non-native plant species over 12 weeks to assess their invasiveness and
habitat impact.
7. 6
2. Materials and Methods
2.1. Plant Sample Collection
Healthy plant samples of Ceratophylum demersum (800g) and Lagarosiphon major
(800g) were collected from two adjacent ponds (A6 and A8) at Penrose Water
Gardens, Truro Cornwall (Figure 2) on October 7th 2015. As pond A8 was larger than
pond A6, samples were only collected within an area of similar size to pond A6
(boundary indicated by dotted line, Figure 2), ensuring that both species derived from
similar depth, light, temperature and growth conditions to limit the degree of variation
in morphology. Samples were collected with a rake and by hand from randomly
selected areas of the two ponds, avoiding sample selection bias and retaining sample
independence. All samples were rinsed thoroughly on site (within ponds), and again
later with settled tap water, to ensure no invertebrates or other plant species were
present; any found were returned to their respective ponds, or a nearby garden pond at
the study site.
2.2. Experimental Set-up
The experiment was conducted outdoors in Truro, Cornwall, between October 2015
and March 2016. Simulating small pond conditions, 200g of L. major and 200g of C.
demersum of similar size and root length were placed into plastic buckets containing 8
A8A6
Figure 2. Satellite map of ponds at Penrose Water Gardens, Truro, where C. demersum (A6) and L.
major (A8) samples were collected. Dotted black line indicates the boundary point of sample collection
(Google Maps,2016).
8. 7
litres of settled tap water. As utilised by Stiers et al (2011), tap water was left to stand
for over 48 hours prior to plant introduction and had a mean DO of 9.6 mg/L (s/d
0.17), with less than 0.5 range difference between buckets; this ensured that simulated
pond conditions were as similar as possible at the start of the experiment. In total, 12
buckets- four replicates of each species and four containing only water as controls-
were left to establish for six weeks (from 29/10/15 to 10/12/16) before any
measurements were taken. All buckets were labeled accordingly (species and replicate
number) and situated on a raised decked area approx. 0.5 m above ground level.
Samples were protected from elements, such as water run-off from heavy rainfall, as
the decked area provided sufficient drainage. As highlighted by many researchers
(Caraco et al., 2006; Desmet et al., 2011; Kemker, 2013; Ribaudo et al., 2014) the
presence and abundance of organisms and organic matter, and their associated
biological processes (e.g., respiration and decomposition) can greatly alter water
conditions. Gauze mesh coverings were considered to exclude organisms and organic
matter. However, in order to retain natural light irradiance and temperature of the
water, and in turn strengthen the integrity of the experiment (since contamination is a
natural occurrence in ponds), the buckets were instead, monitored daily for major
debris contamination (floating leaves, dead insects etc.); anything found was removed
with a sieve promptly.
2.3. Parameters Measured
After six weeks establishment, DO (mg/L [ppm]) and temperature (°C) of the water in
each bucket were measured twice a week for 12 weeks (10th December 2015 to 3rd
March 2016) using Hanna Instruments HI9142 portable waterproof dissolved oxygen
meter and a TPI-315C digital thermometer. Data collection always began at one hour
and 20 minutes after sunrise to control for diurnal effects on DO (informed by pilot
study, Appendix A) and collected in a balanced order to limit the degree of variation
between buckets over time. Another reason for applying a balanced order sampling
technique here was to eliminate systematic bias given the small sample size of 12
(Moore and McCabe, 2006). The DO meter was left for 15 minutes before any
measurements were taken to allow time for calibration, and a one-minute per bucket
time limit was allocated for DO and temperature readings. Along with DO and
temperature parameters, the date, time of day/since sunrise, weather conditions and
9. 8
water volumes of each bucket were also recorded on the relevant data record sheet
(Appendix B). Water levels were controlled every other day and kept at approx.
8000cm3 per bucket. If a significant amount of water was lost or gained as a result of
condensation or rainfall (i.e. +>5cm3 ->5cm3 of original volume), it was replaced with
settled tap water, or removed in order to keep conditions the same, prevent bias
readings, and reduce the potential of algal growth or algae bloom through water
replenishment (Paerl et al., 2001; Stiers et al., 2011). Buckets were rearranged every
other day, again, in a balanced order to provide consistency, and reduce the variation
of light and temperature climate across buckets, further diminishing bias (Stiers et al.,
2011). Plant biomass was measured in grams once every two weeks using Analogue
& Digital’s (A&D’s) EK-300i compact balance scales. Upon removal from water
buckets (by hand), plant samples were left to drain on top of a gauze mesh (placed
over the bucket) for one minute per plant to replenish water and prevent inaccurate
biomass readings due to added weight. It was during this stage of the experiment that
the associated pond life, i.e. invertebrate species were counted and recorded.
Biodiversity (some broadly classified as difficult to identify) included water slaters
(Asellus aquaticus), worms- including bloodworms, sludge worms (Tubifex tubifex)
and flatworms- caddisflies, shrimps (Crangonyx pseudogracilis) and snails (including
ramshorn) (see Appendix C). The weighing process presented an appropriate
opportunity to sufficiently inspect the plants and water bucket contents, whilst
causing the least disturbance to organisms present. Plastic, transparent tubs were used
to transfer plant samples from the gauze mesh to the scales and back into their
corresponding buckets, and also to transfer any organisms found safely to a nearby
pond (approx. 10 meters away from study site).
2.4. Data Analysis
All data was entered into a Microsoft Excel (2011) spread sheet, where all descriptive
statistics were also performed. All statistical analyses were carried out in Minitab® 17
Statistical Software (2010) with a significance value of 0.05. Although both
qualitative (observational) data and quantitative data were collected, only qualitative
data were analysed, reporting central tendencies (means ±) and variations (standard
errors) for each set. Treating the data non-parametrically (as indicated by normality
test results) a General Linear Model (GLM) was performed on all data sets.
10. 9
3. Results
The mean DO of the experimental pond conditions differed significantly in
concentration over the duration of the experiment (Figure 3A; treatment: F11, 264 =
53.4, p = 0.000; days: F24, 264 = 11.1, p < 0.001). Whilst the mean temperature
changed over the course of observations (Figure 3B; F24, 264 = 5003.4, p < 0.001),
there was no significant effect of treatment (Figure 3B; F11, 264 = 1.4, p = 0.160);
therefore temperature effects were not responsible for the DO changes between the
treatments.
0
2
4
6
8
10
12
14
16
Temperature(°C)
B
42
45
49
52
56
59
63
66
70
73
77
80
84
87
91
94
98
101
105
108
112
115
119
122
126
0
2
4
6
8
10
12
14
Days since establishment
Dissolvedoxygen(mg/L)
A
11. 10
Figure 3. Change in DO (A) and temperature (B) with time since start of observations.
Values are means ± S.E (n= 4). Three treatments indicated by colours and symbols (red ○=
control, blue □ = L. major,green △= C. demersum).
The differences in mean pond life abundances associated with native C. demersum
and non-native L. major samples were significant (Figure 4; F7, 42 = 4.6, p = 0.001).
Establishment time of the plants also had a significant effect on the number of
invertebrate species present (Figure 4: F6, 42 = 3.1, p = 0.013). Pond life abundances of
C. demersum reached a total count of 80 over the study period, while L. major totaled
only 14 and exhibited a less diverse array (see Appendix C).
Figure 4. Comparison of pond life abundance associated with the native and non-native
plants over time. Values are means ± S.E (n= 4). Two treatments indicated by colours and
symbols (blue □= L. major,△ green = C. demersum).
The biomass of the two species differed significantly during the experimental period
(Figure 5; F7, 42 = 70.3, p = 0.000). However, establishment time had no significant
effect on growth patterns (Figure 5; F6, 42 = 1.5, p = 0.209). The biomass of L. major
samples increased overall (mean 233.10g), whereas C. demersum samples decreased
slightly (mean 198.27g) from the initial 200g start weight.
0
2
4
6
8
10
12
42 56 70 84 98 112 126
Pondlifeabundance
Days since establishment
12. 11
Figure 5. C Change in biomass with time since start of observations. Values are means ± S.E
(n= 4). Two treatments indicated by symbols and colours (□ blue = L. major, △ green = C.
demersum).
Qualitatively, there was a notable difference, particularly with L. major, in the
appearance of the samples at the start of the experiment compared to the end (Figures
6 and 7). Similarly, after measurements had ceased and the simulated pond conditions
were left to establish for a further two weeks (17/03/16), there were striking visual
differences between native and non-native plants. All L. major replicates were in a
state of decomposition, unlike C. demersum replicates, which appeared to remain in a
healthy condition (Figure 8).
Figure 6. L. major (left) and Ceratophyllumdemersum (right) samples at the start of the experiment.
0
50
100
150
200
250
300
1 42 56 70 84 98 112 126
Biomass(g)
Days since establishment
13. 12
Figure 7. L. major (left) and Ceratophyllumdemersum (right) samples at the end of the experiment.
Figure 8. Water buckets containing L. major samples (left) and C. demersum (right) two weeks post
experiment.
4. Discussion
This research demonstrates for the first time the oxygenating efficacies, growth rates
and associated pond life abundances of two macrophyte species, native, C. demersum,
and non-native, L. major, established in small pond conditions. The results showed
significant differences between the species’ DO concentrations over time, with all
replicates of C. demersum better than L. major in relation to maintaining healthy
levels of DO in its surrounding environment. As exemplified in Figure 3A, while C.
demersum increased DO levels over time, L. major caused levels to decline.
Fluctuations in overall DO concentrations correlated with temperature changes
(Figure 3A and B) and were explicable in terms of the known effects of temperature
14. 13
on DO; as temperature increases, the solubility of oxygen decreases as gases are
typically more soluble at colder temperatures (Tadesse et al, 2004; Desmet et al.,
2011; Hussner et al., 2011; Riis et al., 2012; Kempker, 2013). For example at around
49, 59, 87, and 115 days from establishment, where mean temperatures reached
maxima (13.2, 13.3, 12.5 and 10.7°C respectively), mean DO concentrations of all
samples decreased considerably (Figure 3A). Water temperatures ranged from a
minimum 1°C (December 2015) to a maximum of 13.4 °C (March 2016), which
changed significantly over time. This change with time was another expected
observation that matched literature knowledge relating to the effects of
meteorological conditions on temperature (seasonal shifts) (Tadesse et al., 2004;
Desmet et al., 2011; Hussner et al., 2011; Riis et al., 2012). However, as there were
no differences between the temperatures of the treatments (Figure 3B, Appendix D), it
was clear that temperature was not the causation of the significant DO variations that
occurred between the treatments; these variations were likely as a result of the plants’
differing photosynthetic activities and capabilities.
The pond life associated with L. major and C. demersum (pond life was absent in the
control buckets) ranged from 0 to 13 freshwater invertebrate “species” from a single
sample observation. C. demersum consistently had higher associated pond life
abundance, with a more diverse collection than L. major- often there was no
associated biodiversity (Figure 4, Appendix C). Although little has been published on
the preferences and tolerance levels of native fish and invertebrate species for DO,
previous research has documented that the requirement for most freshwater fish is
greater than 6 mg/L, and around 5 mg/L for freshwater insects (Davis, 1975,
Appendix E). Wurts (1993) proposed that DO levels less than 3 mg/L are insufficient
to support aquatic life (e.g. fish), with more recent literature (Behar, 1996; Leppi et
al., 2016), suggesting that many freshwater organisms will be adversely affected
when DO falls below a level of 2 mg/L of saturation for prolonged periods. Whist DO
reached a maximum of 11.9 mg/L in C. demersum and control buckets over the
course of the experiment, one L. major replicate caused DO to fall to a minimum of
1.1 mg/L (Appendix F); this is well below the level which is classed as sustainable for
most aquatic life. Other L. major replicates often fell below the recommended healthy
requirements for native freshwater invertebrates (5 mg/L, Appendix E), with even the
mean values (4.5 mg/L and 3.9 mg/L) falling to near-lethal levels on several
15. 14
occasions (Figure 3A). As highlighted in the literature, oxygen availability is known
to be a major factor determining the occurrence and abundance of many aquatic
communities (Ruse, 1996; Gabriels et al., 2007; Desmet et al., 2011) as low DO
concentrations characterise diminished water quality and have adverse effects on
associated species (Hussner et al., 2014). This can explain why L. major consistently
had significantly less associated biodiversity than C. demersum- particularly evident
in L. major replicate 1 (Appendix C), which had the lowest mean DO overall (5.4
mg/L) and no associated pond life over the study period. The significant effect of time
on pond life abundance can also be explained by the significant effect of time on DO,
which increased with C. demersum, and decreased with L. major samples. Across all
C. demersum samples, levels never fell below 6.4 mg/L throughout the study period
(Appendix F), thus, were consistently sustainable for aquatic life. Furthermore,
literature states that certain species may be indicators of water quality. For example,
shrimps (Crangonyx pseudogracilis), which were only associated with C. demersum
samples, are often only present in good quality ponds (Freshwater Habitats Trust,
2016). The Freshwater Habitats Trust (2016) also highlighted that the presence of
organisms such as cassisflies, which were abundant in C. demersum pond conditions
but absent in L. major (Appendix C), and water snails may mean that the water
quality is relatively good. On the contrary, the presence of pollution-tolerant species
such as sludge worms (Tubifex tubifex) and water slaters (Asellus aquaticus) may be
indicators of relatively poor water quality (Freshwater Habitats Trust, 2016). Water
slaters and sludge worms were the most abundant species associated with L. major,
with water snails (after shrimps and caddisflies) being the least abundant species
(only 2 compared with the 10 associated with C. demersum). Overall, these findings
demonstrate that the native plant was the preferred habitat for freshwater invertebrates
over the non-native, which created an unhealthy environment and unsuitable living
conditions for such species.
Outside of its normal growing season, L. major outcompeted the native species in
terms of overall growth, with an end mean weight of 233.1g compared to the 198.3g
mean of C. demersum samples. L. major not only increased in biomass, but also
exhibited a wider variability in growth patterns across replicates, deviating quite far
from its initial 200g start-weight at times (Appendix G). In comparison, the growth
patterns of C. demersum plants were less erratic and managed to maintain a relatively
16. 15
steady growth throughout the experiment (Appendix G). This indicates that L. major
has an ability to be more invasive, with high unpredictability in its growth rates,
which poses many issues when implementing guidelines in relation to the trade and
promotion of this species for aquarium and pond use. The biomass findings from this
study are in line with previous research results (Rattray, 2004; Stiers et al., 2011;
Martin and Coetzee, 2014). Rattray et al (2004) revealed that, in comparison to the
macrophyte Myriophyllum triphyllum, L. major has a greater ability to increase both
height and biomass during the colonisation stage. A similar study by Stiers et al
(2011), using a direct comparison of the two species used in this experiment (in
similar pond conditions), found that L. major outperformed C. demersum in relative
growth rate (RGR) (based on total length and weight) under two different sediment
conditions. More recently, in a comparison of the competitive abilities of L. major
and Myriophyllum spicatum, Martin and Coetzee (2014) found that L. major had a
faster RGR and was overall a superior competitor to M. spicatum.
However, as observed in Ranunculus circinatus by Larson (2007) and Myriophyllum
spicatum by Angelstein et al (2009), any treatment used for manipulating the plants
(i.e. by hand when weighing) can be a potential stress factor and impose loss of
vitality. Thus, this may have been an influencer of the weight differences observed
between the two species in this experiment, as well as to the decomposition and
fragmentation observations. After only a few weeks of establishment, although
fragmentation of both species was observed, it was more apparent in C. demersum
samples. By January, one L. major replicate (L.m 1, Appendix E) was beginning to
decompose, and was severely decomposed by February. As stated by Rattray (1994)
and Nault and Mikulyuk (2009) decomposing mats of L. major create extremely low
oxygen levels in the water, which clarifies the consistently low DO concentrations of
that particular sample (lowest DO readings overall- 1.1mg/L). However, these
observations do not concur with the literature that states that L. major is effectively
perennial (Keenan et al., 2009) as none of the samples survived through the winter
and were all heavily decomposed by the end of the experiment (Figure 8). This may
be because of the small, simulated pond conditions representing an over-
simplification of reality, which limits the ability to extrapolate the results to a natural
ecosystem. Furthermore, although many submerged macrophytes are able to tolerate
changes in temperature well (Rooney and Kalff, 2000), L. major is thought to be
17. 16
unable to withstand temperatures below 10°C, dying or becoming dormant when
exposed (Australia Natural Heritage Trust, 2003; CABI, 2016). Therefore, the mean
temperature of 7.29°C over the data collection period may have been a contributing
factor for the decomposing/dying plants.
However, even outside of the species’ usual growth season, and with findings limited
by low temperature (considered minimal given that its optimum is 20-23°C), L. major
still grew rapidly and caused an oxygen depletion. This strongly suggests that the
impacts associated with L. major (rapid growth, diminished DO) will be exacerbated
during its growth season (Wilcock, 1998). Furthermore, although data from the Met
Office (2016) on the provisional mean temperature for the UK was below the 1981-
2010 long-term average, global surface temperate data from NASA (2016) has
reached an all time high, which is predicted to rise. Elevated temperatures and
increased light irradiation are likely to significantly increase L. major growth rates
and heighten invasion risks, further impacting DO and threatening oxygen-dependent
organisms (Hussner et al., 2011).
While longitudinal studies, conducted on natural ponds over the summer months
(typical growth period) are recommended to strengthen the validity of this study’s
findings, the results clearly suggest that invasive non-native L. major has detrimental
impacts on its freshwater environment. As this species was not an efficient
oxygenator (quite the opposite of its sale title) results could inform current practice
and legislation negotiations in relation to the legal trade of L. major in the UK,
offering a safer, more effective alternative (C. demersum) to the aquatics oxygenating
plant industry. Contributions towards a wider body of related research and
organisations/action groups, such as GB Non-Native Species Secretariat (NNSS), the
Department for Environmental food and Rural Affairs (Defra), Natural England, the
Environment Agency, and Student Non-Native Invasive Group (SINNG), may also be
offered from these findings.
In conclusion, it is recommended that invasive non-native L. major, which is already
on the EU draft List of Species of Union Concern, should fall under Order 14 of The
Wildlife and Countryside Act 1981 (prohibition on Sale etc. of Invasive Non-native
Plants) (England) due to its significant negative impacts on biodiversity and
18. 17
ecosystems. A ban on the trade of L. major will help to eliminate further spread and
be a positive move towards dealing with the negative consequences it has had for the
environment, ecosystem services, public health and the economy in Europe.
Acknowledgements
First and foremost, I would like to express my sincere gratitude to my supervisor Dr
Peter McGregor for the continuous support of my BSc honours degree and related
research, for his patience, motivation, and immense expertise. His invaluable
19. 18
guidance assisted me through all stages of the research and writing of this thesis, as
well as eased any apprehensions that arose. I could not have asked for a better advisor
for my study and cannot thank him enough.
My sincere thanks also goes to SINNG project coordinator, and joint-mentor Nicola
Morris, who enlightened me with the initial proposal of this project. Her immense
knowledge on the topic provided me with insight that greatly assisted the research
from the onset to completion. Gratitude is also pledged to Trevor Renals of the
Environment Agency, and UK leading expert on research into aquatic invasive
species, Jonathan Newman, for their shared expertise and technical assistance that
proved fundamental for my literature review and subsequent understanding of the
topic.
Besides my supervisor, I would like to thank the rest of my thesis committee: Dr
Angus Jackson, Kelly Haynes and Thais Martin for their insightful suggestions and
encouragement. Also, thanks goes to Ruth Martin and Andrew Golley for the
conference questioning and comments, which incented me to widen my research from
various perspectives.
Last but certainly not least, I would like to thank my children, Tahia, Remaeus and
Amari for supporting me fully throughout my four years in education, and for your
unconditional love. You are my motivation for success.
References
Angelstein, S., Wolfram, C., Rahn, K., Kiwel, U., Frimel, S., Merbach, L. and
Schubert, H. (2009) ‘The influence of different sediment nutrient content on growth
20. 19
and competition of Elodea nuttalli and Myriophyllum spicatum in nutrient-poor
waters’, Fundamental and Applied Limnology, 175(1), pp.49-57.
Annis, R.B. (2014) Water Resources Institute: Dissolved Oxygen. Available at:
https://www.gvsu.edu/wri/education/instructors-manual-dissolved-oxygen-30.htm
(Accessed: 15 March 2015).
Australia Natural Heritage Trust (2003) Lagarosiphon - Lagarosiphon major. Weed
Management Guide. Australia: Natural Heritage Trust.
Baird, D., Christian, R.R., Peterson, C.H. and Johnson, G.A. (2004) ‘Consequences of
hypoxia on estuarine ecosystem function: Energy diversion from consumers to
microbes’, Ecological Applications, 14, pp. 805–822.
Botanical Society of Britain and Ireland (2011) ‘Grid map for Lagarosiphon major
(Ridl.) Moss ex V.A. Wager [Curly Waterweed]’, Updated 2012-2013. Available at:
<http://data.nbn.org.uk/Taxa/NHMSYS0000460065/GridMap> (Accessed: 2 April
2016).
Bowmer, K.H., Jacobs, S.W.L. and Sainty, G.R. (1995) ‘Identification, biology and
management of Elodea canadensis, Hydrocharitaceae’, Journal of Aquatic Plant
Management, 33, pp. 13-19.
CABI (2016) Largarosiphon major (Afican elodea). Available at:
http://www.cabi.org/isc/datasheet/30548 (Accessed: 2 April 2016).
Caffrey, J.M. (1993) ‘Plant management as an integrated part of Ireland’s aquatic
resource’, Hydroécologie Appliquée, 5, pp. 77-96.
Caffrey, J.M. and Monahan, C. (2006) ‘Control of Myriophyllum verticillatum L. in
Irish canals by turion removal’, Hydrobiologia, 570(1), pp. 211-215.
Caraco, N., Cole, J., Findlay, S. and Wigand, C. (2006) ‘Vascular plants as engineers
of oxygen in aquatic systems’, BioScience, 56(3), pp. 219-225.
21. 20
Carrillo, Y., Guarín, A. and Guillot, G. (2006) ‘Biomass distribution, growth and
decay of Egeria densa in a tropical high-mountain reservoir (NEUSA, Colombia)’,
Aquatic Botany, 85, pp. 7-15.
CBD (The Convention on Biological Diversity) (2011) Information about GB Non-
native Species Risk Assessments. Available at:
file:///Users/annettelumb/Downloads/RA_Lagarosiphon_major_(Curly_Waterweed).p
df (Accessed: 18 November 2015).
Gabriels, W., Goethals, P.L.M., Dedecker, A.P., Lek, S. and De Pauw, N. (2007)
‘Analysis of macrobenthic communities in Flanders, Belgium, using a stepwise input
variable selection procedure with artificial neural networks’, Aquatic Ecology, 41, pp.
427-441.
Google Maps (2016) Penrose Water Gardens. Available at:
https://www.google.co.uk/maps/place/Penrose+Water+Gardens/ (Accessed: 20 March
2016).
Davis, J.C. (1975) ‘Minimal Dissolved Oxygen Requirements of Aquatic Life With
Emphasis on Canadian Species: A Review’, Journal of the Fisheries Research Board
of Canada, 32(12), pp. 2295-2332.
Desmet, N.J.S., Van Belleghem, S., Seuntjens, P., Bouma, T.J., Buis, K. and Meire, P.
(2011) ‘Quantification of the impact of macrophytes on oxygen dynamics and
nitrogen retention in a vegetated lowland river’, Physics and Chemistry of the Earth,
Parts A/B/C, 36(12), pp. 479-489.
Dudgeon, D., Arthington, A.H., Gessner, M.O., Kawabata, Z.I., Knowler, D.J.,
Leveque, C., Naiman, R.J., Priur-Richard, A.H., Soto, D., Stiassny, M.L.J. and
Sullivan, C.A. (2006) ‘Freshwater biodiversity: importance, threats, status and
conservation challenges’, Biological Reviews, 81, pp. 163-182.
22. 21
Dybas, C.L. (2005) ‘Dead zones spreading in world oceans’, BioScience, 55, pp. 552-
557.
EPA (2012) What are Suspended and Bedded Sediments (SABS)? Available at:
http://water.epa.gov/scitech/datait/tools/warsss/sabs.cfm (Accessed: 28 February
2016).
European Parliament (2014) Invasive Alien Species. Available at:
http://www.europarl.europa.eu/RegData/etudes/workshop/join/2014/518746/IPOL-
ENVI_AT(2014)518746_EN.pdf (Accessed: 11 October 2015).
Fenchel, T., King, G.M. and Blackburn, T.H. (1998) Bacterial Biogeochemistry: The
Ecophysiology of Mineral Cycling. San Diego: Academic Press.
Freshwater Habitats Trust (2016) Shrimp. Available at:
http://freshwaterhabitats.org.uk/habitats/pond/identifying-creatures-pond/shrimp/
(Accessed: 10 May 2016).
Gray, J.S., Wu, R.S. and Or, Y.Y. (2002) ‘Effects of hypoxia and organic enrichment
on the coastal marine environments’, Marine Ecological Progress Series. 238,
pp. 249-279.
Hogan, L. (2008) Call for action as pondweed threatens major fishing lake. Available
at: http://www.independent.ie/irish-news/call-for-action-as-pondweed-threatens-
major-fishing-lake-26450894.html (Accessed: 24 March 2016).
Hussner, A., Hofstra, D. and Jahns, P. (2011) ‘Diurnal courses of net photosynthesis
and photosystem II quantum efficiency of submerged Lagarosiphon major under
natural light conditions’ Flora, 206, pp.904-909.
Hussner, A., Hofstra, D., Jahns, P. and Clayton, J. (2014) ‘Response capacity to CO2
depletion rather than temperature and light effects explain the growth success of three
alien Hydrocharitaceae compared with native Myriophyllum triphyllum in New
Zealand’, Aquatic Botany, 120, pp. 205-211.
23. 22
J. Newman (2015) Email from Dr. Jonathan J. Newman, pers comm, 9 October 2015.
Kay, K.H. and Hoyle, S.T. (2001) ‘Mail Order, the Internet, and Invasic Aquatic
Weeds’, Journal of aquatic Plant Management, 39(1), pp.88-91
Keenan, E. Baars, J-R. and Caffrey, J.M. (2009) ‘Changes in littoral invertebrate
communities in lough corrib in response to an invasion by Lagarosiphon major’, in:
Pieterse, A., Rytkonen, A-M. and Hellsten, S. (eds.) Aquatic Weeds 2009. Finland:
Finnish Environment Institute, pp. 24-28.
Kemker, C. (2013) Dissolved Oxygen: Fundamentals of Environmental
Measurements. Fondriest Environmental. Available at: http://www.
http://www.fondriest.com/environmental-measurements/parameters/water-
quality/dissolved-oxygen/#2 (Accessed: 22 February 2016).
Larson, D. (2007) ‘Growth of three submberged plants below different densities of
nymphoides peltara (SG, Gmel) Kuntze’, Aquatic Botany, 86, pp. 280-284.
Lenntech (2015) Why oxygen dissolved in water is important. Available at:
http://www.lenntech.com/why_the_oxygen_dissolved_is_important.htm (Accessed:
27 January 2015).
Leppi, J.C., Arp, C.D. and Whitman, M.S. (2016) ‘Predicting Late Winter Dissolved
Oxygen Levels in Arctic Lakes Using Morphology and Landscape Metrics’,
Environmental Management, 57, pp. 463-473.
Loverde-Oliveira, S.M., Moraes Huszar, V.L., Mazzeo, N. and Scheffer, M. (2009)
‘Hydrology-driven regime shifts in a shallow tropical lake’, Ecosystems 12(5), pp.
807-819.
Martin, G. D. and Coetzee, J.A. (2014) ‘Competition between two aquatic
macrophytes, Lagarosiphon major (Ridley) Moss (Hydrocharitaceae)
24. 23
and Myriophyllum spicatum Linnaeus (Haloragaceae) as influenced by substrate
sediment and nutrients’, Aquatic Botany, 114, pp. 1-11.
Mazzeo, N., Rodríguez-Gallego, L., Kruk, C., Meerhoff, M., Gorga, J., Lacerot, G.,
Quintans, F., Loureiro, M., Larrea, D. and García-Rodríguez, F. (2003) ‘Effects of
Egeria densa Planch. beds on a shallow lake without piscivorous fish’, Hydrobiologia,
506(1), pp. 591-602.
McGregor, P. G., & Gourlay, H. (2002) Assessing the Prospects for the Biological
Control of Lagarosiphon (Lagarosiphon major (Hydrocharitaceae)). New Zealand:
Department of Conservation.
Mitchell-Holland, R. (2016) Lagarosiphon major and Ceratophyllum demersum
[photograph] (authors own collection).
Morris, K., Harrison, K.A., Bailey, P.C.E. and Boon, P.I. (2004) ‘Domain shifts in the
aquatic vegetation of shallow urban lakes: The relative roles of low light and anoxia
in the catastrophic loss of the submerged angiosperm’, Vallisneria Americana’,
Marine and Freshwater Research, 55, pp. 749-758.
NASA (2016) Global temperature. Available at: http://climate.nasa.gov/ (Accessed:
10 May 2016).
Natural Heritage Trust (2003) Lagarosiphon – Lagarosiphon major. Weed
Management Guide. Canberra, Australia: Department of Sustainability, Environment,
Water, Population and Communities. Available at:
http://www.weeds.gov.au/publications/guidelines/alert/pubs/l-major.pdf (Accessed: 6
October 2015).
Nault, M.E. and A. Mikulyuk. (2009) African Elodea (Lagarosiphon major): A
Technical Review of Distribution, Ecology, Impacts, and Management. Madison,
Wisconsin, USA: Wisconsin Department of Natural Resources Bureau of Science
Services.
25. 24
NNSS (2011) Information about GB Non-native Species Risk Assessments. Available
at:file:///Users/annettelumb/Downloads/RA_Lagarosiphon_major_(Curly_Waterweed
).pdf (Accessed: 10 March 2015)
NNSS (2016) England and Wales: The Countryside Act 1981. Available at:
http://www.nonnativespecies.org//index.cfm?pageid=67 (Accessed: 2 January 2016).
Paerl, H.W., Fulton, R.S., Moisander, P.M. and Dyble, J. (2001) ‘Harmful Freshwater
Algal Blooms, With an Emphasis on Cyanobacteria’, The Scientific World Journal, 1,
pp. 76-113.
Pimentel, D., Lach, L., Zuniga, R. and Morrison, D. (2000) ‘ Environmental and
economic costs of nonindigenous species in the United States’, Bioscience, 50(1), pp.
53-65.
Rattray, M.R., Howard-Williams, C. and Brown, J.M. (1994) ‘Rates of early growth
of propagules of Lagarosiphon major and Myriophyllum triphyllum in lakes of
differing trophic status’, New Zealand Journal of Marine and Freshwater Research,
28(3), pp. 235-241.
Ribaudo, C., Bertrin, V. and Dutartre, A. (2014) ‘Dissolved gas and nutrient dynamics
within an Egeria densa Planch. bed’, Acta Botanica Gallica, 161(3), pp. 233-241.
Riis, T., Olsen, B., Clayton, S.J., Lambertini, C., Brix, H. and Sorrell, K.B. (2012)
‘Growth and morphology in relation to temperature and light availability during the
establishment of three invasive aquatic plant species’ Aquatic Botany, 102, pp. 56-64.
Rooney, N., Kalff, J and Habel, C. (2003) ‘The role of submerged macrophyte beds in
phosphorus and sediment accumulation in Lake Memphremagog, Quebec, Canada’,
Limnology Oceanography, 48(5), pp, 1927-1937.
Royal Horticultural Society (2016) Lagarosiphon major (curly waterweed). Available
at: https://www.rhs.org.uk/Plants/9805/Lagarosiphon-
major/Details?returnurl=%2Fplants%2Fsearch-results (Accessed: 29 March 2016).
26. 25
Ruse, L.P. (1996) ‘Multivariate techniques relating macroinvertebrate and
environmental data from a river catchment’, Water Research, 30, pp. 3017-3024.
Sand-Jensen, K. (1989) ‘Environmental variables and their effect on photosynthesis
of aquatic plant communities’, Aquatic Botany, 34, pp. 5-25.
Schwarz, A. and Howard-Williams, C. (1993) ‘Aquatic weed bed structure and
photosynthesis in two New Zealand lakes’, Aquatic Botany, 46, pp. 263-281.
Stiers, I., Njambuya, J. and Triest, L. (2011) ‘Competitive abilities of invasive
Lagarosiphon major and native Ceratophyllum demersum in monocultures and mixed
cultures in relation to experimental sediment dredging’, Aquatic Botany, 95(2), pp.
61-166.
Strayer, D.L., Lutz, C., Malcom, H.M., Munger, K., and Shaw, W.H. (2003)
‘Invertebrate communities associated with a native (Vallisneria americana) and an
alien (Trapa natans) macrophyte in a large river’, Freshwater Biology, 48(11), pp.
1938-1949.
Tadesse, I., Green, F.B. and Puhakka, J.A. (2004) ‘Seasonal and diurnal variations of
temperature, pH and dissolved oxygen in advanced integrated wastewater pond
system® treating tannery effluent’, Water Research, 38(3), pp.645-654.
Wilcock, R.J., Nagels, J.W., McBride, G.G., Collier, K.J., Wilson, B.T. and Huser,
B.A. (1998) ‘Characterisation of lowland streams using a single‐ station diurnal
curve analysis model with continuous monitoring data for dissolved oxygen and
temperature’, New Zealand Journal of Marine and Freshwater Research, 32(1), pp.
67-79.
Westphal, M.I., Browne, M., MacKinnon, K. and Noble, I. (2008) ‘The link between
inter-national trade and the global distribution of invasive alien species’, Biological
Invasions, 10, pp. 391-398.
27. 26
Wetzel, R. G. (2001) Limnology: Lake and River Ecosystems. 3rd edn. San Diego,
CA: Academic Press.
Wurts, W.A. (1993) ‘Dealing with oxygen depletion in ponds’, World Aquaculture,
24(2), pp. 108-109.
Appendices
Appendix A: Pilot study
The pilot study was carried out over 11 hours on 19th October 2015. 100g of each L.
major and C.demersum samples were established for two weeks in small tubs
containing 2.4 L of settled tap water before DO was measured every hour (from 9am-
8pm). This helped to establish the diurnal patterns of DO fluctuations of L. major and
C. demersum and represented the times that were most appropriate to take
measurements (1 hour and 20 minutes from sunrise- indicated by arrow on the figure
below). Although this time indicates the fastest rate of change, this time was chosen
because it represented a point where low DO concentrations were likely to occur,
ultimately determining whether levels fall below that of which are supportive of
aquatic life (aim of the study). It was also chosen for logistical reasons, being the
most suitable time of day for me to take undisturbed measurements, consistently. The
experiment aimed to control for as many confounding variables as possible, including
28. 27
limiting diurnal effects (by taking measurements at the same time after sunrise) and
measuring temperature to determine if any differences between treatments were
present (affecting DO outcomes).
Appendix B: Data record sheet example
Sample:
Day/Date Establ.
time
Hrs since
sunrise
D.O.
mg/L
Temp
(°C)
Weight
(g)
W/level
(cm³)
Weather
conditions
1. 10/12/15
2. 13/12/15
3. 17/12/15
4. 20/12/15
5. 24/12/15
6. 27/12/15
7. 31/12/15
8. 3/01/16
9. 7/01/16
10. 10/01/16
11. 14/01/16
12. 17/01/16
13. 21/01/16
0
2
4
6
8
10
12
DissolvedOxygen
hours from sunrise
Pilot Study
L. major
C.demersum
Sunrise: 7:49 am, measurements taken: 9:09 am
29. 28
14. 24/01/16
15. 28/01/16
16. 31/01/16
17. 4/02/16
18. 7/02/16
19. 11/02/16
20. 15/02/16
21. 18/02/16
22. 21/02/16
23. 25/02/16
24. 28/01/16
25. 3/03/16
Species
Sample
Water slaters Worms Flatworms Caddiflies Shrimps Snails
L. m 1 0 0 0 0 0 0
L. m 2 3 6 1 0 0 2
L .m 3 1 1 0 0 0 0
L. m 4 0 0 0 0 0 0
C. d 1 1 0 1 4 0 6
Pond life abundance/biodiversity:
Additional notes/observations:
30. 29
Appendix C. Associated pond life diversity
The associated biodiversity of pond life found over the duration of the experiment are
displayed in the table below. Species observed (some broadly classified as difficult to
identify) included water slaters (Asellus aquaticus), worms, including bloodworms,
sludge worms (Tubifex tubifex) and flatworms, caddisflies, shrimps (Crangonyx
pseudogracilis) and water snails (including ramshorn). There was a total of 14 pond
life abundances associated with L. major samples, consisting of 4 different species
(water slaters, worms, flatworms and water snails). C. demersum had an abundance
total of 8, consisting of 6 different species (all listed on table).
Appendix D. Temperature variations
As displayed in the below temperature graph, there was very little variation between
the temperatures of L. major, C. dersumum, and the control buckets (means of all
replicates) which indicates that temperature had no significant effect on the DO
variation between the samples. Black arrows indicate where the water of all buckets
froze over during the study (at 119 days, 25/02/16)
C. d 2 1 2 1 4 0 1
C. d 3 25 1 0 4 2 2
C. d 4 21 1 0 2 0 1
Total L.m = 4
C. d =48
L. m = 7
C. d = 4
L. m = 1
C. d = 2
L. m = 0
C. d = 14
L. m = 0
C. d = 2
L. m = 2
C. d = 10
Grand
Total
L. major = 14 C. demersum = 80
31. 30
Appendix E. Dissolvedoxygen (mg/L) tolerance ranges for aquatic life
Previous research has documented that most freshwater fish require DO levels greater
than 6 mg/L, with insets requiring levels around 5 mg/L (Davis, 1975), with lethal
levels potentially occurring below 2 mg/L (Leppi et al., 2016). Below, Behar (1996)
suggests guidelines for the range of tolerance for DO in aquatic life forms (and
interpretation of DO readings) (Adapted from Behar, 1996)
42
45
49
52
56
59
63
66
70
73
77
80
84
87
91
94
98
101
105
108
112
115
119
122
126
0
2
4
6
8
10
12
14
16
Days since establishment
Temperature(°C)
Temperature (mean +/- se)
Control
C. demersum
L. major
0-2 mg/L: not enough oxygen to support life
2-4 mg/L only a few kinds of fish and insects can survive
4-7 mg/L: acceptable for warm water fish
7-11 mg/L: very good for most stream fish including cold water fish
32. 31
Appendix F. Actual numbers of DO from all replicates
The below figure of individual replicates (actual numbers) shows that one replicate of
L. major (sample 1) fell to 1.1 mg/L; this is below the tolerance threshold for most
aquatic life forms, indicated by dotted red line (see Appendix E, Behar, 1996). Other
L. major replicates also fell below the recommended healthy requirements for native
freshwater invertebrates (indicated by green dotted line) (Davis, 1975).
Appendix G. Growthpatterns of all sample replicates
As shown on the figure below, the growth patterns of C. demersum (difference from
initial value) were less variable than those of L. major, which, in some replicates,
expressed erratic growth rates that deviated quite far from the initial weight.
0
1
2
3
4
5
6
7
8
9
10
11
12
13
42
45
49
52
56
59
63
66
70
73
77
80
84
87
91
94
98
101
105
108
112
115
119
122
126
DissolvedOxygen(mg/L)
Days since establishment
DO (actual numbers) Control 1
Control 2
Control 3
Control 4
C.d 1
C.d 2
C.d 3
C.d 4
L.m 1
L.m 2
L.m 3
L.m 4
1.1mg/L
