580
P. J. Ralph, D. Tomasko, K. Moore, S. Seddon, and C. M. O. Macinnis-Ng
exposure. These regional differences could be indicative of chronic sub-lethal acclimation, or that
the plants from the southern site were less tolerant.
This suggests that tolerance is not constant within
a species across a wide geographic distance and
that local microenvironment strongly influences the
tolerance.
6. Recovery from Herbicides
Few seagrass studies have seriously considered recovery from chronic exposure to herbicides. Recovery from acute herbicide exposure is a common response for macrophytes (including seagrasses), but
this needs to be considered from a more ecological
perspective. Bowmer (1986) demonstrated that the
PSII inhibitory effect of herbicides (such as diuron
and atrazine) on macrophytes was reversed once
the contaminant was removed, and photosynthesis
rapidly recovered. This aspect of herbicide toxicology is rarely considered in laboratory experiments
and is a critical consideration in the interpretation
of any field based herbicide exposure experiment.
Haynes et al. (2000b) found three species of seagrass
recovered when returned to fresh seawater after 5days exposure to diuron. While Macinnis-Ng and
Ralph (2003a) found a virtual halt in photosynthetic
activity induced by exposure to atrazine, irgarol and
diuron was reversed on return to fresh seawater in
the laboratory and field. The exposure period was
only 10 h and it is not known if the seagrass would
recover after a longer exposure period. Field conditions would typically resemble a single pulse of contaminant, followed by a period of flushing, followed
by a subsequent pulse of contaminant (Solomon
et al., 1996), and experiments are needed which simulate this effect.
Assessing endpoints such as growth may represent a more ecologically relevant outcome, but it
may not be the most sensitive indicator of stress,
as the plant has a longer period of time for phytoremediating processes to detoxify or adapt to the
contaminant. Conversely, a more sensitive endpoint
may illustrate an impact, which could possibly be reversed by returning the plant to un-contaminated water. For example, Z. marina exposed to 5–10 µg L
−1
atrazine for 5 weeks significantly reduced growth
(Schwarzschild et al., 1994), whereas Z. capricorni
after just 10 h exposure showed a significant impact
on PSII photochemical efficiency, although this impact was completely reversible in situ (Macinnis-Ng
and Ralph, 2003a). This clearly demonstrates the
different sensitivities of these two partially related
endpoints.
D. Heavy Metals
1. Sources of Heavy Metals
Heavy metals can enter the marine environment
naturally via weathering and erosion of rocks and
soils (Batley, 1996) or through urban and municipal runoff, storm water, sewage, industrial effluents, mining operations, atmospheric deposition and
agricultural activity (Batley, 1996; Irvine and Birch,
1998; Haynes and Johnson, 2000). In addition, activities such as dredging can release heavy metal-laden
sediment from anoxic deeper layers allowing metals
to be re-mobilized and available to contaminate the
biota (Batley, 1987; Lytle and Lytle, 2001).
2. Processes of Heavy Metal Contamination
Bio-availability of heavy metals is influenced by water and sediment pH, sediment particle size, redox
potential, dissolved organic matter, sediment cation
exchange capacity, water temperature, salinity, organic content, and the concentrations of other metals (Batley, 1987; Ward, 1989; Batley et al., 1999;
Alutoin et al., 2001). These physicochemical parameters influence processes that affect heavy metal
availability, such as chelation, complexing, precipitation, absorption, and adsorption (Batley, 1987).
For this reason, as well as the more complex nature of the field environment, heavy metals are less
bio-available and generally have a less toxic effect
in the field that the laboratory (Graney et al., 1995),
thus extrapolating laboratory findings to the field can
be unreliable. This is confirmed by recent comparisons between laboratory and field assessments of
toxic metal impacts on photosynthesis, suggesting
that laboratory experiments may overestimate metal
effects (Macinnis-Ng and Ralph, 2002).
The main difficulty with understanding metal impacts on seagrasses is quantifying the exposure, as
it is transient and highly variable in its distribution
between water, sediment, seagrass and epiphytes. It
is important to establish how much heavy metal is
really available in the water column to be absorbed
by the seagrass (Haynes et al., 2000a). Due to the
complex nature of seawater, many compounds have
P. J. Ralph, D. Tomasko, K. Moore, S. Seddon, and C. M. O. Macinnis-Ng
exposure. These regional differences could be indicative of chronic sub-lethal acclimation, or that
the plants from the southern site were less tolerant.
This suggests that tolerance is not constant within
a species across a wide geographic distance and
that local microenvironment strongly influences the
tolerance.
6. Recovery from Herbicides
Few seagrass studies have seriously considered recovery from chronic exposure to herbicides. Recovery from acute herbicide exposure is a common response for macrophytes (including seagrasses), but
this needs to be considered from a more ecological
perspective. Bowmer (1986) demonstrated that the
PSII inhibitory effect of herbicides (such as diuron
and atrazine) on macrophytes was reversed once
the contaminant was removed, and photosynthesis
rapidly recovered. This aspect of herbicide toxicology is rarely considered in laboratory experiments
and is a critical consideration in the interpretation
of any field based herbicide exposure experiment.
Haynes et al. (2000b) found three species of seagrass
recovered when returned to fresh seawater after 5days exposure to diuron. While Macinnis-Ng and
Ralph (2003a) found a virtual halt in photosynthetic
activity induced by exposure to atrazine, irgarol and
diuron was reversed on return to fresh seawater in
the laboratory and field. The exposure period was
only 10 h and it is not known if the seagrass would
recover after a longer exposure period. Field conditions would typically resemble a single pulse of contaminant, followed by a period of flushing, followed
by a subsequent pulse of contaminant (Solomon
et al., 1996), and experiments are needed which simulate this effect.
Assessing endpoints such as growth may represent a more ecologically relevant outcome, but it
may not be the most sensitive indicator of stress,
as the plant has a longer period of time for phytoremediating processes to detoxify or adapt to the
contaminant. Conversely, a more sensitive endpoint
may illustrate an impact, which could possibly be reversed by returning the plant to un-contaminated water. For example, Z. marina exposed to 5–10 µg L
−1
atrazine for 5 weeks significantly reduced growth
(Schwarzschild et al., 1994), whereas Z. capricorni
after just 10 h exposure showed a significant impact
on PSII photochemical efficiency, although this impact was completely reversible in situ (Macinnis-Ng
and Ralph, 2003a). This clearly demonstrates the
different sensitivities of these two partially related
endpoints.
D. Heavy Metals
1. Sources of Heavy Metals
Heavy metals can enter the marine environment
naturally via weathering and erosion of rocks and
soils (Batley, 1996) or through urban and municipal runoff, storm water, sewage, industrial effluents, mining operations, atmospheric deposition and
agricultural activity (Batley, 1996; Irvine and Birch,
1998; Haynes and Johnson, 2000). In addition, activities such as dredging can release heavy metal-laden
sediment from anoxic deeper layers allowing metals
to be re-mobilized and available to contaminate the
biota (Batley, 1987; Lytle and Lytle, 2001).
2. Processes of Heavy Metal Contamination
Bio-availability of heavy metals is influenced by water and sediment pH, sediment particle size, redox
potential, dissolved organic matter, sediment cation
exchange capacity, water temperature, salinity, organic content, and the concentrations of other metals (Batley, 1987; Ward, 1989; Batley et al., 1999;
Alutoin et al., 2001). These physicochemical parameters influence processes that affect heavy metal
availability, such as chelation, complexing, precipitation, absorption, and adsorption (Batley, 1987).
For this reason, as well as the more complex nature of the field environment, heavy metals are less
bio-available and generally have a less toxic effect
in the field that the laboratory (Graney et al., 1995),
thus extrapolating laboratory findings to the field can
be unreliable. This is confirmed by recent comparisons between laboratory and field assessments of
toxic metal impacts on photosynthesis, suggesting
that laboratory experiments may overestimate metal
effects (Macinnis-Ng and Ralph, 2002).
The main difficulty with understanding metal impacts on seagrasses is quantifying the exposure, as
it is transient and highly variable in its distribution
between water, sediment, seagrass and epiphytes. It
is important to establish how much heavy metal is
really available in the water column to be absorbed
by the seagrass (Haynes et al., 2000a). Due to the
complex nature of seawater, many compounds have